Development in Wastewater Treatment Research and Processes: Microbial Ecology, Diversity and Functions of Ammonia Oxidizing Bacteria 0323919014, 9780323919012

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Development in Wastewater Treatment Research and Processes

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Development in Wastewater Treatment Research and Processes Microbial Ecology, Diversity, and Functions of AmmoniaOxidizing Bacteria Edited by

Maulin P. Shah Industrial Wastewater Research Lab, Division of Applied & Environmental Microbiology, Enviro Technology Limited, Ankleshwar, Gujarat, India

Susana Rodriguez-Couto Department of Separation Science, LUT School of Engineering Science, LUT University, Mikkeli, Finland

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 © 2022 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: 978-0-323-91901-2 For information on all Elsevier publications visit our website at https://www.elsevier.com/books-and-journals

Publisher: Susan Dennis Acquisitions Editor: Anita A Koch Editorial Project Manager: Czarina Mae S. Osuyos Production Project Manager: Kumar Anbazhagan Cover Designer: Mark Rogers Typeset by STRAIVE, India

Contents Contributors ............................................................................................................xix

CHAPTER 1

Anammox process: An innovative approach and a promising technology ................................... 1 Komal Agrawal, Maulin P. Shah, and Pradeep Verma

1.1 1.2 1.3 1.4

Introduction ....................................................................................1 Mechanism of anammox process...................................................2 Role of microorganisms in anammox............................................4 Role of various parameters on anammox ......................................5 1.4.1 Ammonium .......................................................................... 5 1.4.2 Nitrite ................................................................................... 6 1.4.3 Organic matter ..................................................................... 6 1.5 The limitations and solutions of the anammox system...............10 1.6 Conclusion....................................................................................10 Conflict of interest....................................................................... 10 References.................................................................................... 10

CHAPTER 2

Abundance of ammonia-oxidizing bacteria and archaea in industrial wastewater treatment systems............................................................... 17 Vidya Sawant and Hitesh S. Pawar

2.1 Introduction ..................................................................................17 2.2 Key enzymes involved .................................................................19 2.3 Physiology and cellular structure.................................................20 2.3.1 Physiology of AOA ........................................................... 20 2.3.2 Physiology of AOB ........................................................... 21 2.4 Diversity in WWTPs....................................................................21 2.4.1 Diversity of AOA .............................................................. 23 2.4.2 Diversity of AOB .............................................................. 23 2.5 Mechanism of action of AOA and AOB.....................................24 2.5.1 Mechanism of AOA .......................................................... 24 2.5.2 Mechanism of AOB........................................................... 25 2.6 Competition and symbiotic relationships between AOMs..........27 2.7 AOA at low DO or in special WWTPs .......................................27 2.8 Factors influencing AOB abundance and diversity.....................28 2.8.1 Ammonia levels ................................................................. 28 2.8.2 FNA and nitrite.................................................................. 28 2.8.3 Process conditions and regime .......................................... 29

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2.9 Quantification techniques.............................................................29 2.9.1 DNA extraction.................................................................. 29 2.9.2 Quantitative PCR and reverse transcriptional qPCR .................................................................................. 29 2.9.3 High throughput sequencing ............................................. 30 2.9.4 Phylogenetic analysis ........................................................ 30 2.10 Environmental factors affecting AOA and AOB ........................31 2.10.1 Ammonia concentration .................................................. 31 2.10.2 Temperature ..................................................................... 31 2.10.3 Oxygen and aeration pressure ......................................... 32 2.10.4 Organic loading ............................................................... 32 2.10.5 Salinity ............................................................................. 32 2.10.6 DO.................................................................................... 33 2.10.7 Sulfide .............................................................................. 33 2.11 Future perspectives.......................................................................33 2.12 Conclusion....................................................................................34 References.................................................................................... 35

CHAPTER 3

Autotrophic nitrification in bacteria ...................... 41 Moupriya Nag, Dibyajit Lahiri, Sougata Ghosh, Sujay Ghosh, and Rina Rani Ray

3.1 Introduction ..................................................................................41 3.2 Symbiotic nitrogen fixers.............................................................42 3.2.1 Molecular mechanism of endosymbionts.......................... 43 3.2.2 Molecular mechanism of nodule formation...................... 43 3.2.3 Mechanism of exchange of nutrients and nitrogen........... 44 3.3 Events of nitrogen fixation ..........................................................46 3.3.1 Nitrification........................................................................ 46 3.3.2 Nitrate and nitrite synthesis during nitrification............... 47 3.3.3 Hydroxylamine oxidoreductase......................................... 47 3.3.4 Nitrous oxide production during nitrification ................... 47 3.4 Genetic regulation of nitrogen fixation .......................................48 3.5 Understanding the balance between Photosynthesis and nitrogen fixation ....................................................................49 3.5.1 Nitrogen fixation by cyanobacteria................................... 49 3.5.2 Nitrogen fixation by rhizobia ............................................ 50 3.5.3 Role of abiotic factors in BNF.......................................... 53 3.6 Conclusion and future aspect.......................................................54 References.................................................................................... 55

Contents

CHAPTER 4

Omics: A revolutionary tool to study ammonia-oxidizing bacteria and their application in bioremediation ............................... 61 Hiren K. Patel, Priyanka D. Sheladiya, Rishee K. Kalaria, Vivek K. Diyora, and Nidhi P. Patel

4.1 4.2 4.3 4.4 4.5 4.6 4.7 4.8 4.9 4.10 4.11 4.12 4.13 4.14 4.15

CHAPTER 5

Introduction ..................................................................................61 Chemolitho-autotrophic ammonia oxidation ...............................61 Role of ammonia-oxidizing bacteria in nitrogen cycling ...........63 Commercial significance and application of ammonia-oxidizing bacteria ....................................................64 Difficulties associated with nitrification and ammonia-oxidizing bacteria..................................................65 Isolation of ammonia-oxidizing bacteria from the environment............................................................................66 Cultivation of new ammonia oxidizers........................................67 Genomics and metabolic models .................................................68 Terminology of environmental proteomics .................................69 Microbial culture proteomic studies techniques..........................70 Potential applications of environmental proteomics ...................72 Enzymology of ammonia-oxidation ............................................72 Ammonia-oxidizers in the environment and production of N2O........................................................................73 Remediation of recalcitrant pollutants.........................................74 Conclusion....................................................................................75 References.................................................................................... 75

Diversity of ammonia-oxidizing bacteria................ 83 Ambreen Ashar, Muhammad Muneeb, Zeeshan Ahmad Bhutta, and Muhammad Shoaib

5.1 Introduction ..................................................................................83 5.2 Emission of nitrous oxide ............................................................85 5.2.1 Potential sources ................................................................ 86 5.2.2 Yield................................................................................... 86 5.3 Niche differentiation ....................................................................87 5.3.1 Oligotrophy ........................................................................ 87 5.3.2 pH....................................................................................... 88 5.4 Conclusion....................................................................................88 References.................................................................................... 89

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CHAPTER 6

Aerobic and anaerobic ammonia oxidizing bacteria................................................ 93 Ayesha Kanwal, Zeeshan Ahmad Bhutta, Moazam Ali, Ambreen Ashar, and Muhammad Shoaib

6.1 Introduction ..................................................................................93 6.2 Ammonia-oxidizing bacteria........................................................94 6.2.1 Ecology .............................................................................. 94 6.2.2 Environmental regulators of ammonia oxidation ............. 95 6.2.3 Strategic functional, anatomical, and biological differentiations among ammonia oxidizers....................... 97 6.3 Anaerobic ammonium oxidation bacteria....................................98 6.3.1 Ecology .............................................................................. 99 6.3.2 Physiology of anammox bacteria .................................... 100 6.4 Microbial interactions and their contribution to enhanced nitrogen removal....................................................103 6.5 Conclusion..................................................................................104 References.................................................................................. 104

CHAPTER 7

Recent advances in biological nitrogen removal from wastewater: Special focus on reactor configuration and nano-mediated microbial nitro-transformation ............................ 111 Asma Musfira Shabbirahmed, Mohanya Kumaravel, Kanti Kusum Yadav, Satya Sundar Mohanty, and Prathap Somu

7.1 Introduction ................................................................................111 7.2 Chemolithotrophs and their diversity ........................................114 7.2.1 Obligate chemolithotroph bacteria .................................. 115 7.2.2 Facultative chemolithotroph bacteria .............................. 116 7.2.3 Sulfur-oxidizing bacteria ................................................. 116 7.2.4 Ammonium-oxidizing bacteria........................................ 117 7.2.5 Nitrite-oxidizing bacteria................................................. 117 7.2.6 Methane-oxidizing bacteria or methanotrophs ............... 117 7.2.7 Ferrous-oxidizing bacteria............................................... 118 7.2.8 Hydrogen-oxidizing bacteria ........................................... 118 7.3 BNR technologies for wastewater treatment.............................118 7.3.1 Nitrification/denitrification.............................................. 119 7.3.2 Nitritation/denitritation .................................................... 119 7.3.3 Sidestream partial nitritation/anammox .......................... 120 7.3.4 Mainstream partial nitritation/anammox......................... 121

Contents

7.3.5 Nitrogen recovery ............................................................ 122 7.3.6 Phototrophic systems ....................................................... 122 7.3.7 Microbial electrochemical cells ...................................... 123 7.4 Advances in the nitrification process ........................................123 7.4.1 Sequencing batch reactor................................................. 123 7.4.2 Activated sludge models ................................................. 124 7.5 Effect of nanomaterials on microbial nitro-transformation ......125 7.6 Conclusion and future perspective.............................................128 References.................................................................................. 129

CHAPTER 8

Diversity of nitrogen-removing microorganisms ................................................. 133 Oscar Franchi, Javiera Toledo-Alarco´n, Jos e Luis Campos, David Jeison Nun˜ez, Annika Vaksmaa, and Estela Tapia-Venegas

8.1 Introduction ................................................................................133 8.2 Nitrogen removal by microorganisms that use sulfur compounds as electron donor ....................................................134 8.2.1 Autotrophic denitrifying sulfur-oxidizing bacteria............................................................................. 136 8.2.2 Growth conditions of ADSOB ........................................ 136 8.2.3 Metabolic pathways involved in sulfur compound oxidation .......................................................................... 138 8.2.4 Molecular tools for assessing microbial diversity in SDAD processes........................................... 139 8.2.5 Technologies used to carry out the SDAD process to treat wastewaters ............................................ 140 8.2.6 Relevant operating conditions in the SDAD process to treat wastewaters ............................................ 142 8.2.7 Projections of using the SDAD process to remove nitrogen in wastewaters ................................................... 143 8.3 Nitrogen removal by microorganisms that use hydrogen as electron donor: Hydrogenotrophic denitrification ................144 8.3.1 Nitrate removal pathway and hydrogen as electron donor ............................................................. 144 8.3.2 Microorganisms and microbial community involved in the process.................................................... 144 8.3.3 Basis of operational conditions ....................................... 145 8.3.4 Possibilities and available technologies for large-scale application ............................................... 146

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8.4 Nitrogen removal by anaerobic nitrate-dependent methanotrophic microorganisms................................................149 8.4.1 Nitrogen removal pathways and ecosystem distribution of the different types of microorganisms .... 149 8.4.2 Activity and factors affecting the enrichment of these microorganisms................................................................ 151 8.4.3 Molecular tools for assessing microbial diversity .......... 153 8.4.4 Application possibilities in sewage and industrial wastewater treatment plants—Main operating conditions description...................................................... 153 Acknowledgments ..................................................................... 156 References.................................................................................. 156

CHAPTER 9

An overview of the anammox process ................. 165 Yan Guo and Yu-You Li

9.1 Introduction ................................................................................165 9.2 The evolution of anammox reaction stoichiometry ..................166 9.3 The existing problems and countermeasures for anammox process application .....................................................................169 9.3.1 The rapid start-up and recovery of anammox-based process.............................................................................. 169 9.3.2 The retention of anammox sludge in the reactor............ 170 9.3.3 The further improvement of NRE................................... 171 9.4 The status of several main anammox-related processes ...........171 9.4.1 Nitritation process............................................................ 173 9.4.2 Pure anammox process .................................................... 174 9.4.3 PNA process .................................................................... 177 9.4.4 Simultaneous nitrogen removal and phosphorus recovery process .............................................................. 181 9.4.5 Denitratation/anammox process ...................................... 183 9.4.6 DAMO/anammox process ............................................... 186 9.5 Conclusion..................................................................................188 References.................................................................................. 188

CHAPTER 10 Aerobic and anaerobic ammonia-oxidizing bacteria: A resilient challenger or innate collaborator ...................................................... 195 Parool Jain, Raunak Dhanker, Aarushi Bhardwaj, Geetanshi Singhla, Kamakshi Saxena, and Touseef Hussain 10.1 Introduction ................................................................................195

Contents

10.2 Physiology and ecology of ammonia-oxidizing bacteria ..........196 10.2.1 Ecology of ammonia-oxidizing bacteria ....................... 197 10.2.2 Physiology of ammonia-oxidizing bacteria .................. 198 10.2.3 Biodiversity of aerobic and anaerobic oxidizing bacteria .......................................................... 199 10.2.4 Species diversity ............................................................ 200 10.3 Factors affecting aerobic and anaerobic oxidizing bacteria .......................................................................................201 10.3.1 Ammonia levels ............................................................. 202 10.3.2 Organic carbon .............................................................. 203 10.3.3 Temperature ................................................................... 203 10.3.4 Salinity ........................................................................... 204 10.3.5 DO levels ....................................................................... 204 10.3.6 pH................................................................................... 205 10.3.7 Sulfide levels ................................................................. 206 10.3.8 Phosphate ....................................................................... 206 10.4 Role of aerobic and anaerobic ammonia-oxidizing bacteria in wastewater treatment plants.....................................207 10.5 Application of anammox in wastewater treatment....................212 10.5.1 Advantages..................................................................... 212 10.5.2 Disadvantages ................................................................ 213 10.6 Ammonia-oxidizing microorganisms: Key players in the promotion of plant growth...............................................213 10.6.1 Autotrophic nitrification ................................................ 213 10.6.2 Heterotrophic nitrification ............................................. 214 10.6.3 Diversity of ammonia oxidizers .................................... 214 10.7 Mechanism of ammonia oxidation by ammonia-oxidizing microorganisms ..........................................................................214 10.8 Function and activity of ammonia-oxidizing microbes in different soil types .................................................................215 10.8.1 pH................................................................................... 215 10.8.2 Bioavailability of nutrients............................................ 216 10.8.3 Temperature ................................................................... 216 10.8.4 Soil water content .......................................................... 216 10.9 Conclusion..................................................................................216 References.................................................................................. 217

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CHAPTER 11 A technique to boost the nitrogen-rich agricultural ecosystems efficiency by anaerobic ammonium oxidation (anammox) bacteria ............................ 223 Bablesh Ranawat, Freny Shah, Sonam Dubey, Aneesha Singh, and Sandhya Mishra 11.1 11.2 11.3 11.4 11.5

Introduction ................................................................................223 Role of anaerobic ammonium oxidation in nitrogen cycle ......225 Diversity and richness of anammox bacteria ............................226 Uncovering anammox bacteria and its reaction........................227 Role of anammox in agricultural soil........................................228 11.5.1 Anammox in paddy soil ................................................ 228 11.5.2 Anammox in arable high grounded soil........................ 229 11.5.3 Anammox in special sites.............................................. 229 11.6 Factors affecting anammox .......................................................230 11.6.1 Nitrate, nitrite, and ammonium ..................................... 230 11.6.2 Salinity and pH of the soil ............................................ 230 11.6.3 Rhizosphere effect ......................................................... 231 11.7 Outlook for anammox research and concluding remarks .........231 11.8 Future prospects .........................................................................232 Acknowledgments ..................................................................... 232 References.................................................................................. 232

CHAPTER 12 Genomics of ammonia-oxidizing bacteria and denitrification in wastewater treatment plants ................................................ 237 Martha In es V elez-Mercado, Brayan Arturo Pin˜a-Arroyo, Carlos Antonio Espinoza-Lavenant, Aldo Sosa-Herrera, Edgar Ramirez-Ramirez, Aldo Almeida, Miriam Paulina Luevanos-Escaren˜o, Ayerim Yedid Herna´ndez-Almanza, Javier Ulises Herna´ndez-Beltran, Cristo´bal Noe Aguilar-Gonza´lez, and Nagamani Balagurusamy 12.1 Introduction ................................................................................237 12.2 Nitrogen cycle............................................................................238 12.3 Role of ammonia-oxidizing bacteria (AOB) and ammoniaoxidizing archaea (AOA) in nitrogen cycle ..............................240 12.4 Factors that influence AOM abundance and distribution .........241 12.5 Other ammonia-oxidizing microorganisms in wastewater treatment.....................................................................................242

Contents

12.6 Genetic regulation for ammonia oxidation by AOMs ..............242 12.6.1 Genes involved in the oxidation of ammonia to hydroxylamine ............................................................... 242 12.6.2 Genes involved in the oxidation of hydroxylamine to nitric oxide and further to nitrite .............................. 244 12.6.3 Genes involved in the direct oxidation of ammonia to nitrate .................................................... 245 12.7 Gene amoA as a functional marker for AOM ..........................245 12.7.1 Evolutionary relation of AMO and pMMO .................. 245 12.8 Denitrification ............................................................................247 12.8.1 Nitrate reduction to nitrite............................................. 247 12.8.2 Nitrite reduction to nitric oxide..................................... 249 12.8.3 Nitric oxide reduction to nitrous oxide......................... 249 12.8.4 Nitrous oxide reduction to molecular nitrogen............. 249 12.9 Nitrifier denitrification...............................................................250 12.10 Conclusions ................................................................................251 References.................................................................................. 251

CHAPTER 13 Genomic modules of the nitrifying and denitrifying bacterial population in the aerated wastewater treatment systems.............................................. 257 Boobal Rangaswamy, Amirthavarshini Muralidharan, Aishwarya Subramani, Divya Mayilsamy, and Hari Hara Sudhan Palanisamy 13.1 Introduction ................................................................................257 13.2 Microbial association and biofilm formation in the aerated bioreactors .....................................................................258 13.3 Mutualism between the microbial communities .......................259 13.4 Factors influencing the microbial shift .....................................260 13.4.1 Substrate availability ..................................................... 260 13.4.2 Carbon-nitrogen (C/N) ratio .......................................... 261 13.4.3 Dissolved oxygen (DO) concentration.......................... 261 13.4.4 Temperature ................................................................... 262 13.4.5 pH................................................................................... 262 13.5 Population dynamics of the bacterial groups ............................262 13.5.1 Functional plasticity and functional redundancy .......... 262 13.6 Microbial community in the biofilm .........................................263 13.7 Heterotrophic nitrification and aerobic denitrification .............264 13.8 Functional genomics of the microbial community ...................264 13.8.1 Nitrifiers......................................................................... 265

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13.8.2 Comammox.................................................................... 267 13.8.3 Anammox....................................................................... 267 13.8.4 Denitrifiers ..................................................................... 268 13.9 Molecular approaches and bioinformatics tools—Dynamics of the microbial population ..........................268 13.10 Conclusion..................................................................................271 References.................................................................................. 271

CHAPTER 14 Influence of the different operational strategies on anammox processes for the sustainable ammonium wastewater treatment .......................................................... 277 Rahul Jadhav, Chetan Aware, Ranjit Gurav, Yung-Hun Yang, and Jyoti Jadhav 14.1 14.2 14.3 14.4

14.5 14.6 14.7 14.8

Introduction ................................................................................277 Microorganisms involved in the anammox process..................279 Mechanism of anoxic removal of ammonia..............................280 Factors affecting Anammox process and operational strategies.....................................................................................281 14.4.1 Temperature ................................................................... 281 14.4.2 pH................................................................................... 282 14.4.3 Dissolved oxygen (DO) ................................................. 283 14.4.4 Nitrogen loading ............................................................ 284 14.4.5 Carbon sources............................................................... 284 14.4.6 Organic toxicants........................................................... 285 14.4.7 Effect of toxic metals on anammox process................. 287 Recent advancement in anammox process................................288 Diverse applications of anammox process ................................289 Future prospectus of anammox process ....................................290 Conclusion..................................................................................292 Acknowledgments ..................................................................... 292 Competing interests ................................................................... 292 References.................................................................................. 292

CHAPTER 15 Anammox processes in marine environment: Deciphering the roles and applications ............... 297 Jakir Hossain, Md. Foysul Hossain, and Roksana Jahan 15.1 Introduction ................................................................................297 15.2 Overview of the anammox process ...........................................298

Contents

15.3 Anammox bacteria in marine environment...............................300 15.4 Anammox processes in different marine ecosystems ...............302 15.4.1 Marine sediment ............................................................ 302 15.4.2 Oxygen minimum zone ................................................. 303 15.4.3 Marine sponges.............................................................. 304 15.4.4 Arctic sea ice ................................................................. 304 15.5 Role of anammox in marine environment ................................304 15.5.1 Anammox and marine biogeochemical cycles ............................................................................. 304 15.5.2 Marine nitrogen cycling and anammox: A global perspective ..................................................................... 307 15.6 Application of anammox process in marine environment and its potential ....................................................308 15.6.1 Application in marine aquaculture ................................ 308 15.6.2 Application in wastewater treatment............................. 310 15.7 Conclusion..................................................................................312 References.................................................................................. 313

CHAPTER 16 Diversity and versatility of ammonia-oxidizing bacteria ............................................................ 319 G. Anjali and P.C. Sabumon 16.1 Introduction ................................................................................319 16.2 Evolution and classification of ammonium-oxidizing microorganisms (AOMs) ...........................................................320 16.2.1 AOB and AOA .............................................................. 321 16.2.2 Ammonium oxidizer in Comammox ............................ 322 16.2.3 Anaerobic ammonium oxidizer in anammox................ 322 16.2.4 Heterotrophic nitrifying bacteria (HNB) as ammonium oxidizers................................................. 323 16.3 Diversity, specificity, and adaptability of AOB........................324 16.3.1 Diversity of AOB .......................................................... 324 16.3.2 Specificity of AOB ........................................................ 327 16.3.3 Adaptability in cohabitation with other species ........... 330 16.4 Tolerance and inhibition of AOB..............................................332 16.4.1 Ammonia ....................................................................... 332 16.4.2 Carbon............................................................................ 333 16.4.3 Other inhibitory substances ........................................... 334 16.5 Recent applications and challenges of AOB.............................335 16.5.1 Novel and Hybrid reactors involving AOBs................. 335 16.5.2 Challenges on employing AOB .................................... 338

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16.6 Future research prospects employing the versatile ammonium oxidizers...................................................339 16.7 Conclusions ................................................................................339 References.................................................................................. 340

CHAPTER 17 Role of ammonia oxidizers in performing simultaneous nitrification and denitrification process in advanced SBR plants......................... 347 Ghazal Srivastava and Absar Ahmad Kazmi 17.1 Introduction ................................................................................347 17.2 Theory of SND and practical applications in different WWTPs/technologies .............................................351 17.3 Advantages of SND over anammox and other biological nitrogen removal processes ......................................355 17.3.1 Nitrification–denitrification........................................... 356 17.3.2 Nitritation–Denitritation ................................................ 356 17.3.3 Simultaneous nitrification and denitrification (SND) (or aerobic denitrification)................................. 356 17.3.4 Nitritation–ANAMMOX ............................................... 359 17.3.5 CANON process ............................................................ 359 17.4 Types and characteristics of different ammonia oxidizers and nitrate reducers encouraging SND mechanism prevailing in these systems ...........................361 17.5 Operational parameters/factors that control the diversity of nitrifiers (ammonia and nitrite oxidizers) and denitrifiers (nitrate reducers) during the SND mechanism in advanced SBR plants ..............362 17.5.1 Carbon source: Readily biodegradable COD, soluble COD, soluble BOD5 ............................... 362 17.5.2 C/N ratio ........................................................................ 362 17.5.3 Floc size and PHB storage ............................................ 363 17.5.4 Dissolved oxygen control .............................................. 363 17.5.5 ORP................................................................................ 364 17.5.6 pH................................................................................... 364 17.5.7 Temperature ................................................................... 364 17.5.8 HRT and SRT ................................................................ 364 17.6 Effect of free ammonia (FA), nitrate concentrations, and some metals on AOBs ........................................................365 17.7 Conclusion..................................................................................366 References.................................................................................. 366

Contents

CHAPTER 18 Diversity and functional role of ammonia-oxidizing bacteria in soil microcosms....................................................... 371 Jintu Rabha, Sashi Prava Devi, Sukanya Das, Amrit Kumar, and Dhruva Kumar Jha 18.1 Introduction ................................................................................371 18.2 Diversity and distribution of ammonia-oxidizing bacteria in soil............................................................................372 18.2.1 Diversity of ammonia-oxidizing bacteria in different ecological niches ........................................ 375 18.2.2 Determination of soil AOB diversity ............................ 377 18.3 Factors affecting ammonia oxidation in soil.............................378 18.3.1 Ammonia concentration ................................................ 379 18.3.2 Soil pH ........................................................................... 379 18.3.3 Temperature ................................................................... 380 18.3.4 Moisture content ............................................................ 380 18.3.5 Fertilizers and manures ................................................. 380 18.3.6 Contaminants ................................................................. 381 18.3.7 Salinity ........................................................................... 382 18.4 Molecular biology of ammonia oxidation in bacteria ..............382 18.5 Economic importance of AOBs ................................................385 18.6 Conclusion and prospects ..........................................................386 References.................................................................................. 387

CHAPTER 19 Anaerobic ammonia oxidation: From key physiology to full-scale applications................... 393 Sumira Malik, Shristi Kishore, Shradha A. Kumar, and Vinay Kumar 19.1 Introduction ................................................................................393 19.2 Anammox bacteria: Diversity and cell biology ........................394 19.3 Physiological parameters and the metabolic pathway involved in anammox..................................................397 19.4 Possible reaction mechanism for the anammox process and the factors influencing the reaction.......................399 19.5 Anammox culture in the laboratory ..........................................402 19.6 Full-scale applications of the anammox process ......................405 19.7 Conclusions ................................................................................408 References.................................................................................. 409

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CHAPTER 20 Ammonification in the oral microbiome with plausible link to diet and health and their systemic role in the salivary entero-nitrate channel—A reality or farce ................................ 415 Jesse Joel Thathapudi, R.S. David Paul Raj, Gomez Levin Anbu, Ritu Shepherd, Prathap Somu, and John Jobin 20.1 Introduction ................................................................................415 20.2 Ammonification and chemolithotrophs .....................................418 20.2.1 Gluconeogenesis ............................................................ 418 20.2.2 Role of ammonia-oxidizing bacteria............................. 418 20.2.3 Ammonia flux................................................................ 419 20.3 Oral microbiome ........................................................................420 20.3.1 The oral microbiota ....................................................... 420 20.3.2 Streptococcus mutans group.......................................... 421 20.3.3 Role of enzymes ............................................................ 421 20.3.4 Halitosis ......................................................................... 422 20.3.5 Impact on daily life ....................................................... 422 20.4 Plausible link to diet and health ................................................423 20.5 Contemporary scenario and future perception ..........................423 20.5.1 Quorum sensing (QS) .................................................... 423 20.5.2 Inhibition mechanism .................................................... 424 20.6 Conclusion..................................................................................424 References.................................................................................. 425

CHAPTER 21 Nitritation kinetics and its application in wastewater treatment..................................... 429 Ying Song, Yan Guo, and Yu-You Li 21.1 Introduction ................................................................................429 21.2 Factors affecting kinetics of ammonia oxidation microorganisms and nitritation performance ............................430 21.2.1 Aerobic ammonia-oxidizing microorganisms ............... 431 21.2.2 Temperature ................................................................... 431 21.2.3 Free ammonia, and free nitrous acids and pH .............. 435 21.2.4 Aeration control............................................................. 436 21.2.5 DO concentration........................................................... 437 21.3 Unit processes of nitritation ......................................................438 21.3.1 Suspended growth systems............................................ 439 21.3.2 Attached growth systems............................................... 440 21.3.3 Hybrid systems .............................................................. 444 21.4 Conclusions and perspectives ....................................................445 References.................................................................................. 445 Index ......................................................................................................................451

Contributors Komal Agrawal Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India Cristo´bal No e Aguilar-Gonza´lez Directorate of Research & Postgraduate Studies, Autonomous University of Coahuila, Saltillo, Coahuila State, Mexico Moazam Ali Department of Clinical Medicine and Surgery, University of Agriculture, Faisalabad, Pakistan Aldo Almeida Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico; Department of Plant and Environmental Science, University of Copenhagen, Frederiksberg C, Denmark Gomez Levin Anbu Department of Biotechnology, School of Agriculture and Biosciences, Karunya Institute of Technology and Sciences (Deemed to be University), Coimbatore, Tamil Nadu, India G. Anjali School of Civil Engineering, Vellore Institute of Technology (VIT), Chennai, Tamil Nadu, India Ambreen Ashar Department of Chemistry, University of Agriculture, Faisalabad, Pakistan Chetan Aware Department of Biotechnology, Shivaji University, Kolhapur, Maharashtra, India Nagamani Balagurusamy Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Aarushi Bhardwaj Department of Basic and Applied Sciences, School of Engineering and Sciences, GD Goenka University, Gurugram, Haryana, India Zeeshan Ahmad Bhutta Laboratory of Biochemistry and Immunology, College of Veterinary Medicine, Chungbuk National University, Cheongju, Chungbuk, Republic of Korea Jos e Luis Campos Facultad de Ingenierı´a y Ciencias, Universidad Adolfo Iban˜ez, Vin˜a del Mar, Chile

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Contributors

Sukanya Das Microbial Ecology Laboratory, Department of Botany, Gauhati University, Guwahati, Assam, India Sashi Prava Devi Microbial Ecology Laboratory, Department of Botany, Gauhati University, Guwahati, Assam, India Raunak Dhanker Department of Basic and Applied Sciences, School of Engineering and Sciences, GD Goenka University, Gurugram, Haryana, India Vivek K. Diyora School of Science, P. P. Savani University, Surat, Gujarat, India Sonam Dubey Applied Phycology and Biotechnology Division, Central Salt and Marine Chemicals Research Institute, Council for Scientific and Industrial Research (CSIR), G.B. Marg, Bhavnagar, Gujarat, India Carlos Antonio Espinoza-Lavenant Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Oscar Franchi Facultad de Ingenierı´a y Ciencias, Universidad Adolfo Iban˜ez, Vin˜a del Mar, Chile Sougata Ghosh Department of Microbiology, School of Science, RK University, Rajkot, Gujarat, India Sujay Ghosh AMH Energy Pvt. Ltd., Kolkata, West Bengal, India Yan Guo Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Miyagi, Japan Ranjit Gurav Department of Biological Engineering, College of Engineering, Konkuk University, Seoul, South Korea Ayerim Yedid Herna´ndez-Almanza Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Javier Ulises Herna´ndez-Beltran Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Jakir Hossain Department of Marine Fisheries and Oceanography, Sher-e-Bangla Agricultural University, Dhaka, Bangladesh

Contributors

Md. Foysul Hossain Department of Aquatic Environment and Resource Management, Sher-e-Bangla Agricultural University, Dhaka, Bangladesh Touseef Hussain Department of Botany, Aligarh Muslim University, Aligarh, Uttar Pradesh, India Jyoti Jadhav Department of Biotechnology; Department of Biochemistry, Shivaji University, Kolhapur, Maharashtra, India Rahul Jadhav Department of Biotechnology, Shivaji University, Kolhapur, Maharashtra, India Roksana Jahan Department of Marine Fisheries and Oceanography, Sher-e-Bangla Agricultural University, Dhaka, Bangladesh Parool Jain SRM Institute of Science and Technology, Ghaziabad, Uttar Pradesh, India David Jeison Nun˜ez Escuela de Ingenierı´a BioquI´mica, Pontificia Universidad Cato´lica de Valparaı´so, Valparaı´so, Chile Dhruva Kumar Jha Microbial Ecology Laboratory, Department of Botany, Gauhati University, Guwahati, Assam, India John Jobin Department of Biotechnology, School of Agriculture and Biosciences, Karunya Institute of Technology and Sciences (Deemed to be University), Coimbatore, Tamil Nadu, India Rishee K. Kalaria Aspee Shakilam Biotechnology Institute, Navsari Agricultural University, Surat, Gujarat, India Ayesha Kanwal Institute of Biochemistry, Biotechnology and Bioinformatics, The Islamia University of Bahawalpur, Bahawalpur, Pakistan Absar Ahmad Kazmi Department of Civil Engineering, Environmental Engineering Group, Indian Institute of Technology Roorkee, Roorkee, Uttarakhand, India Shristi Kishore Amity Institute of Biotechnology, Amity University Jharkhand, Ranchi, Jharkhand, India Amrit Kumar Microbial Ecology Laboratory, Department of Botany, Gauhati University, Guwahati, Assam, India

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Contributors

Shradha A. Kumar Amity Institute of Biotechnology, Amity University Jharkhand, Ranchi, Jharkhand, India Vinay Kumar Modern College of Arts, Science and Commerce, Pune, India Mohanya Kumaravel Department of Biotechnology, School of Agriculture and Bioscience, Karunya Institute of Technology and Sciences (Deemed-to-be University), Coimbatore, Tamil Nadu, India Dibyajit Lahiri Department of Biotechnology, University of Engineering & Management, Kolkata, West Bengal, India Yu-You Li Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Miyagi, Japan Miriam Paulina Lu evanos-Escaren˜o Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Sumira Malik Amity Institute of Biotechnology, Amity University Jharkhand, Ranchi, Jharkhand, India Divya Mayilsamy Department of Biotechnology, PSG College of Arts & Science, Coimbatore, Tamil Nadu, India Sandhya Mishra Applied Phycology and Biotechnology Division, Central Salt and Marine Chemicals Research Institute, Council for Scientific and Industrial Research (CSIR), G.B. Marg, Bhavnagar, Gujarat; Academy of Scientific and Innovative Research (AcSIR), CSIR-HRDC Campus, Ghaziabad, Uttar Pradesh, India Satya Sundar Mohanty Department of Biotechnology, School of Agriculture and Bioscience, Karunya Institute of Technology and Sciences (Deemed-to-be University), Coimbatore, Tamil Nadu, India Muhammad Muneeb Department of Pathology, University of Agriculture, Faisalabad, Pakistan Amirthavarshini Muralidharan Department of Biotechnology, PSG College of Arts & Science, Coimbatore, Tamil Nadu, India Moupriya Nag Department of Biotechnology, University of Engineering & Management, Kolkata, West Bengal, India

Contributors

Hari Hara Sudhan Palanisamy Department of Biotechnology, PSG College of Arts & Science, Coimbatore, Tamil Nadu, India Hiren K. Patel School of Science, P. P. Savani University, Surat, Gujarat, India Nidhi P. Patel School of Science, P. P. Savani University, Surat, Gujarat, India Hitesh S. Pawar DBT-ICT Centre for Energy Biosciences, Institute of Chemical Technology, Mumbai, India Brayan Arturo Pin˜a-Arroyo Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Jintu Rabha Microbial Ecology Laboratory, Department of Botany, Gauhati University, Guwahati, Assam, India R.S. David Paul Raj Department of Biotechnology, School of Agriculture and Biosciences, Karunya Institute of Technology and Sciences (Deemed to be University), Coimbatore, Tamil Nadu, India Edgar Ramirez-Ramirez Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Bablesh Ranawat Applied Phycology and Biotechnology Division, Central Salt and Marine Chemicals Research Institute, Council for Scientific and Industrial Research (CSIR), G.B. Marg, Bhavnagar, Gujarat; Academy of Scientific and Innovative Research (AcSIR), CSIR-HRDC Campus, Ghaziabad, Uttar Pradesh, India Boobal Rangaswamy Department of Biotechnology, PSG College of Arts & Science, Coimbatore, Tamil Nadu, India Rina Rani Ray Department of Biotechnology, Maulana Abul Kalam Azad University of Technology, Haringhata, West Bengal, India P.C. Sabumon School of Civil Engineering, Vellore Institute of Technology (VIT), Chennai, Tamil Nadu, India Vidya Sawant DBT-ICT Centre for Energy Biosciences, Institute of Chemical Technology, Mumbai, India

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Contributors

Kamakshi Saxena SRM Institute of Science and Technology, Ghaziabad, Uttar Pradesh, India Asma Musfira Shabbirahmed Department of Biotechnology, School of Agriculture and Bioscience, Karunya Institute of Technology and Sciences (Deemed-to-be University), Coimbatore, Tamil Nadu, India Freny Shah Applied Phycology and Biotechnology Division, Central Salt and Marine Chemicals Research Institute, Council for Scientific and Industrial Research (CSIR), G.B. Marg, Bhavnagar, Gujarat, India Maulin P. Shah Industrial Wastewater Research Lab, Division of Applied & Environmental Microbiology, Enviro Technology Limited, Ankleshwar, Gujarat, India Priyanka D. Sheladiya School of Science, P. P. Savani University, Surat, Gujarat, India Ritu Shepherd School of Arts and Creative Sciences, Nehru Arts and Science College, Coimbatore, Tamil Nadu, India Muhammad Shoaib Institute of Microbiology, University of Agriculture, Faisalabad, Pakistan Aneesha Singh Applied Phycology and Biotechnology Division, Central Salt and Marine Chemicals Research Institute, Council for Scientific and Industrial Research (CSIR), G.B. Marg, Bhavnagar, Gujarat; Academy of Scientific and Innovative Research (AcSIR), CSIR-HRDC Campus, Ghaziabad, Uttar Pradesh, India Geetanshi Singhla Department of Basic and Applied Sciences, School of Engineering and Sciences, GD Goenka University, Gurugram, Haryana, India Prathap Somu Department of Biotechnology, Saveetha School of Engineering, Saveetha Institute of Medical and Technical Sciences (Deemed to be University), Chennai, Tamil Nadu, India Ying Song Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Miyagi, Japan Aldo Sosa-Herrera Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico

Contributors

Ghazal Srivastava Department of Civil Engineering, Environmental Engineering Group, Indian Institute of Technology Roorkee, Roorkee, Uttarakhand, India Aishwarya Subramani Department of Biotechnology, PSG College of Arts & Science, Coimbatore, Tamil Nadu, India Estela Tapia-Venegas Departamento de Ciencias de la Ingenierı´a para la sostenibilidad, Facultad de Ingenierı´a, Universidad de Playa Ancha, Valparaı´so, Chile Jesse Joel Thathapudi Department of Biotechnology, School of Agriculture and Biosciences, Karunya Institute of Technology and Sciences (Deemed to be University), Coimbatore, Tamil Nadu, India Javiera Toledo-Alarco´n Escuela de Ingenierı´a BioquI´mica, Pontificia Universidad Cato´lica de Valparaı´so, Valparaı´so, Chile Annika Vaksmaa Department of Marine Microbiology and Biochemistry, Royal Netherlands Institute for Sea Research, Texel, The Netherlands Martha In es V elez-Mercado Bioremediation Laboratory, Faculty of Biological Sciences, Autonomous University of Coahuila, Torreon, Coahuila State, Mexico Pradeep Verma Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India Kanti Kusum Yadav Department of Biotechnology, School of Agriculture and Bioscience, Karunya Institute of Technology and Sciences (Deemed-to-be University), Coimbatore, Tamil Nadu, India Yung-Hun Yang Department of Biological Engineering, College of Engineering, Konkuk University, Seoul, South Korea

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CHAPTER

Anammox process: An innovative approach and a promising technology

1

Komal Agrawala, Maulin P. Shahb, and Pradeep Vermaa Bioprocess and Bioenergy Laboratory, Department of Microbiology, Central University of Rajasthan, Ajmer, Rajasthan, India bIndustrial Wastewater Research Lab, Division of Applied & Environmental Microbiology, Enviro Technology Limited, Ankleshwar, Gujarat, India a

1.1 Introduction Anammox also referred to as the “Anaerobic Ammonium Oxidation” is a process where the ammonium is transformed into N2 (dinitrogen) with nitrite as the electron (e ) acceptor. The anammox is a chemoautotrophic biological process and relies on bacteria that belong to the phylum Planctomycetes, e.g., Brocadia (Puyol et al., 2013), Kuenenia (Speth et al., 2012), Jettenia (Hu et al., 2012), Scalindua, etc. (Woebken et al., 2008; Sinninghe Damst
e et al., 2005; Van de Graaf et al., 1995). In the past decade, anammox has received immense attention from the scientific community for the treatment of WW (Cui et al., 2020; Kokabian et al., 2018). The establishment of full-scale anammox sewage treatment plants did see a rise and a total of 114 plants were established by the year 2015 globally (Ali and Okabe, 2015; Ali et al., 2014). The process has gained more acceptance for commercial use because of the cost-effectiveness as aeration and supply of exogenous carbon source is not required during the functioning of the process and has high efficiency of the process. The anammox bacteria are very easy to find in both natural and artificial systems such as marine ecosystems (Nakajima et al., 2008), wastewater (Wang et al., 2011), freshwater (Ding et al., 2020), natural ecosystems (Hu et al., 2011), terrestrial ecosystems (Humbert et al., 2010), wetland soils (Humbert et al., 2012). However, the anammox system has restricted use due to various parameters such as inhibition factors, e.g., free ammonia, nitrite, and organic matter along with physiological parameters such as pH and temperature. These can be overcome by optimization but then commercialization would be the biggest barrier in the practical implementation of the process. Thus the present chapter would discuss the working principle, the relationship between anammox bacteria, and the parameters in the anammox process. Further, the limitation and solutions for effective implementation will be elaborated. Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00002-4 Copyright # 2022 Elsevier Inc. All rights reserved.

1

2

CHAPTER 1 Anammox process

1.2 Mechanism of anammox process The anammox reaction in the case of bacteria occurs within the anammoxosome that is the intracytoplasmic compartment. The catabolic reaction creates a proton gradient across the anammoxosome membrane (van Niftrik et al., 2004). The anammox reaction consists of series of steps of the reaction. In the first step, the nitrite is reduced to nitric oxide in the presence of nitrate reductase. In the presence of hydrazine hydrolase, ammonia combines with nitric oxide and forms hydrazine which is finally converted to dinitrogen gas via oxidation reaction in the presence of hydrazine dehydrogenase. The reaction is represented in Fig. 1.1. The anammox process has been used for the treatment of contaminated WW and is more feasible at commercial or scale-up processes due to its low cost and high efficiency (Table 1.1). Of the various anammox process, the one-stage system is the most commonly studied (Li et al., 2018; Connan et al., 2018; Shah, 2021). Despite the advantages of the anammox process, the main hindrance is the doubling time of the bacteria used in the system. In the study by Okabe et al. (2021) the doubling time for “Candidatus Brocadia sinica” and “Candidatus Scalindua sp.” was 1255 and 184 days, respectively. As a result, over time various methodologies and techniques have been adapted for the enhancement of the anammox process such as effects of aerobic and microaerobic conditions (Strous et al., 1997), upflow anammox sludge bed reactor (Reino et al., 2018), enrichment of anammox (Third et al., 2005), immobilization (Magrı´ et al., 2012), etc. Also in addition to various parameters, the inoculation of the system with partially or fully enriched anammox sludge, the reactors can enhance the performance of the system in some weeks (Li et al., 2018; Ali et al., 2015; Guo et al., 2010; Yu et al., 2013). However, a more detailed study is required in this area of research for better understanding and implementation of the system.

FIG. 1.1 Schematic representation of anammox reaction.

1.2 Mechanism of anammox process

Table 1.1 Various anammox systems reported in the literature and their application. Sl. no.

Anammox system

Application

Reference

1.

Two reactors, i.e., Reactor 1 (containing biochar and anammox bacteria) and Reactor 2 (containing biochar as a control) were studied for 4 months Rapid biofiltering anammox reactor and the mode of operation continuous was upward flowing along with gradually shortened hydraulic retention time

Nitrogen removal was better in Reactor 1 over Reactor 2

Mojiri et al. (2020)

The used operational conditions where carmine anammox granular sludge and thick biofilm coexisted The strain Candidatus Brocadia was dominant The total nitrogen removal was stable (86.2%) under a nitrogen loading rate of 4.86 kg N/m3 day and HRT of 32 min. The system contributed toward N loss in both aerobic and anoxic tanks Mass balance exhibited that anammox rate was higher than denitrification. Granular sludge had a high purity Nitrate-to-nitrite transformation and P removal were attained by optimizing anaerobic/anoxic/ aerobic durations and influent nitrate concentration with suitable NO2 -N/NH4+-N The nitrogen removal and mechanisms of quorum sensing or quorum quenching were determined The nitrogen removal was high in Anammox H The strain Candidatus Kuenenia was dominant in both the reactors Dense biofilm development was observed in Anammox H More active quorum sensing was in Anammox H The influent concentrations of ammonium affected the quorum-sensing and quorumquenching activities

Lu et al. (2020)

2.

3.

Developed spontaneous anammox process in a fullscale swine wastewater system

4.

An integration of anammox and endogenous partial denitrification and phosphorus removal in two-stage sequencing batch reactors

5.

Two lab-scale Anammox biofilm reactors fed with influent ammonium concentrations of two concentrations/load, i.e., 110 mg L 1 (Anammox-H) and 50 mg L 1 (Anammox-L) were operated

Wang et al. (2019a)

Wang et al. (2019b)

Sun et al. (2018)

Continued

3

4

CHAPTER 1 Anammox process

Table 1.1 Various anammox systems reported in the literature and their application—cont’d Sl. no.

Anammox system

Application

Reference

6.

An anaerobic fixed bed reactor was A-stage, the anammox moving bed biofilm reactor was B2-stage, and the sequencing batch reactor was B1-stage. Effect of nanoscale Zero-Valent Iron and hydraulic shock during high-rate anammox WW treatment The use of anammox bacteria as biological cathodes/ catalysts in microbial desalination cell

COD removal, total inorganic nitrogen (TIN) removal, and sludge reduction

Gu et al. (2018)

Treatment of wastewater

Xu et al. (2018)

Simultaneous generation of electricity, desalinate saltwater, and remove nitrogenous compounds in the cathode chamber of microbial desalination cell using wastewater Removal of chemical oxygen demand (COD)

Kokabian et al. (2018)

The nitrogen loading rates of the reactor increased as well as total nitrogen removal efficiencies

Qiao et al. (2008)

7.

8.

9.

10.

Anammox nitrification was studied in an anoxic/aerobic granular sludge reactor for 390 days at 18°C  3 One column-type reactor with net type acrylic fiber (Biofix) as support material was used for anammox treatment at 25°C and 330 days

Winkler et al. (2012)

1.3 Role of microorganisms in anammox In the case of anammox bacteria two major components are required for the functioning, i.e., ammonia and nitrite. The ammonia is attained through the wastewater and nitrite via the metabolism of the microorganisms in the system. For the production of nitrite, ammonia is consumed by ammonia monooxygenase during nitrification. The nitrite so produced can be easily converted to nitrate by hydroxylamine oxidoreductase in the presence of O2 by nitrite-oxidizing bacteria. Also, the heterotrophic bacteria have the potential to produce nitrite as an intermediate product in the presence of nitrate reductase, nitric oxide reductase, and nitrous oxide reductase constituting the entire denitrification process. However, here too the heterotrophic bacteria may possibly convert the nitrite to dinitrogen, thus making it clear that the microbial community operating in the system has a very complex system and competition (Li et al., 2018) (Fig. 1.2).

1.4 Role of various parameters on anammox

FIG. 1.2 Schematic representation of the role of microorganisms in the anammox process.

1.4 Role of various parameters on anammox The various parameters impact the functioning of the anammox system and are as follows:

1.4.1 Ammonium The ammonia in the anammox process has been provided by wastewater (Van Hulle et al., 2010) and it can tolerate concentrations up to 1 g N L 1. However, high levels suppress the process (Strous et al., 1999a,b; Ferna´ndez et al., 2012; Dapena-Mora et al., 2007). Dapena-Mora et al. (2007) stated that an ammonium concentration of 700 mg L 1 was the hemiinhibitory concentration (IC50) where 50% activity loss is attained. Also, the major inhibitor was free ammonia and not ammonium. Thus the main basis of ammonium inhibition is free ammonia and not ammonium. The work by Waki et al. (2007) stated that free ammonia at 13–90 mg L 1 exhibited toxicity toward anammox organisms, and later Ferna´ndez et al. (2012) reported that the optimal free ammonia concentration for the anammox process was 20–25 mg L 1. Further, it has been stated that the anammox microorganisms can tolerate high concentration of free ammonium rather than free ammonia (Leita˜o et al., 2006) and could be due to variation in operating conditions, flocculent, sludge, and the microorganism operating in the system ( Jin et al., 2012).

5

6

CHAPTER 1 Anammox process

1.4.2 Nitrite The nitrite as one of the substrates of anammox beyond a certain threshold level is toxic to the microorganisms (Zhou et al., 2011) and the level is lower than ammonium (Isaka et al., 2007; Dapena-Mora et al., 2007; Van Hulle et al., 2010; Jaroszynski et al., 2011; Bettazzi et al., 2010; Kimura et al., 2010). The threshold concentration of nitrite has been reported to vary from 5 to 280 mg N L 1 depending upon the operating and experimental conditions ( Jaroszynski et al., 2011; Isaka et al., 2007; Shah, 2020). However, in the study by Dapena-Mora et al. (2007) the concentration of nitrite at 350 mg L 1 the anammox activity has only decreased by 50% as this can be inferred that variation in threshold concentration for nitrite as in for ammonia can be due to variation in operating conditions, flocculent, sludge, and the microorganism operating in the system ( Jin et al., 2012). Further, in order to determine the effect of nitrite toxicity, the anammox bacteria was entrapped in gel and immersed in a solution with nitrite concentration of 400 mg L 1 where the activity was not inhibited whereas at 750 mg L 1 the anammox activity depleted by 90% gradually over a period of 7 days (Kimura et al., 2010). In general, the nitrite inhibition of anammox is irreversible ( Jetten et al., 1998; Egli et al., 2001). The impact of nitrite can be reduced by immersing the anammox bacteria in a gel. It can help in the enhancement of mass transfer resistance and decrease nitrogen removal efficiency; however, the concentration of nitrite is a limiting factor and has to be experimentally studied further to overcome the drawback in the system (Isaka et al., 2007; Tsushima et al., 2007).

1.4.3 Organic matter The anammox bacteria use carbon dioxide (CO2) as the carbon source and thus the bicarbonate is a necessity for the functioning of the system (Strous et al., 1999a,b; Dexiang et al., 2008). The organic matter at low concentration does not affect the anammox process and promoted biological reactions but high concentrations have an inhibitory effect (Tang et al., 2010a,b; Dapena-Mora et al., 2007). The organic matter has an inhibitory effect on the anammox system and after the study, two mechanisms were proposed. In the first “out competition” the heterotrophic bacteria outcompete the autotrophic anammox bacteria that is present in the anammox system as the growth of the prior of is faster. This leads to the removal of the autotrophic anammox bacteria from the system and eventually reducing the capability of nitrogen removal (Chamchoi et al., 2008; G€uven et al., 2005; Molinuevo et al., 2009; Lackner et al., 2008). On the second the anammox bacteria predominate but rather than utilizing the nitrite/ammonium it utilizes organic matter (G€uven et al., 2005; Kartal et al., 2008) and reduces nitrogen removal potential. However, an advantage here is that the nitrogen removal process is an independent process and can be expressed again when the organic matter is completed from the system and this is referred to as “metabolic pathway conversion inhibition.” The other group of organic matter such as alcohol, aldehydes, antibiotics, and phenol also exist that have a toxic and mostly irreversible impact on the anammox

1.4 Role of various parameters on anammox

process via enzyme inactivation or poisoning of the microorganisms in the anammox system (G€ uven et al., 2005; Jin et al., 2012). The methanol concentration of 3–4 mmol L 1 completely inhibited the anammox bacteria in the natural system of marine sediment (G€uven et al., 2005; Jensen et al., 2007). However, under artificial conditions the anammox process was inhibited but at varying concentrations/values and could be due to varying conditions and species ( Jin et al., 2012). In the study by Isaka et al. (2008) it was stated that if the anammox activity was inactive the methanol inhibition was also not observed. In the case of methanol inactivation, the methanol is intracellularly converted to formaldehyde via hydroxylamine oxidoreductase—an anammox enzyme (Isaka et al., 2008; Schalk et al., 2000). The formaldehyde so formed inactivates the enzyme and irreversibly cross-links the peptide chains, thereby irreversibly inhibiting the anammox system (G€ uven et al., 2005; Metz et al., 2004; Isaka et al., 2008). Though a more detailed study is required for a better and detailed understanding of the anammox process. The other inhibitor phenol is found in industrial effluent (e.g., pulp, textile, petrochemical, pharmaceutical, etc.) (Collins et al., 2005; Dosta et al., 2011). A study was conducted where strain Anammox Candidatus Brocadia anammoxidans was acclimated in phenol-containing wastewater. As the concentration increased from 50 (10) to 550 (10) mg L 1 initially the nitrogen removal decreased after which it gradually acclimatized (Toh and Ashbolt, 2002; Toh et al., 2002). Thus it can be stated that though phenol limited the anammox process it gradually recovered and the process enhanced after acclimatization ( Jin et al., 2012). The other toxic inhibitor is the antibiotic (Tang et al., 2011; Van de Graaf et al., 1995; Fernandez et al., 2009) and high salinity where at high salt concentrations the cell would die due to osmotic pressure or plasmolyze and become dormant ( Jin et al., 2007) (Table 1.2).

Table 1.2 Various optimization studies reported in the literature to enhance the functioning of the anammox system. Sl. no. 1.

System Partial nitritation/ Anammox process in sequencing batch reactor

Type of wastewater Ammonium wastewater

Parameter

Reference

The feed distribution in the sequencing batch reactor cycle and subcycles was optimized using a response surface methodology (RSM) The nitrogen removal efficiency and rate of 88% and 0.84 kg/m3/ day were attained

Choi et al. (2019)

Continued

7

8

CHAPTER 1 Anammox process

Table 1.2 Various optimization studies reported in the literature to enhance the functioning of the anammox system—cont’d Sl. no.

System

Type of wastewater

2.

Anammox baffled reactor

Partially nitrified ammonia-rich wastewater

3.

Pilot plant

Municipal wastewater

4.

Anammox Sequence Batch Reactor

Mature landfill leachate

Parameter

Reference

AnBR was tested under different nitrogen loading rate values The maximum nitrogen removal rate of 3.39  0.16 at a nitrogen loading rate of 4.04  0.10 kg N/m3/ day was attained A 1-year pilot study of a nitritationdenitritation process followed by anammox polishing An intermittent aeration pattern was controlled based on effluent ammonia and nitrate + nitrite concentration Total inorganic nitrogen removal was 75  15% at a chemical oxygen demand/NH4+-N ratio of 8.9  1.8 within the hydraulic retention time range of 3.1–9.4 h. The anammox polishing further increased the total inorganic nitrogen removal by 11% The factors were feeding mode and influent organics The influent ammonia (250  20 mg L 1) and nitrite concentrations (320  20 mg L 1) at continuous feeding mode reduced the

Ismail et al. (2019)

Regmi et al. (2015)

Miao et al. (2015)

1.4 Role of various parameters on anammox

Table 1.2 Various optimization studies reported in the literature to enhance the functioning of the anammox system—cont’d Sl. no.

5.

System

Laboratoryscale sequencing batch reactor

Type of wastewater

The anammoxenriched granular sludge was obtained from nitritation/ anammox system

Parameter adverse effects of high influent nitrite concentration A small amount of organics improved nitrogen removal With the increase in effluent COD (from 600  50 to 800  50 mg L 1), the anammox was slowly inhibited. Influent organics concentration below 800 mg L 1 facilitated Anammox. Effect of dissolved oxygen (O2) and pH was determined for the nitrogen removal The highest specific nitrogen removal rate of 1.1 gN gVSS 1 day 1 was in nonaerated conditions, with a nitrogen removal efficiency of 81.6%. The nitrogen removal was inhibited under aerated conditions, an increased specific nitrogen removal rate occurred at the dissolved O2 concentration of 0.5 mg O2 L 1. The highest specific nitrogen removal rate was at a pH range of 6.5–8.5 with low concentrations of free ammonia and free nitrous acid

Reference

Yin et al. (2016)

9

10

CHAPTER 1 Anammox process

1.5 The limitations and solutions of the anammox system The “solution” to any inhibitory effect on the system is its “reduction” and “enhancement” of nitrogen removal. The first would be the pretreatment of the wastewater to remove the inhibitors and the second would be optimizing the operating parameters such as pH, controlling substrate concentration, temperature, dissolved oxygen, oxidation-reduction potential, and sludge control. The optimization of these parameters by varying its range can significantly impact the nitrogen removal potential and functioning of the anammox system. Further, the role of various inhibitors has to be studied in detail as they have their respective roles in the functioning of the anammox, e.g., antibiotics. As currently there is limited study on the role of antibiotics and organic compounds, the detailed analysis would pave new ways in the functioning of the anammox system and should primarily focus on the in-depth analysis, removal of inhibitors, molecular level study, enrichment, and cultivation of the dominant microorganism as per the suitability of the system, combination/interdisciplinary approach, and lastly scale-up of the system.

1.6 Conclusion The anammox system is fairly new and has tremendous scope in the future. The various parameters along with the microbes have a significant role to play in the functioning of the anammox system. The detailed study along with the integration of the technological advancement may pave new ways in the future for the development of a better and enhanced anammox system.

Conflict of interest The authors declare no conflict of interest.

References Ali, M., Okabe, S., 2015. Anammox-based technologies for nitrogen removal: advances in process start-up and remaining issues. Chemosphere 141, 144–153. Ali, M., Oshiki, M., Okabe, S., 2014. Simple, rapid and effective preservation and reactivation of anaerobic ammonium oxidizing bacterium “Candidatus Brocadia sinica”. Water Res. 57, 215–222. Ali, M., Oshiki, M., Rathnayake, L., Ishii, S., Satoh, H., Okabe, S., 2015. Rapid and successful start-up of anammox process by immobilizing the minimal quantity of biomass in PVA-SA gel beads. Water Res. 79, 147–157.

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CHAPTER 1 Anammox process

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CHAPTER

Abundance of ammoniaoxidizing bacteria and archaea in industrial wastewater treatment systems

2

Vidya Sawant and Hitesh S. Pawar DBT-ICT Centre for Energy Biosciences, Institute of Chemical Technology, Mumbai, India

2.1 Introduction Nitrogen-mediated biotransformation by nitrogen cycle plays a significant role in human lives and ecological systems. Assimilation, ammonification, nitrification, denitrification, anaerobic ammonium oxidation (anammox), denitrifying anaerobic methane oxidation (DAMO), and dissimilatory nitrate reduction to ammonium (DNRA) are all part of the nitrogen cycle (Canfield et al., 2010; Raghoebarsing et al., 2006). Several physiologically distinct classes of microorganisms are involved in these processes (Fig. 2.1). Gaseous nitrogen is converted to ammonia for assimilation by nitrogen-fixing bacteria at the time of nitrogen fixation. Ammonification is a more complex process that involves bacteria to convert organic nitrogen to ammonia (NH3 + ) or ammonium (NH4 + ). Nitritation (conversion of ammonia to nitrite) mediated by AOB is the primary driver of chemolithotrophic nitrification. Denitrification, a process by which nitrate (NO3
) is converted to nitrite (NO2
), nitric oxide (NO), nitrous oxide (N2O), and dinitrogen gas (N2), allows mixotrophic denitrifiers to utilize organic carbon sources to gain energy under oxygen deficient conditions. Anammox bacteria have the ability to convert nitrite and ammonium to dinitrogen gas under complete anoxemia ( Jetten et al., 2009). Based on the currently available information on ammonia oxidation, the ammonia-oxidizing microorganisms (AOM) can be broadly classified into three types that involve the autotrophic ammonia-oxidizing bacteria (AOB), heterotrophic ammonia-oxidizing bacteria (anammox), and mixotrophic ammonia-oxidizing archaea (AOA). Nitrogen-containing contaminants are one of the most important environmental contaminants found in different forms of wastewater and are a major cause of eutrophication (Yin et al., 2018). Nitrification is mediated by the ammonia oxidation, facilitated by specialized groups of bacteria such as “AOB” and archaea “AOA” Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00007-3 Copyright # 2022 Elsevier Inc. All rights reserved.

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

FIG. 2.1 General pathways for microorganism-mediated nitrogen transformation. The square emphasizes the close relationship between nitrous oxide emission and the process for ammonia oxidation. The different genes involved in different processes are shown in the figure. Steps 5–8 mentioned in the figure represent ammonification, with gdh representing glutamate dehydrogenase and ure representing urease. Steps 2–4 mentioned in the figure depict ammonia oxidation (autotrophic, heterotrophic, and archaeal): hao stands for bacterial hydroxylamine oxidoreductase (unknown gene/enzyme in AOA); amo stands for ammonia monooxygenase. Nitrite oxidation, the second stage of nitrification, is also depicted within Steps 2–4. Steps 5–8 labeled in the figure represent key steps in denitrification: nar stands for nitrate reductase, nir stands for nitrite reductase, nor stands for nitric oxide reductase, and nos stands for nitrous oxide reductase. The nitrogen fixing enzyme nif is represented by Step 1. The denitrifying anaerobic methane oxidation (DAMO) (unknown gene/enzyme in DAMO) is shown in the form of triangle. Steps 9 and 10 represent dissimilatory nitrogen reduction, which includes nas, which codes for nitrate reductase, and nir, which codes for nitrite reductase.– AOA, ammonia-oxidizing archaea; AOB, ammoniaoxidizing bacteria (Guo et al., 2013).

(Stein, 2019). The first report on “comammox,” a bacterium that carries out complete ammonia oxidation from ammonia to nitrate was published in 2015 (Daims et al., 2015; Van Kessel et al., 2015). The rate-limiting step is microbe-mediated nitrification, which involves the ammonia oxidation to nitrite catalyzed by the autotrophic AOB and mixotrophic AOA. For the purpose of balancing the depleted and oxidized nitrogen reservoirs in nature, which links the removal and mineralization of environmentally accessible nitrogen together (Nishizawa et al., 2016).

2.2 Key enzymes involved

The removal of nitrogen from wastewater is important to prevent adverse effects on marine life and depletion of oxygen in the receiving water stream. Currently, all countries have established strict discharge requirements for the removal of nitrogenmediated contaminants from water environments. Biological Nitrogen Removal treatments are widely adopted among all treatment methods as it is more efficient and a relatively less expensive process (Guo et al., 2007). Microorganisms being the key drivers of the process affect the success and stability of WWTPS and play a significant role in the pollutant removal process (Daims et al., 2006). Conventional BNR needs a lot of aeration energy and carbon source consumption from wastewater for full oxidation to nitrate followed by heterotrophic denitrification (Kartal et al., 2006). Not only do AOB, AOA, anammox, and heterotrophic nitrifiers open up new routes in the global nitrogen cycle, but they also offer new options and alternatives for BNR from wastewater. Ammonia-oxidizing bacteria were once thought to be a major player in the ammonia oxidation process. The amo A gene, which is found in large numbers of AOB and AOA, has been discovered to be an indicator gene of ammonia oxidation due to recent advances in molecular biology techniques. In this book chapter we summarize the diversity and abundance of ammonia-oxidizing bacteria (AOB) and ammonia-oxidizing archaea (AOA) in WWTPS. In addition, the mechanism for the process of ammonia oxidation adapted by these ammonia-oxidizing microorganisms (AOM), key enzymes involved in the process and techniques available for their quantification are discussed in detail.

2.2 Key enzymes involved The two main enzymes required for conserving energy during oxidation of ammonia are ammonia monooxygenase (AMO) and hydroxylamine oxidoreductase (HAO). Since they both produce substrate and electrons, these enzymes are codependent in vivo (Guo et al., 2013). Current investigation shows that AMO contains a trinuclear structure as its active site rich in Cu2+ ions (Gilch et al., 2010). AMO is made up of three membrane-bound polypeptides, Amo A, Amo B, and Amo C, which are generated by the amo operon’s genes amo A, amo B, and amo C, respectively (Guo et al., 2013). The amo A gene derived from ammonia monooxygenase is a kind of indicative gene for ammonia oxidation (Yin et al., 2018). A polypeptide of mass 31.99 kDa is predicted by the amo A gene. A conclusion has been made that the amo A gene contains site for substrate oxidation based on the fact that amo A polypeptide can covalently bind 14CH-14CH (Hyman and Wood, 1985). The amo A gene is important of all the three genes because it provides energy required for the bacterial growth and also the gene forms the product that retains the active site of enzyme (Guo et al., 2013). The amo A gene sequence studies can determine the quantities, species, and activities of AOB. Following the cleavage of periplasmic signal peptide, the amo B gene predicts a polypeptide of 44.26 kDa. amo B gene is most likely to form a complex with amo A. amo C protein may behave like a molecular

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

chaperon to assist the proper folding of amo A and amo B gene to facilitate their appropriate integration into the membrane (Klotz et al., 1997). In addition to this, a detailed analysis of the transmembrane alpha helices of the amo protein shows that amo A, amo B, and amo C consist of six, two, and six transmembrane alpha helices, respectively (Arp and Stein, 2003). However, the particular role of amo B and amo C subunit involved in ammonia oxidation is still unclear. HAO is a catalytic enzyme that catalyzes the oxidation of hydroxylamine to nitrite in the second step of the process of nitrification (Guo et al., 2013). HAO being a water-soluble periplasmic protein is much easier to understand in comparison with AMO. HAO is a trimeric enzyme with a highly complex structure and a single subunit with a molecular weight of 63 kDa. As the most likely components of the active site, each subunit contains seven one uncommon residue of heme P460 and seven ctype hemes (Bergmann and Hooper, 1994; Pearson et al., 2007). P460 is a peculiar part of HAO and an uncommon feature of c-type heme. P460, a c-type heme is a crucial part of HAO which is involved in bond formation between one of the cytochrome centers and a meso carbon of heme. A physiological electron acceptor, namely cytochrome c-554, plays a significant role in the ammonia oxidation process by bacteria (Yamanaka and Shinra, 1974). By transferring electrons from HAO to membrane bound cytochrome c-552, cytochrome c-544 plays a role in the generation of energy by a respiratory system (Kim et al., 2008). In contrast to Nitrosococcus europaea cytochrome c-554, which is a highly basic protein with an isoelectric point of 10.7, Nitrosococcus oceani cytochrome c-554 is an acidic protein with an isoelectric point of 4.6 (Hozuki et al., 2010). Bergmann et al. (2005) discovered a molecular evolutionary basis for orthogonal heme stacking in HAO, tetrathionate reductase, and cytochrome c nitrate reductase by analyzing the sequences of HAO gene clusters from gammaproteobacterial AOB N. oceani and betaproteobacterial AOB Nitrosomonas europaea and Nitrosospira multiformis. Their findings revealed that the hao gene cluster developed as a structural and functional unit from pentaheme nitrate reductase with more effective ammonia catabolism, i.e., under the pressure of increasing hydroxylamine concentrations. The existence of distinct and evolutionarily conserved regions within the HAO and cytochrome c-554 proteins will predict many new molecular ecology approaches to understanding the diversity of AOB and AOA.

2.3 Physiology and cellular structure 2.3.1 Physiology of AOA Individual archaea have sizes ranging from 0.1 to 15 μm and have variable shapes in the form of plates, rods, and spirals. Electron microscopy structure of Nitrosopumilus maritimus reveals that it is a straight rod with a length of 0.5–0.9 μm and a diameter of 0.17–0.22 μm which is identical in size to Crenarchaeota, but smaller than cultivated AOB (K€ onneke et al., 2005). The existence of flagella or intracellular

2.4 Diversity in WWTPs

compartments is not confirmed by electron microscopy, which also means that the cells exist in loose aggregates. Likewise, Nitrososphaera viennensis has a diameter of 0.5–0.8 μm and appears as a spherically shaped slightly small, irregular cocci (Tourna et al., 2011). Hatzenpichler et al. (2008) discovered the first thermophilic ammonia-oxidizing archaea Nitrososphaera gargensis which appears as an irregular shaped cocci with a diameter of 0.9  0.3 μm based on catalyzed reporter deposition (CARD)-FISH and combined with micro autoradiography (MAR) analysis.

2.3.1.1 Kinetics stoichiometry of ammonia oxidation

Initially, N. maritimus strain SCM1 was the only viable strain for physiological research, and it produced valuable insights into the ecological performance of AOA (Stahl and De La Torre, 2012). SCM1 can survive the extreme environmental stresses and still carry out efficient ammonia oxidation in open oceans (MartensHabbena et al., 2009). Under mesophilic conditions, N. maritimus can grow chemolithotrophically by oxidizing ammonia to nitrite. It also has important genes for all three subunits (amo A, amo B, and amo C) of the bacterial ammonia monooxygenase needed for ammonia oxidation (Park et al., 2006). Martens-Habbena et al. (2009) investigated the ammonia oxidation kinetics and stoichiometry of N. maritimus and discovered that it has an evident half-life constant (Km) of 132 nM for total content (ammonium plus ammonia) (3 nM NH3 at near-neutral pH). NH3 + 1:5O2 ! NO2
+ H2 O + H+

The molecular evidence that AOA N. maritimus strain SCM1 occurs in activated sludge bioreactors to extract nitrogen from wastewater was examined and reported by Park et al. (2006).

2.3.2 Physiology of AOB During the ammonia oxidation process, AOBs are chemolithotrophic bacteria that use carbon dioxide as a source of carbon and molecular oxygen as an electron acceptor to convert ammonia to nitrite as their primary source of energy. The physiology of AOB can be determined by inspecting the 16S rDNA gene and the amo A gene, which encodes for key enzyme ammonia monooxygenase present in common for all three groups of ammonia oxidizers (Lehtovirta-Morley, 2018). Table 2.1 describes the morphological and ecophysiological characteristics of AOB in different environments.

2.4 Diversity in WWTPs Several researchers reported higher abundance of AOA than AOB in domestic WWTPs whereas the situation was vice versa for industrial WWTPs. Switch in the abundances and diversity of AOB and AOA has been seen in engineered and

21

Table 2.1 Morphological and ecophysiological features of AOB (Guo et al., 2013).

Bacteria genus

Cell morphology

Nitrosomonas

Rod to ellipsoid

Nitrosospira

Nitrosospira (close heliciform) Nitrosovibrio (acerose arc shaped) Nitrosolobus (polymorphic)

Species

Salt requirement

Urease activity

Substrate (NH3) affinity (Ks) (μM)

Nitrosomonas europaea Nitrosomonas eutropha Nitrosomonas halophile Nitrosococcus mobilis Nitrosomonas ureae Nitrosomonas oligotropha

Halotolerant or moderately halophilic No salt requirement




30–61

WWTPs, marine, biofilms, high ammonia wastewater

+

1.9–4.2

Nitrosomonas Nitrosomonas Nitrosomonas cryotolerans Nitrosomonas Nitrosomonas

marina, aestuarii,

Obligatory halophilic

+

50–52

Oligotrophic lakes, moderately acid soils, WWTPs Marine sediments, free waters, estuary, marine, WWTPS

communis, nitrosa

No salt requirement

+/


42–49

No salt requirement

+/


19–46

Nitrosospira briensis, Nitrosospira spp., Nitrosovibrio tenuis

Habitats

agricultural soils, WWTPs eutrophic environments Activated sludge, biofilms, acid soils, and neutral arable fields.

2.4 Diversity in WWTPs

natural ecosystems, thus providing a relevant evidence of their contribution to ammonia oxidation.

2.4.1 Diversity of AOA While previous studies had confirmed the existence of AOA only in extremophilic conditions, a gene analysis of ammonia-oxidizing archaea (AOA) recently revealed that they thrive in a variety of natural habitats with varying characteristics. In reality, AOA are the most diverse and important players in the ammonia oxidation process in both terrestrial ecosystems (Leininger et al., 2006) and oceanic ecosystems (Wuchter et al., 2006). AOA was reported in WWTPs in the early 2006 (Yin et al., 2018). The factors which influence the community structures and diversity of AOA are dissolved oxygen concentration, pH value, and chemical oxygen demand (Chen et al., 2017). Candidatus Cenarchaeum symbiosum, a marine sponge symbiont, encodes amoA, a gene needed for ammonia oxidation, and is thus classified as an AOA (Hallam et al., 2006a; Delong et al., 2006). Chen et al. (2017) investigated the diversity and abundance of AOA and AOB in a WWTP operated in biological aerated filter type (BAF) mode and carried out quantification studies using PCR. According to their results the AOA clearly exceeded the population of AOB and the relative abundance of amoA AOA varied from 6.32  103 to 3.8  104 copies/ng DNA whereas the abundance of AOB amoA genes was 1.32  102 copies/ng DNA in the filter layer of BAF. Large amounts of antibiotic spiramycin in wastewater result in a high growth and abundance of AOA in the nitrifying population of pharmaceutical WWTPs, according to Zhang et al. (2015).

2.4.2 Diversity of AOB In general, Υ proteobacterial AOB can be found in aquatic ecosystems, while β proteobacterial AOB can be found in terrestrial habitats and WWTPs (Lehtovirta-Morley, 2018). The only identified gammaproteobacterial species are N. oceani and Nitrosococcus halophilus while the genera Nitrosomonas and Nitrosospira belong to betaproteobacterial class. Betaproteobacterial AOB have been discovered in activated sludge, terrestrial habitats, soil, coastal, and marine ecosystems and are used as molecular markers in microbial and ecological research (Nold et al., 2000). Conversely gammaproteobacterial AOB have only been discovered in the aquatic environments (Kowalchuk and Stephen, 2001; O’Mullan and Ward, 2005). Nitrifying WWTPs consist of ammonia-oxidizing bacteria (AOB) as their crucial communities (Rowan et al., 2003). Rowan et al. (2003) used polymerase chain reaction (PCR) and phylogenetic studies to examine the diversity and community structures of betaproteobacterial AOB in two different treatment reactors—a biological aerated filter (BAF) and a trickling filter receiving identical wastewater. They found variable community structures of AOB in separate sections of each reactor. Tsuchiya et al. (2020) used quantitative PCR to examine amo A genes of AOB, AOA, and a

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

total ammonia oxidizer (comammox) in the biofilm produced on the carrier used to extract ammonia in a full-scale WWTP. They discovered that the biofilm carrier had a higher relative abundance of AOB and comammox than the suspended activated sludge in the WWTP, while the activated sludge had no relative abundance of AOA.

2.5 Mechanism of action of AOA and AOB 2.5.1 Mechanism of AOA Currently only a few reports are available for pathways of ammonia oxidation in archaea. Prior research into archaeal lipid biological markers, as well as subsequent analyses, revealed the autotrophic growth capability in planktonic Crenarchaeota and a chemolithotrophic metabolism based on the concentration of ammonia was observed in case of N. maritimus (K€onneke et al., 2005). Genes for a near-complete oxidative tricarboxylic acid cycle and a modified 3-hydroxypropionate cycle have been found in Cenarchaeum symbiosum. As a result, the uncultivated C. symbiosum will act as an obligatory autotroph or a mixotroph, consuming both organic matter and carbon dioxide as a substrate (Hallam et al., 2006a; Delong et al., 2006). Tourna et al. (2011) studied the growth rates of N. viennensis in the presence of pyruvate, suggesting that this organism has the ability to develop in a mixotrophic environment. N. maritimus uses a modified 3-hydroxypropionate/4-hydroxybutyrate (HP/ HB) cycle to autotrophically fix inorganic carbon while retaining a small capacity for organic carbon assimilation, according to genomic sequence studies (Walker et al., 2010). Based on the lack of hydroxylamine oxidoreductase genes in N. maritimus, Walker et al. (2010) proposed a hypothesized mechanism for ammonia oxidation by AOA in which ammonia is oxidized to nitrite through nitroxyl (HNO) instead of hydroxylamine as in AOB. Ammonia oxidation is the first step in nitrification, in which ammonia is converted to nitrite through nitrate. Molecular evidence indicates that the archaeal amoA gene and transcripts, which code for the enzyme ammonia monooxygenase (AMO), outnumber their bacterial equivalents in the oceans and terrestrial hot springs (Nishizawa et al., 2016). Because of its tertraether lipid-based membrane, AOA has adapted to extreme conditions by providing lower ion permeability, resulting in less futile ion cycling and lower levels of maintenance energy than AOB (Valentine, 2007). The data needed to comprehend AOA’s ecological adaptations, such as the stability of ammonia monooxygenase (AMO), mRNAs, and ribosomal proteins, has yet to be discovered (Yin et al., 2018). In AOA, the enzyme that catalyzes the conversion of NH2OH to NO2 or HNO to NO2 is still not known (Guo et al., 2013). Though a much larger information is available related to AOA, the patterns and molecular mechanisms of ammonia oxidation and autotrophic carbon assimilation are still unresolved. Further investigation is required to understand the different pathways and diversity of metabolism in AOA (Fig. 2.2).

2.5 Mechanism of action of AOA and AOB

FIG. 2.2 Schematic illustration of ammonia-oxidizing archaea pathway (Nishizawa et al., 2016).

2.5.2 Mechanism of AOB Mukherjee et al. (2002) revealed the biochemistry of ammonia-oxidizing nitrifiers oxidize the ammonium to nitrite; in contrast, nitrobacteria oxidize nitrite to nitrate. Enzyme ammonia monooxygenase (AMO) is involved in the process of ammonia oxidation which utilizes ammonia and dioxygen as substrates during which an atom of oxygen is reduced to form water while the second atom is utilized in the formation of hydroxylamine. Hydroxylamine oxidoreductase (HAO) is a periplasmic multiheme enzyme which oxidizes hydroxylamine to nitrite. The oxidation of hydroxylamine produces a change in concentration of protons, which induces ATP synthesis. AMO : NH3 + 2H+ + 2e
+ O2 ! NH2 OH + H2 O HAO : NH2 OH + H2 O ! NO2
+ 5H+ + 4e


The metabolic pathways involved in mixotrophic ammonia oxidation are currently unclear (Guo et al., 2013). Paracoccus denitrificans (Moir et al., 1996), Alcaligenes faecalis ( Joo et al., 2005), Pseudomonas putida (Daum et al., 1998), and many other different microbial species are currently used in research on metabolic pathways required for heterotrophic ammonia oxidation. According to some study, the biochemical processes of heterotrophic nitrification vary from those of autotrophic nitrification (Wehrfritz et al., 1993). Ammonium oxidation by these heterotrophs; an

25

26

CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

inorganic and an organic pathway has been proposed by Koops and PommereningR€ oser (2001). These pathways are represented as follows:
(a) Inorganic Pathway: NH4 + ! NH2OH ! NOH ! NO
2 ! NO3 (b) Organic Pathway: R-NH2 ! R-NHOH ! R-NO ! R-NO2 ! NO3


There are four major variations in metabolic pathways between heterotrophic and autotrophic AOB (Guo et al., 2013). To begin with, most heterotrophic nitrifiers are able to utilize both organic and inorganic nitrogen (Brierley and Wood, 2001). Second, the strains of heterotrophic nitrifiers develop different intermediate and final products during heterotrophic nitrification (Guo et al., 2013). Third, certain mixotrophic nitrifiers aerobically denitrify part or all of the nitrite, turning it to dinitrogen gas or trace nitrogen oxides at the same time ( Joo et al., 2007). Fourth, aerobic nitrite oxidation from ammonia or other organic nitrogen sources provides no metabolic energy or reducing capacity, while autotrophic nitrite oxidation provides Adenosine triphosphate (ATP) for AOB multiplication and conservation reactions including carbon dioxide reductive fixation (Fig. 2.3). NH2OH HNO2 + H2O + 4H+ H2O2 2H2O

HNO2 + 4H+

HAO

4e–

Periplasm Cc554 NH2OH NH3 + H2O + O2

Cc552

4e–

NAD

NADH +

nH

HNO2 NO

+ Ccm552 4H

QH2 AMO

Cu NiR

4e– ?

RH ROH

NADH dh

CCP

C-P460

QH2

AMO

alternative substrate oxidation

4H+

2H+

bc1

Cuaa3

2H+ Cytoplasm

1/2 O2 4H+

N2O

NOR

H2O

FIG. 2.3 Nitrosomonas europaea electron transport. The solid lines represent known electron transfer pathways, while the dashed lines represent hypothetical electron transfer pathways. HAO, hydroxylamine oxidoreductase; AMO, ammonia monooxygenase; CP460 stands for cytochrome P460; and Q stands for ubiquinone-8. Cu NiR stands for copper-containing nitrite reductase; CCP stands for cytochrome c peroxidase; NOR stands for nitric oxide reductase; CcM552 stands for membrane cytochrome c552. Substrates other than ammonia are called alternative substrates and do not generate hydroxylamine (Whittaker et al., 2000).

2.7 AOA at low DO or in special WWTPs

2.6 Competition and symbiotic relationships between AOMs The competition and symbiotic associations within Ammonia-Oxidizing Microorganisms (AOMs) is relatively complicated. AOB and AOA compete for the same substrates and ecological habitats, but AOB supplies nitrite, which allows anammox bacteria to thrive in anoxic microenvironments. Coolena et al. (2007) and Lam et al. (2007) stated that anammox bacteria are present at the similar levels in the Black Sea, so it is presumed that they will compete for ammonium, and that AOA could supply nitrite for anammox. Interaction between AOA and NOB is also an intriguing research subject. These possibilities and subjects, on the other hand, need to be explored further. The conversion often necessitates a number of different combined procedures. N is converted to dinitrogen gas from ammonium by a variety of processes, including nitrification by bacteria, nitrification by AOA + denitrification by denitrifiers, nitrification by AOB + denitrification by denitrifiers, simultaneous nitrification and denitrification by heterotrophic denitrifiers, nitrification and denitrification by heterotrophic bacteria nitrifiers, anammox + AOB nitrification, and anammox + AOA nitrification + anammox (Guo et al., 2013). More study is needed to figure out which pathways are the most common and under what circumstances each one works best. It is also important to figure out how much nitrogen is lost in different ecosystems and habitats as a result of each combined pathway. Finally, an inhibitor must be detected to distinguish each player’s position in the global N cycle, allowing for the differentiation of bacterial vs archaeal and autotrophic vs heterotrophic ammonia oxidation contributions (Guo et al., 2013). New processes for treating high-ammonium wastewaters could be established based on syntrophic and competitive interactions among AOB, AOA, and anammox.

2.7 AOA at low DO or in special WWTPs Gene sequence analysis using quantification techniques recently revealed the universality of AOA in nitrifying bioreactors from municipal WWTPs all across the globe (Park et al., 2006; Limpiyakorn et al., 2011; Ye and Zhang, 2011; Zhang et al., 2011). In nitrifying bioreactors, 50 of 75 archaeal amoA gene sequences were collected and found to be provisionally distributed in 4 major evolutionary clusters (Park et al., 2006). Since all of the archaeal amo A sequences were found in sludge samples collected from WWTPs with long SRTs and low DO levels, DO concentration and SRT can play a role in the presence of AOA (Wells et al., 2009). Wells et al. (2009) examined the occurrence of ammonia-oxidizing populations in a full-scale activated sludge bioreactor with high aeration. Quantitative bacterial and archaeal amo A gene analysis was carried out using PCR and the results stated that in about 15% of the samples archaeal amoA was present. Based on their results, the researchers concluded that AOB had a greater contribution to nitrogen removal in highly aerated bioreactors than AOA.

27

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

Kayee et al. (2011) carried out studies on relative abundances of AOB and AOA in WWTPs of Thailand involved in autotrophic ammonia oxidation and depending on the amo A gene estimation it was examined that AOA was found in higher proportion than AOB. Guo et al. (2013) looked for 52 nitrifying WWTPs in Germany, Switzerland, United Kingdom, and Austria, for AOA using PCR and FISH. According to their study, AOA were only found in WWTPs treating wastewater from leather factories and oil refineries, and not in any WWTPs treating municipal wastewater. Furthermore, CARD FISH analysis showed that under nitrifying conditions, no CO2 fixation of AOA occurred. Based on their findings the researchers concluded that AOA have a relatively minor role to play in WWTPs but they play a major role in other functions such as alkane degradation during special wastewater treatment. In principle, AOA are used in WWTPs with long Solid retention times (SRTs) and low DO, as well as those that handle special wastewater and have the potential for ammonia oxidation. Hence to convert ammonia to nitrite it is possible to employ the use of AOA. The censorious point of AOA’s application in nitrogen removal, however, is identifying an effective method for enriching their development in engineered systems, due to their lower growth rates. In a sequencing batch reactor (SBR), stable enrichment of anammox bacteria has been achieved, although it is still uncertain if SBR is feasible for the enrichment of AOA; indeed, anammox bacteria are still slow growers in comparison to AOA (Guo et al., 2013).

2.8 Factors influencing AOB abundance and diversity 2.8.1 Ammonia levels The dominance of a particular AOB species depends on the concentration of ammonia level as the affinity for substrate (ks) values differs among species. In general, the members of N. Nitrosospira are dominant at low ammonia concentrations in contrast to the members of N. europaea found to be dominant in environments rich in ammonia levels. As ammonia levels rise, so does the amount of AOB and the amount of ammonia oxidation activity. A large excess of free ammonia, on the other hand, can inhibit the growth of AOB (Guo et al., 2013).

2.8.2 FNA and nitrite While the effect of nitrite on AOB’s ammonia-oxidizing rate has been extensively studied, little is known about the physiological mechanisms and gene regulation of AOB at higher nitrate concentrations. At a concentration of 30 mM, nitrite inhibited the growth of the AOB Nitrosomonas sp. (Painter, 1970). Anthonisen et al. (1976) studied the inhibition of nitrification caused by unionized ammonia (FA) and unionized nitrous acid (FNA) in activated sludge treating municipal WWTPs and the inhibitory concentrations for FA ranged from 0.1 to 150 mg/L for FA and 0.2 to 2.8 mg/L for FNA although the inhibition was not permanent and could be lifted by adjusting the levels of FA and FNA.

2.9 Quantification techniques

Stein and Arp (1998) discovered that FNA inhibited the N. europaea enzyme AMO. Cua and Stein (2011) examined the effect of relatively high nitrite concentration (10 and 20 mM) on oxidation rates of ammonia and the results stipulated that the rate of oxidation of ammonia decreased notably only in NaNO2-amplified incubations of N. eutropha whereas no significant inhibitory effects were observed in the ammonia oxidation rate of N. europaea or N. multiformis whereas nitrite concentration of 20 mM significantly reduced the mRNA levels of amoA in N. multiformis and norS in the two different Nitrosomonas spp.

2.8.3 Process conditions and regime Nevertheless of the initial inoculum sludge concentration, for long-term operation varying environmental conditions in WWTPs results in selection of variable types of AOB species. Dytczak et al. (2008) observed the dominance of slower nitrifiers such as Nitrosospira in a strictly anaerobic reactor while a dominance of AOB species such as Nitrosomonas was observed in an anoxic/aerobic reactor with higher rate of ammonia oxidation. Rowan et al. (2003) compared the community structure and diversity of AOB belonging to betaproteobacterial class in two separate reactors handling similar wastewater and discovered that a biologically aerated filter (BAF) had a comparatively lower diversity of AOB as compared to a tricking filter. The diversity and community structures of AOB are influenced by a number of factors, including reactor form, wastewater characteristics, and process activity regimes.

2.9 Quantification techniques Since biomass yield rates are poor and it’s difficult to distinguish mixotrophic contaminants from solid and liquid media, the age-old method of assessing AOB by isolation and culture is inefficient. The quantification techniques most popularly used and their subsequent methodologies are discussed.

2.9.1 DNA extraction Genomic DNA can be extracted from activated sludge sample. Bourrain et al. (1999) conducted optimization studies of the cell lysis stage for DNA extraction from activated sludge samples by implementing two floc dispersion methods (sonication versus stirring with a cation exchange resin) and three cell lysis treatments (lysozyme + SDS, sonication in a water bath, and thermal shock), concluding that the cell lysis treatment yielding the maximum DNA yield is not identical for all sludges.

2.9.2 Quantitative PCR and reverse transcriptional qPCR A quantitative polymerase chain reaction with a specific set of primers can be conducted to detect the abundances of AOA and AOB (Zheng et al., 2021). For the first time in 2009 Well et al. used quantitative PCR to examine the presence of AOA in

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

WWTPS (Yin et al., 2018). Without the time-consuming laboratory cultures, PCR amplification of 16S rDNA segments and subsequent nucleotide sequencing can be used as a novel technique for investigating evolutionary affinities of members of natural populations (Stephen et al., 1996). Park et al. (2006) reported the existence of AOA in activated sludge bioreactors used to remove ammonia from WWTPs using PCR primers targeting archaeal ammonia monooxygenase subunit A (amoA) from five wastewater treatment plants operating with low dissolved oxygen levels and longer residence periods.

2.9.3 High throughput sequencing Zheng et al. (2021) sequenced the purified PCR products on the Illumina MiSeq platform to transverse the phylogeny and transcripts of amo A genes in AOA and AOB. Raw sequences can be screened and filtered using QIIME (acronym stands for Qualitative Insights into Microbial Ecology) software and public scripts. QIIME (statutory pronounced as chime) is a program that analyzes microbial communities. QIIME has been used to decipher nucleic acid sequencing data from fungi, viruses, archaea, and cultures (Kuczynski et al., 2011). Hundreds of communities can be simultaneously analyzed by means of pyrosequencing. Major barrier in traditional sequencing approach is the unavailability of software to manage massive data for downstream analyses. QIIME, an open source program developed using the PyCogent toolkit, is used to handle the issue of acquiring sequencing data from raw sequences to analysis and database deposition (Caporaso et al., 2010). QIIME supports a broader range of visualizations and microbial community analyses that are crucial for high profile studies including network analyses and histograms. To allow user to interact with the data QIIME also provides a graphical insight (Caporaso et al., 2010).

2.9.4 Phylogenetic analysis Since the first complete isolation of chemolithotrophic AOB N. europaea at the end of the nineteenth century, extensive research has been carried out to assess changes in AOB population structures, isolate and improve strains, classify molecular properties, and investigate the evolutionary relationship of these species (Guo et al., 2013). Phylogenetic analysis of AOB revealed that they primarily constrained to β-Proteobacteria and Υ-Proteobacteria classes (Kowalchuk and Stephen, 2001). Generally the analysis of diversity and evolutionary relationships for AOB is based on comparative analyses of the active site of ammonia monooxygenase (amo A) and their genes encoding 16S rDNA (Purkhold et al., 2000). Aakra et al. (2001) performed a comparative phylogenetic analysis of 31 AOB strains and discovered that phylogenetic trees based on 16S rDNA and amo A sequences of most AOB groups were highly consistent. However, due to the expansion of 16S rDNA genes from closely related nonnitrifying bacterial species, a comparative analysis approach

2.10 Environmental factors affecting AOA and AOB

FIG. 2.4 In the BAF filter layer, phylogenetic tree of various amoA gene sequences from AOA (Chen et al., 2017).

using the 16S rDNA gene for physiological inspection of AOB species may result in an exaggeration of AOB both in terms of diversity and numbers (Fig. 2.4).

2.10 Environmental factors affecting AOA and AOB 2.10.1 Ammonia concentration Nitrogen in the form of ammonia is a common substrate that has a significant impact on ammonia-oxidizing microorganisms’ growth. AOA has a stronger ammonia affinity than AOB (Martens-Habbena et al., 2009). The earlier statement indicates the amount of AOA amoA gene will decrease significantly with gradual increase of ammonia concentrations in WWTPs. Thus when exposed to higher ammonia concentrations AOA might face a suppression as compared to AOB. As per the information provided by Gao et al. (2016) AOB have a higher level of competition than AOA which means abundance of AOB under high ammonia concentrations is greater as compared to AOA. Therefore AOB dominates over AOA when higher amount of nitrogen is available as a substrate in the form of ammonia. It can be concluded that differences in the quantity and structures of AOB and AOA can be due to the amount of ammonia in the wastewater, which is determined by the form of wastewater.

2.10.2 Temperature The effect of temperature on ammonia-oxidizing microorganisms is determined by the action of an enzyme ammonia monooxygenase, which mediates the first step of ammonia oxidation by converting ammonia to hydroxylamine. Lee et al. (2011) determined the acclimation rate and growth rate at steel industry wastewater plant by utilizing AOB and free ammonia as a substrate and they found that oxidation rate was increased by two to three times in granules than flocs when temperature

31

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

increases from 10 to 30°C. A moving bed biofilm reactor (MBBR) was investigated as a tertiary treatment stage for removal of ammonia in high temperature (35–45°C) effluents (Shore et al., 2012). They discovered that bench scale reactors functioning at 35–40°C could eliminate more than 90% of influent ammonia (up to 19 mg/L NH3-N) in both synthetic and industrial wastewater, but that when the temperature was increased to 45°C, no biotreatment actually occurred. Thus temperature affects the activity, growth, and community structure of ammonia-oxidizing microbes.

2.10.3 Oxygen and aeration pressure The nitrification process requires oxygen as a substrate. Since nitrifying microbes have a different affinity for oxygen (AOA > AOB > NOB (nitrite-oxidizing bacteria)), the oxygen level can affect the nitrification mechanism (Yin et al., 2018). The earlier relation indicates that AOA has a strong affinity for oxygen as compared to AOB which makes them at most competitive to survive in hypoxic environments. Li et al. (2018) used a one-dimensional multispecies model to compare the effects of aeration pressure and total nitrogen removal efficiency for wastewater treatment in membrane-aerated biofilms (MAB) containing ammonia-oxidizing archaea (AOA), nitrite-oxidizing bacteria (NOB), and heterotrophic bacteria (HB). The findings showed that lowering the aeration pressure inhibited NOB activity, allowing nitrogen to be removed via simultaneous nitrification and denitrification. Thus nitrogen removal by AOA and AOB could be carried out by monitoring the amount of dissolved oxygen levels which effect their community structures.

2.10.4 Organic loading The growth of ammonia-oxidizing microorganisms indirectly depends on the organic loading. AOB are known to be autotrophic microorganisms, but it’s still unclear if AOA are purely autotrophic or mixotrophic (Yin et al., 2018). Hallam et al. (2006b) carried out genome analysis studies using marine archaea Crenarchaeota, C. symbiosum and proved that the strain had two varied carbon utilization processes: 3-hydroxypropionic acid/4-hydroxybutyric acid cycle (autotrophic metabolism) and tricarboxylic acid cycle (mixotrophic metabolism), indicating the potential of AOA for both autotrophic and heterotrophic metabolism. Prior research on Crenarchaeota based on lipid biomarkers and isotopic analyses had only found autotrophic growth. To conclude, the AOA have more complex metabolic pathways compared to AOB under mixotrophic conditions resulting in varying oxidation capacities.

2.10.5 Salinity Salinity has an effect on the abundance of both AOA and AOB (Bernhard et al., 2010). Salinity stress will weaken the nitrification efficiency by changing the ammonia-oxidizing microbial community structure (Lin et al., 2020). As per the study conducted by Bernhard et al. (2010), along an estuarine salinity gradient for

2.11 Future perspectives

potential nitrification rates AOA were found to be greater as compared to AOA and the abundance of AOA was highest at intermediate salinity. High concentrated organic matter found in high saline industrial wastewaters makes the autotrophic AOB less competitive as compared to heterotrophic AOA. AOB are more dominant in high saline industrial WWTPs as compared to AOA because of autotrophic growth.

2.10.6 DO Heterotrophic nitrifiers have the capacity to survive at lower DO as compared to autotrophic AOB (Ahmad ud din et al., 2008; Jetten et al., 1997; Patureau et al., 2000). Conversely, heterotrophic nitrifiers’ denitrification components may be influenced by DO levels ( Jetten et al., 1997). Under high DO conditions, N2O can make up a greater proportion of Thiosphaera pantotropha’s final denitrification products, while both N2 and N2O can be generated under low DO conditions (Robertson and Kuenen, 1990). As a result, while DO concentration has almost no impact on heterotrophic nitrification activity, it is strongly linked to denitrification activity and product types. AOA can thrive at 0.1 mg/L DO levels (Park et al., 2006; Coolena et al., 2007), in addition to being able to perform at DO levels of 5–7 mg/L (Hatzenpichler et al., 2008). However, it has been proposed that AOA can only play a minor role in ammonia oxidation when DO levels are relatively high.

2.10.7 Sulfide Sulfide-containing estuarine sediments and water columns (Lam et al., 2007; Caffrey et al., 2007), as well as the bio fabrics of a sulfidic geothermal mine (Spear et al., 2007) and sulfide-related hot springs (Reigstad et al., 2008; Weidler et al., 2007), were all found to contain AOA. AOA was found in hypoxic zones of the Black Sea, where sulfide concentrations reached a maximum value of 5 mM. A negative relationship between the abundance of the AOA amoA gene and sulfide concentration has been discovered (Caffrey et al., 2007). These observations of in situ archaeal oxidation of ammonia in the existence of sulfide relate to AOA’s sulfide tolerance. As a consequence, AOA, or at least some ecotypes, are likely to possess unique enzymes or metabolic pathways that enable them to survive and oxidize ammonia in the presence of sulfide (Erguder et al., 2009). However, less information is available about sulfide’s inhibitory effects on AOA growth, especially the inhibitory threshold concentration, necessitating additional study.

2.11 Future perspectives As compared to AOB, AOA can thrive under complex environments like low DO levels or low temperatures in WWTPs, thus solving the problem of inefficient nitrification under these conditions. Hence studies on enrichment of AOA in need to be conducted further. Currently hypothetical mathematical models have been

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CHAPTER 2 Abundance of ammonia-oxidizing bacteria and archaea

developed by taking AOA as a model organism along denitrifying bacteria or anammox bacteria; further research regarding the actual removal of ammonia from WWTPs is still needed. Thus contribution of different microorganisms involved in nitrogen removal from wastewater under variable environmental conditions must be studied to better understand the coexistence, collaboration, and competition processes between the microbes associated with the feature of removal of nitrogen. At the moment, AOA’s capacity and practical application in WWTPs is minimal. Studies on the competition between AOB and AOA, as well as AOA and anammox, should be performed in particular. Since both AOA and anammox develop at a consistent rate, enrichment takes a lot of practice. Therefore a detail of their enrichment methods is required for their broad application in WWTPs. However, whether MBR and SBR can be used to solve the bottleneck linked with slower ammonia oxidation in DO limiting conditions in systems controlled solely by AOB is still uncertain. It is important to establish a novel method that makes effective use of some of AOA’s specific characteristics in order to handle industrial wastewater containing inhibitors or wastewater at higher temperatures. CANON is a process which couples AOB and anammox for nitrogen removal under completely autotrophic conditions. A process using a combination of AOA and anammox is also expected to be produced, similar to this one. The relationship of AOB and NOB is a fascinating subject for study. However, further research into these topics is needed. However further exploration is required in these topics (Guo et al., 2013). There is a need for a more systematic census of AOA and physiological verification of their nitrifying activity in WWTPs since there is no agreement on the functions of AOA in wastewater treatment. More expertise in this area will help in elucidating the role of AOA in nitrogen removal from WWTPs and extending the applications of AOA for environmental protection.

2.12 Conclusion Due to the increasing interest on studies related to global climate change and eutrophication, our understanding about the process of ammonia oxidation and nitrogen cycle has been enhanced with respect to biochemical, ecological, and microbiological perspectives. We’ve gradually modified our understanding of oxidation of ammonia pathways and players in the microbial N cycle thanks to discoveries like anammox, AOB, and AOA, thanks to modern molecular biotechnology. Currently many novel species of AOB and AOA mandatory for ammonia oxidation are discovered. The present book chapter mainly focuses on the discussion of the diversity and abundance of AOB and AOA in WWTPs. In most of the instances the abundance AOB are found to dominate AOA in several WWTPs. Recent studies have concluded that Nitrososphaera sister cluster is found to be the dominant archaeal species involved in ammonia oxidation whereas N. europaea are the most active ammonia-oxidizing bacterial species (Zheng et al., 2021). A detailed coordinated discussion is made on the key enzymes involved

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(AMO and HAO) and the mechanism of action of AOB and AOA needed for ammonia oxidation. Further the book chapter also scrutinizes the quantification techniques and environmental factors that affect several parameters like ammonia oxidation, salinity, temperature, organic loading, and oxygen concentration required for ammonia oxidation.

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Autotrophic nitrification in bacteria

3

Moupriya Naga, Dibyajit Lahiria, Sougata Ghoshb, Sujay Ghoshc, and Rina Rani Rayd Department of Biotechnology, University of Engineering & Management, Kolkata, West Bengal, India bDepartment of Microbiology, School of Science, RK University, Rajkot, Gujarat, India c AMH Energy Pvt. Ltd., Kolkata, West Bengal, India dDepartment of Biotechnology, Maulana Abul Kalam Azad University of Technology, Haringhata, West Bengal, India a

3.1 Introduction Nitrogen is an essential component for the sustenance of life on earth as majority of biomolecules viz. proteins or nucleic acids possess it. Due to its abundant availability in chemically inert form (N2) in the atmosphere, it becomes very difficult to convert this element into more useful components and enter into biogeochemical cycle. It has been reported that biological N2 fixation (BNF) is the main natural pathway for nitrogen (N2) to enter biogeochemical cycles (Fowler et al., 2013) during the Archean era. However, during the industrial revolution, Haber-Bosch process helped in the massive production of ammonia from atmospheric nitrogen involving huge consumption of fossil fuels and subsequent elevated productions of nitrogenous fertilizers. This has led to disruption of N2 cycles owing to the massive distribution of mineral nitrogen to agricultural lands, having deleterious imbalances in the biosphere. Hence, there is an urgent need for more sustainable approaches of soil enrichment by organic carbon and nitrogen instead of free ammonia during nitrogen fertilization involving renewable resources. This will ensure the long-term soil fertility preservation and right balance of elements in the pedosphere ( Jones et al., 2013). Atmospheric N2 is fixed naturally by diazotrophs (few bacteria and archaea species) while photosynthetic eukaryotes especially leguminous plants rely on symbiotic relation with diazotrophs for their nitrogen demand (Hartmann and Barnum, 2010; Kneip et al., 2007). This helps in reducing or removing the need for application of nitrogenous fertilizers. Thus improved knowledge of biological nitrogen fixation (BNF) will help the farmers to have better understanding of N management for optimal crop products along with reduced release of harmful substances to the environment (Liu et al., 2011). NF is performed by a small fraction of prokaryotes commonly termed as diazotrophs and thus limits the global primary productivity (Lebauer and Treseder, 2008; Gaby and Buckley, 2011; Raymond et al., 2004). Each year BNF is Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00003-6 Copyright # 2022 Elsevier Inc. All rights reserved.

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reported to contribute 40–100 Tg (1012 g) nitrogen to the terrestrial ecosystem (Vitousek et al., 2013). Soil N2 fixing ability is chiefly determined by the abundance, diversity, and community composition of diazotrophs (Hsu and Buckley, 2009; Lindsay et al., 2010; Reed et al., 2010; Stewart et al., 2011) revealing proteobacteria (α, β, and γ; 56%) and cyanobacteria (10%) to be the most abundant soil nitrogen fixers. There are four different types of phototrophic diazotrophs: photosynthetic green and purple bacteria; cyanobacteria; anaerobic heliobacteria containing bacteriochlorophyll, Bchl g; and the chlorophyll-containing rhizobia that can form nodules on the stems and roots of leguminous plants. Green and purple sulfur bacteria are reported to fix nitrogen and hence anaerobes, whereas purple nonsulfur bacteria such as Rhodospirillum rubrum and Rhodobacter capsulatus are photosynthetic in nature (Gallon, 2001). Green nonsulfur bacteria do not fix nitrogen. Cyanobacteria are oxygen evolving phototrophs that are reported to make a huge contribution to global nitrogen fixation. Strains like BTAi1 can grow in nitrogen-free medium and can reduce acetylene, which is diagnostic of nitrogen fixation.

3.2 Symbiotic nitrogen fixers Understanding the process of BNF involves studying the biology and biochemistry of host plants and diazotrophs association. This includes identification of desired microorganism, searching for intracellular accommodation, studying the plant microbiome including the harmful microbes, and finding ways to control their ill effects on BNF. Plants have developed several associations with diazotrophs to acquire atmospheric nitrogen depending on the interrelationship between plant and microbe. Mainly three categories exist (Fig. 3.1): loose associations with free-living nitrogen fixers—the diazotrophs, intercellular endophytic associations—plant growthpromoting rhizobacteria (PGPR), and endo-symbioses-cyanobacteria.

FIG. 3.1 Classification of symbiotic N2 fixers.

3.2 Symbiotic nitrogen fixers

Diazotrophs are widely divided into two groups: root nodule bacteria and plant growth promoting bacteria (PGPR) (Fig. 3.1). Root nodule bacteria involve rhizobia, frankia, and cyanobacteria. While rhizobia include various types of α and β proteobacteria having symbiotic relationship with legume plants and Parasponia species, Frankia undergo close association with actinorhizal plants. PGPR are endophytic colonizers that do not invade plant tissues but use chemotaxis to respond to root exudates (Santi et al., 2013; Compant et al., 2010). Many PGPRs (Herbaspirillum, Azoarcus, and Gluconacetobacter) develop associative or endophytic associations with cereals (Pedraza, 2008).

3.2.1 Molecular mechanism of endosymbionts The cross talk between legume and rhizobia involves secretion of various phenolic molecules such as flavonoids and isoflavonoids into rhizosphere that are successfully taken up by rhizobia. Upon intake, these molecules bind with transcription regulators involved in NodD and produces lipo-chito-oligo-saccharides (LCOs) called Nod factors. These Nod factors are very important components and signal molecules that are involved in nodule organogenesis (Oldroyd and Downie, 2008). Activation of Nod factor molecules mediated by LysM domain receptor kinase triggers plant cell division and formation of meristem. The process helps rhizobia invade legume roots via cracks by endocytosis and continue with their intercellular colonization of epidermal cells (Sprent and James, 2007) leading to differentiation into nitrogen-fixing bacteroids within plant organelle called the symbiosome that controls nutrient exchange. Formation of nodules can be of two types: indeterminate or determinate depending on the active functioning of meristem during the lifetime of nodule. Bacteria can also interact with its host plants directly via the bacteria surface polysaccharides which includes exopolysaccharides (succinoglycans and galactoglucans), capsular and lipopolysaccharides, and cyclic β-glucans.

3.2.2 Molecular mechanism of nodule formation Plant rhizosphere comprises of unique carbon sources in the form of low molecular weight organic compounds like sugars, fatty acids, amino acids, sterols, vitamins, growth factors, etc. known as exudates. Root exudates offer a rich nutrient source for a unique population of prokaryotes and eukaryotes to thrive on (Turner et al., 2013). They also help fight against various infection causing microorganisms via secretion of varied phytochemicals (Baetz and Martinoia, 2014). Thus diazotrophs surviving in the root nodules possess important signal molecules like Nod factors that help them plant immunity and permit the invasion within plant. After successful invasion and nodule formation, it has been found that proliferation of disease causing microbes decreases but this does not help in efficient nitrogen fixation. Thus several adaptations have taken place in order to overcome these shortcomings such as: many plant genera have short peptide known as defensin that controls the behavior of the symbiotic microorganism; some plant clades have small, nodule-specific and

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CHAPTER 3 Autotrophic nitrification in bacteria

cysteine-rich peptides that interferes with cell cycle of both plant and bacterial genomes, perturbs membrane stability, alters gene expression, and promotes terminal differentiation of the Rhizobium species (Maro´ti et al., 2015).

3.2.3 Mechanism of exchange of nutrients and nitrogen In the symbiotic relationship of plant and diazotroph, the most critical step is the exchange of nutrient between the two. The plant provides the bacterial symbiont with necessary carbon sources such as fatty acids, amino acids, sterols, vitamins, growth factors, whereas in return bacteria fix the atmospheric nitrogen thereby satisfying the nitrogen demand of plant. Both organisms change their metabolic routines in order to accommodate to each other’s needs. It has been observed that during symbiosis, genes involved in various metabolic pathways, such as photosynthesis, glycolysis, purine and redox metabolism, amino acid biosynthesis, and metabolite transport, are all upregulated during symbiosis (Benedito et al., 2010) (Fig. 3.2). Nitrogen fixation takes place by nodule formation in the legume root along with differentiation of microbe into rhizobium bacteroids. Once the bacteroid formation is initialized, ammonium assimilation produced by the action of nitrogenase is reduced due to efficient delivery to the plant cell in the form of NH4 and/or NH3 (Fig. 3.2). The mechanism of transport of NH3 involves crossing the bacteroid membrane via diffusion into the plant cytoplasm. The symbiosome space overlapping the bacteroid and plant cell membrane ensures the successful protonation of ammonia to NH4 due to the acidic environment (Day et al., 2001). The transfer of NH3 to symbiosome membrane may take place via two mechanisms: (1) NH3 channel and (2) ion channel (K, Na+, and NH4). At last NH4 enters the infected plant cell cytoplasm where it is converted to organic forms such as amino acids, etc. by the action of aspartate aminotransferase and GS, glutamate synthase (GOGAT) enzymes whose expression is well regulated by Nod genes activated during nodule development (Colebatch et al., 2004.). Lastly, cytosolic GS interacts with nodulin in transporting NH3 and its conversion to glutamine (Masalkar et al., 2010). Nodule formation on legume takes place by two possible ways: Nod factordependent or Nod factor-independent method. The former process includes release of plant signal molecules like flavonoids which are taken by the rhizospheric bacteria that enter in the plant via epidermal cracks. This in turn activates the nod genes of rhizobia bacterial signal molecules such as Nod factors lipo-chito-oligosaccharides (LCOs), thus triggering the early nodulation event within the infection zones leading to accumulation of cytokinin. Nod factor comprises of nodABC gene cascade present in all rhizobia. Once mature nodules are formed, lower O2 concentrations cause differentiation of bacteria to bacteroid where N2 fixation can take place by nitrogenase enzyme complex producing NH3. This NH3 can be transformed into amino acids via the glutamine synthetase-glutamate synthase (GS-GOGAT) pathway or else can be directly diffused through the bacterial cell membrane and be transported to plant cytoplasm via ammonia transporters (e.g., AmtB). Here NH3 can be converted into nitrogenous compounds (viz. amino acids, proteins, and alkaloids) in exchange for

3.2 Symbiotic nitrogen fixers

FIG. 3.2 Symbiotic N2 fixation between bacterial and plant cell. (AmtB, ammonia transporter; cyt bd, cytochrome bd; DctA, dicarboxylate transporter; Glu, glutamate; Gln, glutamine; GS, glutamine synthetase; Mo, molybdenum; NH3, ammonia; Nod factors, nodulation factors; P, phosphorus; S, sulfur.)

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CHAPTER 3 Autotrophic nitrification in bacteria

food molecules, e.g., amino acids, glucose, and other saccharides. Thus plants provide amino acids for maintaining a symbiotic relationship with bacteria and in return bacterial cell cycles amino acids back to the plant for asparagine synthesis (Mus et al., 2016).

3.3 Events of nitrogen fixation The whole nitrogen fixation process is believed to evolve prior to the Great Oxygenation Event. It involves conversion of atmospheric N2 into more biologically accessible forms such as ammonia to legume plants by proteobacteria known as rhizobia sharing a symbiotic relation (Poole et al., 2018; Oldroyd and Dixon, 2014). The process involves nitrogenase enzyme complex acting in near-anoxic conditions (Seefeldt et al., 2009; Leister, 2019). However, rhizobia are obligate aerobes that must respire to meet the high energy demands of nitrogen fixation (Millar et al., 1995; Talbi et al., 2012). This pose as a challenge in symbiotic N2 fixation vaguely known as “oxygen paradox” (Marchal and Vanderleyden, 2000; Schulze, 2004). A solution to this paradoxical situation is dealt efficiently by cooperation between legume plant and rhizobia. During this symbiotic relationship (bacteria enter the plant root via infection threads), alterations in gene expression lead to the development of nitrogen-fixing bacteroids within the host legume root organs forming nodules (Gage and Margolin, 2000). These nodules provide anoxic environment (oxygen concentration is as low as 20–50 nM) for N2 fixation by nitrogenase using oxygenbinding leghemoglobin, oxygen diffusion barrier, and mitochondrial localization (Rutten and Poole, 2019).

3.3.1 Nitrification The process of oxidation of ammonia (NH3) to nitrite (NO2
) and nitrate (NO3
) by formation of intermediates such as hydroxylamine (NH2OH) and nitric oxide (NO) is termed as nitrification (Stein and Klotz, 2016). Due to low retention in soil, nitrites and nitrates can be easily leached out to nearby water bodies causing water pollution leading to downstream eutrophication. Nitrification is a complex process that involves various enzymatic interventions (Hooper et al., 2005). For example, Cubound enzyme ammonia monooxygenase (AMO) catalyzes oxidation of NH3 to NH2OH which is further processed by multiheme enzyme hydroxylamine oxidoreductase (HAO) to NO2
. Nitrite oxidoreductase (NXR) enzymes catalyze the conversion of NO2
to NO3
. In another study, an anaerobic ammonium-oxidizing (anammox) bacterium Kuenenia stuttgartiensis was found to use NO2
in place of O2 as an electron acceptor (Kartal et al., 2011a,b). It was observed that NO and hydrazine were formed as intermediates in the nitrogen production with the help of novel enzymes like nitrite reductase (NIR) and hydrazine synthase (Kartal et al., 2011a,b; Kartal and Keltjens, 2016).

3.3 Events of nitrogen fixation

3.3.2 Nitrate and nitrite synthesis during nitrification Nitrification reaction takes places within the soil by autotrophic ammonia-oxidizing bacteria (AOB) and ammonia-oxidizing archaea (AOA) (Kuypers et al., 2018). For example, in AOB Nitrosomonas europaea, a membrane-bound protein ammonium monooxygenase (AMO) catalyzes the conversion of NH3 to NH2OH in the presence of oxygen and releases the product into the periplasm (Eq. 3.1) (Robertson and Groffmann, 2007). NH3 + O2 + 2e
+ 2H+ Ð NH2 OH+ H2 O

(3.1)

3.3.3 Hydroxylamine oxidoreductase The NH2OH thus produced by AMO is oxidized by hydroxylamine oxidoreductase (HAO) in the periplasm (Eq. 3.2). NH2 OH+ H2 O Ð NO2
+ 5H+ + 4e


(3.2)

For example, successful purification and crystallization of N. europaea HAO reveals it as a 200 kDa homotrimer protein with each subunit containing eight c-type hemes and making it a very complex enzyme (Cedervall et al., 2013). Out of these eight heme groups, seven possesses octahedral Fe centers for efficient electron transfer, while the remaining single heme is catalytically active and is known as the P460 center. The oxidation of NH2OH takes place via the electron transfer through the c-type hemes before reaching the acceptor cytochrome c 554 (Arciero et al., 1991). HAO generally catalyzes the 3e
oxidation of NH2OH to NO under both anaerobic and aerobic conditions (Eq. 3.3) suggesting involvement of a third, unidentified enzyme in the final 1e
oxidation of NO to NO2
. NH2 OH Ð NO+ 3H+ + 3e


(3.3)

Enzymes play a dominant role in nitrification reactions as can be seen from the above-mentioned examples. Thus the role of steric orientation of various amino acids of these enzymes plays a very important role in the entire process. For example, Tyr491 crosslink present at the catalytic heme of HAO is known to disrupt the π-conjugation in the heme, causing substantial ruffling (Cedervall et al., 2013) and leaving the metal center electron-rich, thus helping in the formation of Fe III-NO complex. In another example, enzyme isolated from the anaerobic anammox bacterium K. stuttgartiensis is catalytically active for NH2OH oxidation to NO (Maalcke et al., 2014).

3.3.4 Nitrous oxide production during nitrification The monoheme enzyme CytP460 found in AOB N. europaea has an active site very similar to that found in P460 heme of HAO. Only difference being a Lys70 N-C crosslink heme c 130 -meso-carbon atom exists in the former (Pearson et al.,

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CHAPTER 3 Autotrophic nitrification in bacteria

2007). At high concentration, Cyt460 protects against NH2OH toxicity. The mechanism involves binding of CytP460 to one equivalent of NH2OH to yield Fe II-NO intermediate followed by formation of more stable Fe III-NO species. The NO ligand triggers a NO-based signaling reaction inducing the dissociation of the proximal His ligand to form a five-coordinate complex (Goodrich et al., 2013; Berto et al., 2009; Lehnert et al., 2010). The release of His from Heme of CytP460 causes the fivecoordinate Fe II-NO species to be catalytically inactive that then couples with a second equivalent of NH2OH to generate N2O (Eq. 3.4) (Caranto et al., 2016; Vilbert et al., 2018). NH2 OH Ð N2 O+ H2 O+ 4H+ + 4e


(3.4)

Thus P460 has a purely protective (scavenging) function and not a metabolic one. Another work reported the important role played by Cyt P460s from AOB Methylococcus capsulatus in NH2OH oxidation. Another mechanism of denitrification involves adjusting the metabolism by a cascade of genes like NIRs (NirK) (Beaumont et al., 2005) and NORs (NorB) (Beaumont et al., 2004) for detoxifying NO2
as observed in AOB strains N. europaea. The process is sometimes referred to as nitrifier that includes the release of electron acceptor NO2
into soil for further oxidation to NO3
(Eq. 3.5) by an 180 kDa membrane-bound nitrite oxidoreductase in nitrate-oxidizing bacteria (NOB). NOD contains multiple hemes and Fe-S clusters for electron transport, as well as a Mo metallo-cofactor for efficient catalysis (Daims et al., 2016; Lehnert et al., 2018). NO2
+ H2 O Ð NO3
+ 2H+ + 2e


(3.5)

3.4 Genetic regulation of nitrogen fixation Nitrogen fixation involves a constant supply of active nitrogenase enzyme with ATP and reductant. Moreover, synthesis of nitrogenase and its protection from O2 create an extra metabolic demand within the diazotrophs. It has been observed that diazotrophs grow more slowly on N2 than on NH4 + or NO3
indicating that N2 fixation has an “extra” energy cost. Phototrophs, on the other hand, have evolved in a way better than the diazotrophs employing solar energy to support nitrogenase activity and synthesis. One reason for being interested in these organisms therefore is to unravel the interrelations that exist between photosynthesis and N2 fixation and understanding the role played by phototrophic diazotrophs in global nitrogen cycle. Most widespread O2 sensors detected within rhizobia include: (i) membranebound protein FixL, (ii) hybrid FixL (hFixL), and (iii) FnrN transcription factor. FixL is known to phosphorylate its receiver FixJ protein under aerobic conditions, thus triggering expression of the fixK transcription factor (Wright et al., 2018; Dixon and Kahn, 2004) FixK induces expression of downstream genes by binding as a dimer to an “anaerobox” motif (TTGAT-N 4-ATCAA) upstream of their promoters

3.5 Photosynthesis and nitrogen fixation

(Mesa et al., 2009). On the other hand, FxkR acts as the receiver protein of hFixL inducing expression of fixK, by binding to an upstream “K-box” motif (GTTACA-N 4-GTTACA) (Zamorano-Sa´nchez et al., 2012; Reyes-Gonza´lez et al., 2016; Tsoy et al., 2016). The transcription factor FnrN contains an N-terminal cysteine-rich cluster that makes the protein a direct sensor of O2. The incorporation of biologically fixed nitrogen to terrestrial ecosystems is mostly being done by the rhizobium-legume symbiosis that involves reduction of atmospheric N2 to plant usable forms catalyzed by nitrogenase enzyme complex by energy supplying ATP molecules and electrons required for bond breakage. The entire enzymatic reaction is highly sensitive to the presence of oxygen that is known to inactivate the enzyme. Thus microorganisms have evolved unique regulatory circuits (nitrogen fixing nif genes) to fix the atmospheric N2. The protein NifA works with RNA polymerase sigma factor RpoN (σ 54) to activate the expression of nif genes by binding to an upstream activating sequence (UAS; 50 -TGT-N10-ACA30 ). The N-terminal GAF domain found in NifA protein is ubiquitous and found in all three kingdoms of life, the central domain interacts with σ 54, and C terminal shows a helix-turn-helix (HTH) motif involved in DNA binding (Souza et al., 1999; Steenhoudt and Vanderleyden, 2000). The protein NifA is generally found as a single copy in rhizobia but exceptions are there such as Mesorhizobium loti which possesses two nifA genes, nifA1 and nifA2 (Nukui et al., 2006; Wongdee et al., 2018). In addition to these O2 sensors, rhizobia also possess O2 sensor NifA transcription factor for regulating their final differentiation into nitrogen fixing bacteroids (Dixon and Kahn, 2004; Martinez-Argudo et al., 2005). NifA has a metal-binding cysteine-rich motif in an interdomain linker and possesses nitrogenase component nifH (Salazar et al., 2010; Martı´nez et al., 2004) required for its autoregulation (Th€ony and Hennecke, 2006). It has been observed that regulation of N2 fixation in cyanobacteria in the presence of ammonia does not involve NtrB/NtrC proteins, rather they involve protein NtcA encoded by gene ntcA (see, for example, cyanobacteria are known to form specialized N2 fixing cells called heterocyst during N2 starvation). Strains lacking NtcA were unable to produce heterocysts in the absence of ammonia. NtcA was also required for the transcription of gene hetR encoding a serine-type protease that is a key regulator of heterocyst differentiation. This assumes that NH4 + , or a related metabolite, binds to NtcA and induces a conformational change that prevents this protein from interacting with appropriate promoters.

3.5 Understanding the balance between Photosynthesis and nitrogen fixation 3.5.1 Nitrogen fixation by cyanobacteria Nonsymbiotic bacteria such as cyanobacteria have remained a subject of interests to researchers for a long time in respect of understanding the balance between

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photosynthesis and N2 fixation. Firstly, studies have been aimed at determining whether photosynthesis directly provides ATP and reductant to N2 fixation. Secondly, how cyanobacteria have evolved such efficient strategy to keep aside active, O2-sensitive nitrogenase while simultaneously producing O2. There are two, well-documented strategies that cyanobacteria use in order to minimize contact between nitrogenase and photosynthetically generated O2. First, various strains of cyanobacteria are known to produce specialized N2 fixing cells called heterocyst during N2 starvation. Heterocysts are essentially anaerobic and maintain a spatial separation between N2 fixation and photosynthetic O2 production. Heterocyst generates ATP via cyclic photophosphorylation. They do not fix CO2 and are dependent upon a fixed carbon supply from adjacent, photosynthetically active vegetative cells. For example, Anabaena variabilis, a heterocystous cyanobacterium, can clearly produce nitrogenase in vegetative cells that are capable of simultaneously evolving O2. Second, nonheterocystous cyanobacteria are also found to be capable of N2 fixation by maintaining a temporal separation between the two processes viz. N2 fixation during night and photosynthesis during daytime. Thus there exists an indirect connection between N2 fixation and photosynthetically generated ATP and reductant. N2 fixation in the dark is found to be supported by catabolism of carbon reserves synthesized during the daytime. Moreover, the distribution of intracellular nitrogenase within nonheterocystous cyanobacteria differs. For example, in Oscillatoria limosa nitrogenase can be found in cells throughout the light: dark cycle, even when cultures are not fixing N2 (Gallon, 2001). Many heterocystous species are having biofertilizer potential due to its spatial and temporal separation of nitrogenase activity and oxygenic photosynthesis. Examples include species of Anabaena, Nostoc, Aulosira, Tolypothrix, Scytonema, Cylindrospermum in rice fields which contribute significantly to soil fertility. Thus it can help the economically weak farmers who are unable to invest on costly chemical nitrogen fertilizer.

3.5.2 Nitrogen fixation by rhizobia The most widely studied symbiotic-N2 fixing family of bacteria is Rhizobia which includes genera: Bradyrhizobium, Rhizobium, Azorhizobium, Sinorhizobium, Allorhizobium, and Mesorhizobium (Graham and Vance, 2000). The roots of leguminous plants are inhabited by bacteria leading to the formation of new organ called nodule that helps in N2 fixation by maintaining a symbiotic relationship. Bacterial endosymbionts fix the atmospheric N2 in a plant usable form and in turn plants supply them with an environment rich in carbon as an energy source. The sequence of events during this process involves steps such as preinfection, root colonization and adhesion, hair branching and curling, nodule initiation, bacterial release followed by bacteroid development, nodule function by nitrogen fixation, and nodule persistence (Gage, 2004). For better chances of nodulation, rhizobia compete with a variety of soil microflora including indigenous rhizobial species with respect to nodulation space,

3.5 Photosynthesis and nitrogen fixation

nutrition, and other essential limiting factors (Stephens and Rask, 2000). Thus improving the rhizobial inoculants is a much needed step from ecological standpoint. Two possible solutions include: (1) modifying the environment by varying various abiotic and biotic factors like temperature, soil type, desiccation, pH, fertilizers, fungicides, nutrient supply to suit the bioinoculant strain and (2) modifying the rhizobial bioinoculant strains by doing laboratory or field trials with resident protozoa, fungi, actinomycetes, rhizobacteria to suit the environment (Geetha and Joshi, 2013).

3.5.2.1 Nitrogenase and its mode of action Nitrogenase enzyme needed in biological N2 fixation is known to be sensitive to molecular oxygen. It comprises of a dimeric Fe protein (the dinitrogenase reductase) functioning as an electron carrier to the tetrameric MoFe protein (the dinitrogenase) which reduces molecular nitrogen to ammonia (Fig. 3.3). Both the enzymes viz. dinitrogenase, reductase, and dinitrogenase are highly oxygen sensitive. High oxidative stress is reported to cause proteolysis of nitrogenase, reduced enzyme synthesis, and leads to a shortage of respiratory substrates and reductants necessary for nitrogen fixation and assimilation. In filamentous heterocystous cyanobacteria, anaerobic environment causes diffusion barrier to gases due to multilayered wall of polysaccharides and glycolipids preventing the oxygen entry. They also possess elevated respiration rates to use up the defused O2 and no photosynthetic production of O2 occurs due to lack of

FIG. 3.3 Structure of nitrogenase.

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CHAPTER 3 Autotrophic nitrification in bacteria

photosystem II. Photosynthates transferred from vegetative cells to heterocysts ensure reductant supply of nitrogenase via Fe-S containing small proteins known as ferredoxin. ATP needed for N2 fixation is generated by oxidative phosphorylation in the heterocyst. The carbohydrates are imported by heterocysts (thick walled cells maintaining low oxygen tension inside) from the vegetative cells. Following substrates were found to be nitrogenase active by the heterocyst extract: maltose, glycogen, glucose, glucose 6-phosphate (G6P) and fructose, glyceraldehyde 3-phosphate (GAP), dihydroxyacetone phosphate (DAP), and fructose-1,6bisphosphate (FBP). Substrates such as phosphoenolpyruvate (PEP) and pyruvate (Pyr) were found to be inactive by the heterocyst extract. During night, activity of oxidative Polyenylphosphatidylcholine (PPC) and isocitrate dehydrogenase produces reductant for nitrogen and oxygen. NADPH thus generated donates electron by ferredoxin: NADP reductase (FNR) to a heterocyst-specific ferredoxin (FdxH) to nitrogenase (having two components viz. Fe protein and FeMo protein). Simultaneously, electrons are donated to the respiratory electron transport cycle (RET) via the same NADPH, thus producing the required ATP for the nitrogenase reaction. ATP is formed in the presence of light by cyclic photophosphorylation mediated by photosystem I (PSI). In some nonheterocystous cyanobacteria, N2 is fixed within the vegetative cells containing nitrogenase which can be deactivated in the presence of O2. In cyanobacteria, N2 fixation can take place in light under combined nitrogen-deficient conditions and in the presence of combined nitrogen source where the enzyme nitrogenase remains repressed. For example, cyanobacteria-rich rice fields decrease N2 losses via metabolization of the applied combined nitrogen forms. The metabolized combined N2 and biologically fixed N2 become available gradually through exudation and decomposition of these algae. A wide range of nitrogenase activity has been reported in cyanobacteria (Issa et al., 2014). In contrast to the known mechanism of artificial BNF to form NH3 via nitrogenase enzyme catalysis (Milton et al., 2017), another work involves microbial electrosynthesis in biological N2 fixation process (Fig. 3.4). In this method, inorganic C and N2 were fixed via electro-autotrophic metabolism that involves synthesis of new microbial biomass from air, water, and electricity. Microbial biofilm deposited on the solid electrodes is an essential source of electron donors during the process of C and N2 fixation. In another work, electro-enzymatic approach of BNF was tested on glassy carbon electrode coated with a mixture of dried and N-starved cells of A. variabilis and Nafion. Intracellular redox mediators such as Fd/NADPH mediated electron transfer from electrode to nitrogenase/nitrite reductase pool of A. variabilis leading to NH3 formation. Thus the new approach of electro-BNK (eBNF) involves complex biofilm formation on the electrode surface changing the electrochemical overpotentials of electrode reactions due to various biotic/abiotic factors affecting the extracellular electron transfer (EET) (Kalathil and Pant, 2016; Nevin et al., 2010). A series of complex interactions between various metabolic pathways offer a sustainable biomass production along with C and N2 fixation. This may help open a newer path in the field of sustainable and local on-demand N-fertilizer production.

3.5 Photosynthesis and nitrogen fixation

FIG. 3.4 Microbial electrosynthesis in biological N2 fixation process (DIET, direct intracellular electron transfer).

Thus e-BNF will help in finding alternate ways to enrich soil organic matter storage options thus preserving soil fertility in deserted localities (Rago et al., 2019).

3.5.3 Role of abiotic factors in BNF Biological nitrogen fixation is affected by several direct as well as indirect factors such as water content of soil, temperature, nitrogen concentration near the root zone, pH, plant nutritional status such as carbon, nitrogen, phosphorus, potassium levels that are inherently linked to the nitrogenase activity and nodule growth, and lastly genetic variations in the efficacy of nitrogen fixation. Soil temperature near the root zone is an important factor for maintaining the growth of nodule formation. It has been found that nitrogenase activity is highest around 12–35°C and reaches maximum at 20–25°C in most legumes. Legumes such as Trifolium vesiculosum Savi. and Glycine max (L.) Merr. show maximum nodulation in the temperature range 18°C and 32°C. Soil water deficit in the root zone is related to retardation in BNF (Goh and Bruce, 2005) by inhibiting the activity of nodules as it is also responsible for efficient percolation of dissolved gases through the root-soil interface. Carbon demand for nitrogen fixation is also a very important factor that determines the rate

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CHAPTER 3 Autotrophic nitrification in bacteria

of nodule growth, maintains a functional population of rhizobia, and helps in preparing amino acids produced from N fixation (Lichtfouse et al., 2011). For healthy growth of legume and nonlegume plants, inorganic mineral ions such as potassium (K); ions of nitrogen (N), calcium (Ca), sulfur (S); phosphorus (P); and trace amounts of other elements such as nickel (Ni), iron (Fe), manganese (Mn), chlorine (Cl), boron (B), zinc (Zn), molybdenum (Mo), and copper (Cu) are required to be supplied through soil. It has been found that N2 fixation is altered by the presence of mineral ions at various stages of symbiotic association including infection and nodule development, nodule function, and host plant growth. For example, adequate amount of phosphorus supply is required during nodule development and inadequate supply of the same may eventually lead to deficiency of nitrogen. Elements like calcium are required during early symbiotic events and boron is needed for nodule maturation. Copper, nickel, molybdenum, and iron serve the purpose of cofactors of various enzymes needed for nitrogen fixation including nitrogenase complex, ferredoxin, and hydrogenases (Weisany et al., 2013). Altogether the mineral nutrients play a huge role in various plant metabolic pathways such as cell wall synthesis, protein and nucleic acid synthesis, maintenance of osmotic concentration of cell sap, component of the chlorophyll molecule, electron transport systems, and enzymatic activity.

3.6 Conclusion and future aspect Nitrogen is an essential and limiting factor in plant growth. However, plants are unable to access the atmospheric dinitrogen gas directly and have developed efficient mechanisms that help them uptake the atmospheric nitrogen in the form of ammonium and nitrates from soil with the help of roots. This ability is seriously restricted to few of the plant species that has forced us toward the enhanced usage of nitrogenbased chemical fertilizers to majority of crop plants. Presently, 60% of the chemically synthesized nitrogen fertilizers are being used for cereals of which 50% is being truly taken up by plants and rest being leached to soil and nearby water sources causing pollution, eutrophication, and serious health hazards. The atmospheric nitrogen is converted to plant usable form ammonium by a highly conserved enzyme complex known as nitrogenase present in diazotrophs during the process of biological nitrogen fixation (BNF) (Franche et al., 2009). During BNF, the host plant is benefitted by a continuous supply of fixed nitrogen from the indwelling symbiotic partner which in return benefits by receiving nutrients from its host. Moreover, the plant develops nodules during the process of BNF which provides favorable conditions for protecting the nitrogenase complex from oxygen exposure. Thus rhizosphere associations between diazotrophs and plants allow organisms to cope with a variety of environmental stresses, spread across the biosphere and occupy new niches (Santi et al., 2013). Nitrogen fixing plants have the competitive advantage of receiving all or part of its requirements from BNF as compared to the nonnitrogen fixing plants owing to the endosymbiotic, endophytic, and associative interactions with various

References

microorganisms. Thus understanding the mechanism of symbiotic relationship between legume and rhizobium offers a better agronomic opportunity and also enables us with reduced application of N2 fertilizers. As majority of crop plants are nonlegumes, the knowledge of mechanism of action in BNF will help transfer the same endosymbiotic nitrogen fixation capacities to major nonlegume crops. Limited supply of nitrogen to crop plants affects the productivity of crop plants and successfully emphasizes the need for nitrogenous fertilizers for efficient growth of various crop plants. This leads to the development of various chemical pollution of soil owing to the excessive use of fertilizers. Hence, there is an urgent need for the reduction in usage of chemical fertilizers and increase in BNF by bacteria and archaea for conversion of atmospheric N2 to NH3, a form that can be used by plants. BNF is a symbiotic relationship between plants and microbes where plants provide the ecological niche and much needed carbon to the indwelling microbes within root nodules and microbe in turn provides the fixed nitrogen. BNF is strictly restricted to legume plants and there are significant research interests extending the possibility toward nonlegumes which produce bulk of human food. Thus research involving the knowledge of synthetic biology methods is needed for engineering the symbiotic relationships between nonlegume plants and nitrogen fixing microorganisms.

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CHAPTER

Omics: A revolutionary tool to study ammonia-oxidizing bacteria and their application in bioremediation

4

Hiren K. Patela, Priyanka D. Sheladiyaa, Rishee K. Kalariab, Vivek K. Diyoraa, and Nidhi P. Patela School of Science, P. P. Savani University, Surat, Gujarat, India bAspee Shakilam Biotechnology Institute, Navsari Agricultural University, Surat, Gujarat, India

a

4.1 Introduction Nitrification is a microbial technique that converts alkali to nitrate, and it is one of the primary tasks of the global nitrogen cycle, alongside nitrogen absorption and denitrification (Stein and Klotz, 2016). Nitrification is begun by the oxidation of smelling salts through particular gatherings of microscopic organisms AOB and archaea AOA. For around 125 years, alkali-oxidizing microorganisms should yield just nitrite from their digestion, which was free as the substrate for nitrite-oxidizing microbes to yield nitrate. In 2015, the principal reports of “comammox,” a bacterium that finishes smelling salt oxidation from alkali to nitrate, were distributed (van Kessel et al., 2015; Daims et al., 2015). This discovery has improved portraying and segregating the enzymology, guidelines, and individual spots of AOB, AOA, and comammox in common and designed conditions. Understandings of pathways and abiotic measures that cause nitrous oxide (N2O) are presented to explain why alkali oxidizers are linked to the solid ozone-depleting chemical N2O.

4.2 Chemolitho-autotrophic ammonia oxidation Chemolithotrophic nitrification is a two-step process that converts smelling salts to nitrite, which is then converted back to nitrate. These stages are acknowledged by two different gatherings of organic entities, the smelling salt-oxidizing microorganisms and the nitrite-oxidizing microbes (AOB and NOB), correspondingly (Warrington, 1878; Winogradsky, 1891). There are no distinguished autotrophic microscopic organisms that can catalyze the development of nitrate from alkali. AOB are chemolitho-autotrophs that require aerobes while a few animal categories Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00001-2 Copyright # 2022 Elsevier Inc. All rights reserved.

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation might be amazingly lenient toward low oxygen or anoxic conditions (Bodelier et al., 1996; Shah, 2021). In addition to their physiological consistency, Head et al. (1993) used 16S rRNA quality grouping data to show that all high-impact AOB isolates were limited to two developmentally distinct ancestries of the class proteobacteria. Similarly, all strains isolated from local and freshwater environments were assigned to a single (monophyletic) developmental group within the subclass of proteobacteria provided their ancestry was traced back to a single chemolithoautotrophic alkali oxidizing ancestor. The exclusive change of ammonia to nitrite through AOB includes a multiple reaction arranged through hyroxylamine, as shown in the subsequent two reactions. 2H+ + NH₃+2e
+ O₂ ! NH₂OH + H₂O


NH₂OH + H₂O ! HONO + 4e + 4H

(4.1)

+

(4.2)

(NH+4 )

It is ordinarily acknowledged that alkali (NH3) and not ammonium is utilized as a substrate, and the smelling salts/ammonium extent may consequently disturb development (Suzuki et al., 1974; Wood, 1986). A layer-bound, alkali monooxygenase (AMO) multisubunit compound catalyzes the main reaction; a periplasm-related chemical, hydroxylamine oxidoreductase (HAO), catalyzes the second. Two of the electrons made in the subsequent reaction are utilized as reward for the electron commitment of the primary reaction while the additional two go through an electron transaction chain to the terminal oxidase, thus producing a proton rationale power (Fig. 4.1). 2H+ + ½O + 2e
! H₂O

(4.3)

Reaction (4.4) gives the quantity reply: NH₃+1½O₂ ! HONO + H₂O

(4.4)

In spite of the inability to cultivate on carbon-based substrates, the hydrolysis of urea (Eq. 4.5) can be used through some species in place of a primary ammonia source. ðNH₂Þ₂C ¼ O + H₂O ! 2NH₃ + CO₂

(4.5)

In any event, two options occur for the anaerobic oxidation of smelling salts. The first has been established through ordinary oxygen-consuming alkali-oxidizing strains, which can diminish nitrite or nitrogen dioxide by hydroxylamine or smelling salts instead of an electron contributor (Schmidt and Bock, 1997; Bohn, 1992). While this methodology should uphold cell development, it might deliver enough energy to permit a presence under anaerobic conditions. A second anaerobic-autotrophic alkali oxidation instrument, the anammox method, has been credited to a bunch of life forms inside a profound stretching clade of the planctomycetales (Schmid et al., 2000; Strous et al., 1999). Devices conquered through these anaerobes yield N₂ from smelling salts and nitrate (Mulder et al., 1995) by the intermediates hydroxylamine and hydrazine. It isn’t yet recognized to what particular sum anaerobic alkali oxidation occurs in indigenous habitats.

4.3 Role of ammonia-oxidizing bacteria in nitrogen cycling

FIG. 4.1 Role of ammonia-oxidizing microorganisms in the nitrogen cycle.

4.3 Role of ammonia-oxidizing bacteria in nitrogen cycling The main natural sources of ecological smelling salts are nitrogen interest and mineralization, with the latter being most directly linked to autotrophic alkali oxidation. In aggregation, the amount of smelling salts saved either straight or circuitously for farming and modern occasions has intensely improved in current periods (Bobbink et al., 1998; Galloway, 1995; Grennfelt and Hultberg, 1986). The major phase in the oxidation of alkali to nitrate is smelling salt oxidation, which is hence dominating in the entire nitrogen cycle. In many game plans, smelling salt oxidation is regarded to be the rate-limiting stage for nitrification; for example, nitrite is rarely produced to accumulate in the environment (De Boer et al., 1990, 1992; El-Demerdash and Ottow, 1983; Prosser, 1989; Shah, 2020). Nitrification can emphatically or adversely upset nitrogen holding in a framework, dependent on the natural circumstances. Smelling salts are extremely unstable and consequently promptly lost from specific frameworks. In such cases, nitrification may work with nitrogen retaining through oxidizing smelling salts to a more modest sum of unstable nitrogen structures.

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation Nitrification can be integral to a net harm of nitrogen from ecological frameworks through three assorted contraptions. To begin with, the ultimate manufacture of nitrate can fuel denitrification, which is the main pathway for the expulsion of secure nitrogen from ecological plans. Second, the formation of alkali oxidation, nitrite, can attempt chemodenitrification, and it isn’t recognized how much this procedure disturbs nitrogen misfortunes. A third significant path to nitrogen misfortune is finished nitrate spillage from soil. While anionic NH+4 particles bind strongly to cationic soil iotas, they are far more mobile and hence quickly spill into adjacent channels. There are a lot of critical mixtures in the nitrogen cycle that have a standard of approximations for the work of nitrification. However, such estimations don’t imitate the nuances of nitrogen adjustments, along these lines giving a net outcome, best-case scenario. Because of nitrate absorption by plants, microbial throughput, immobilization, and many other factors, standing nitrate stages do not reveal nitrification phases (Stark and Hart, 1997).

4.4 Commercial significance and application of ammonia-oxidizing bacteria Organic smelling salt oxidation is the main stage in the disposal of nitrogen through the activity of waste. The treatment of nitrogenous waste has become a critical issue in ecological association as a result of the influenced rise in nitrogenous squanders due to the improvement of animal farming, human activities, and nitrogen-delivering industries. Smelling salts are the principle inorganic nitrogen wellspring of influent material, and nitrogen harm is characteristically because of the improvements of nitrification, shadowed through denitrification. The removal of nitrogen from trash the board is incredibly important from an ecological standpoint, as the declaration of rough waste might result in astonishing eutrophication of the climate, particularly in heavily populated areas. Indeed, even in situations where the board doesn’t integral to viable denitrification, nitrification assists with staying away from ecological contamination with theoretically harmful alkali salts (Painter, 1986). Squander treatment is through distant the most huge biotechnological use of AOB. The administration of such squanders typically happens specifically reactors of numerous proposal and assurance, whose steadfastness and cost-viability are of main anxiety to assembling and shared wellbeing. The broad explicitness of the alkali monooxygenase compound, common to all AOB, habitually allows the cooxidation of many unruly aliphatic, fragrant, and halogenated particles (Hooper et al., 1997). A preliminary oxidation stage is frequently the rate-restricting variation in the microbiological filth of several boisterous hydrocarbons, and AOB may operate with bioremediation systems along these lines. Limited evidence from pure culture considers suggests that AOB may play a role in methane oxidation (Steudler et al., 1996), therefore easing to evacuation of this critical ozone harming substance, yet results from the field still can’t seem to demonstration important methane ingesting through AOB (Bodelier and Frenzel, 1999; Bodelier et al., 2000).

4.5 Nitrification and ammonia-oxidizing bacteria

AOB may likewise play a vital part in biofilter schemes, which normally contain microbial gatherings that are used in influent substrates to endure a comparatively continuous biomass and produce waste liberated from input debasements. In such scenarios, AOB may collaborate closely with other gathering members to burn through noxious structures such as smelling salts or CO2, reuse supplements, and ensure mass reliability. Biofilter plans have been utilized for a variety of options to date, usually connecting the end of smells related with squander treatment and fertilizing the soil (Bohn, 1992) yet additionally for judgments as outlandish as given that long haul sifting estimations reasonable for operated spacelab ( Joshi et al., 2000).

4.5 Difficulties associated with nitrification and ammonia-oxidizing bacteria Nitrification can affect a few conditions, notably those offered to raise nitrogen affidavit levels, in order to express scorn for the existence of a characteristic approach (Painter, 1986; Van Breemen and Van Dijk, 1988). One of the limiting issues of nitrification is nitrate contamination of groundwater and freshwater, attributable to nitrate spillage from agrarian locales. The use of nitrate by humans and animals can pose serious health hazards, and government restrictions restrict the quantity of nitrate allowed in drinking water. Nitrate penetration in freshwater may likewise be fundamental to the eutrophication of such conditions, increasing the development of some phototrophic and heterotrophic lifeforms. This can be key to diminished biodiversity and the development of anoxic conditions. The high production of contemporary agricultural practices is reliant on the usage of nitrogen-rich manure, which can be created artificially or by repurposing natural squanders, such as animal excreta. Nitrogen deficiency in such manures, whether before or after application, compromises their viability. Nitrification can be key to nitrogen misfortune through two significant components: (a) aggregate nitrogen leakage due to the change of smelling salts to nitrate, or (b) associating nitrification by denitrification activities (Abbassi and Adams, 1998). Blemished denitrification is likewise vital to assembling the ozone-exhausting gas NO or the ozone-depleting substance N2O. AOB can likewise yield these destructive gases either at low levels as side effects of regular nitrification (Zart et al., 2000) or at higher levels (Bremner and Blackmer, 1978; Goreau et al., 1980; Lipschultz et al., 1981; Poth and Focht, 1985) through incomplete denitrification developments below reduced oxygen conditions (Colliver and Stephenson, 2000). Numerous collections of microorganisms contribute to the development of these follow gases (Bouwman, 1990); however, the relative guidelines of assorted microbial gatherings aren’t yet completely comprehended. Research suggest that the amount of AOB in NO and N2O is small, yet this can’t be set up in situ. Although approaches have been made to differentiate among the various wellsprings of these gases, for example, denitrifying (Kester et al., 1996), it is unknown as to what sum AOB gives to the universal construction of these gases. However, it is now

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation established that the production of N2O under difficult ventilation varies substantially throughout AOB groups, highlighting the importance of considering aspects such as population size and the construction of ecological AOBs ( Jiang and Bakken, 2000). The smelling salts technique leads to a net fermentation of the climate due to oxidation indicators. When alkali explanation and nitrification levels are high, the ecological pH may be lowered. Where alkali explanation and nitrification levels are high, this may bring down the ecological pH (Beiderbeck et al., 1996). The fermentation of woodland soils can have a harmful consequence on tree well-being, and extraordinary degrees of nitrification may increase difficulties, including corrosive downpour. Likewise, Nitrogen Trans forms is the key to an improved proton burden, which can be major issue of metals like aluminum, which can guarantee to root harm and timberland debilitating. Such challenges can be impaired through nitrification in modifying the overall pools of NH4 + and NO3
, the most public nitrogen structures taken up by plants. Some tree species have a strong preference for taking up NH4 + over, and nitrification may disrupt seed-lethargy breaking, germination, seedling foundation, and development in terms (Kronzuker et al., 1997). In distinction, many weed classes show an inclination for NO3
and may prevail under conditions of increased nitrification. AOB may subsequently influence vegetation series by the near bounty of inorganic N sources (Kronzuker et al., 1997). AOB movement has additionally been a concern in the disintegration of characteristic stones, authentic landmarks, and building resources through creating raised levels of corrosive nitrous (Meincke et al., 1989; Spieck et al., 1992). The transition from chlorites to chloramines in the disinfection of drinking-water storage and distribution systems has created a problem for AOB in maintaining the transportability of decontaminated water (Baribeau et al., 2000). Chloramines infiltrate biofilms more effectively than typical chlorites, yet clogging causes the declaration of free alkali, resulting in the compelling development of AOB. Nitrite produced through this action responds through remaining choramine residuals, approving the quick development of extra bacterial populaces, including coliform microbes. Nitrification and the development of coliforms bring about an extreme drop in water quality, with the end goal that water needs be eliminated and the framework desterilized by additional activity, comprising a fundamental monetary weight on the water business.

4.6 Isolation of ammonia-oxidizing bacteria from the environment AOB start in most oxygen-consuming conditions where alkali is possible through the mineralization of natural matter or anthropogenic nitrogen bases, such as compost and waste. As a rule, carbon core, CO2, should be limited to alkali oxidizer development and development. AOB are generally in soils, freshwater, and marine conditions and have been separated from these and extra territories. While AOB are typically viewed as vigorous, they have likewise been separated or upgraded from low-oxygen conditions such as pungent water (Smorczewski and Schmidt, 1991),

4.7 Cultivation of new ammonia oxidizers

generally anoxic soil and silt layers, and the subsurface portions of building assets (Bock and Sand, 1993). Certain confined classes will in general be limited to specific conditions. AOB unadulterated societies are characteristically achieved through picking settlements from a strong medium (Ford et al., 1980) or through utilizing weakening techniques in fluid culture (Aakra et al., 1999; Schmidt and Belser, 1982). The analytical medium used should be free of natural carbon bases, contain inhibitors of heterotrophic organic entities, a source of smelling salts, and important basics (MacDonald and Spokes, 1980). AOB are disliked for their sluggish growth and poor maximum growth yield, making isolation and support in pure culture risky and tedious (Schmidt and Belser, 1982). In like manner, the way of life media may decision for explicit strains that may not unavoidably signify the most bountiful populaces present ( Juretschko et al., 1998; Stephen et al., 1996; Wagner et al., 1993). Likewise, similarly as with extra microorganisms, the negligible part of practical cells in a natural example that is truly agreeable to research facility culture conditions might be minuscule (Amann et al., 1995; Wagner et al., 1993). Culture-subordinate strategies such as specific plating (Ford et al., 1980) and the most extreme conceivable number (MPN) strategy (Alexander, 1982) have been utilized for the rundown of AOB; however, such methods should underestimate genuine cell numbers. In adding to medium selectivity and inclination, MPN underestimation may likewise come from insufficient postponement of cells from strong substrates in the natural example or scattering of herds and miniature settlements (Aakra et al., 2000). Imprecisions may also result from cell debilitation caused by difficult disruption methods or osmotic stun, as well as the conceivable reliance on between-species or interspecies connections for development. The investigation of unadulterated societies of AOB has clearly increased our comprehension of their physiology; however, the physiological assets of secluded strains regularly can’t explain the nitrification activities found in the conditions from which they were improved. Two fundamental possibilities can be engaged to clarify this division: (a) Laboratory conditions utilized for physiological experiments are not illustrative of the in situ natural conditions. AOB may require other organisms for monitoring, substrate supplies, intra- or interspecific sign atoms, or specific climate features in order to have the action sex periential in the environment. (b) The life forms naturally separated in unadulterated culture are not typical of the main physiological kinds of AOB in these conditions. Clearly, the sole reliance on unadulterated culture investigations of AOB is deficient to characterize their numbers, variety, and exercises.

4.7 Cultivation of new ammonia oxidizers Our thoughts on alkali oxidizers and the challenges of their physiology have been intensely rearranged through examining axenic societies. Sergei depicted the major isolation of a smelling salts oxidizer, Nitrosomonas European, in 1890

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation (Winogradsky, 1890). Various strains of AOB, implying two groups of Proteobacteria (Beta proteobacteria and Gamma proteobacteria), AOA in the subphylum Thaumarchaeota, and comammox microbes in the Nitrospirae phylum, have remained passed on into culture from a variety of biological systems including soils, estuaries, marine associations, underground aquifers, freshwater, residue, aquaria, wastewater treatment offices, heated water pipe biofilms, drinking water plans, and various others (Stein and Nicol, 2018; Kits et al., 2017). Isolation of smelling salts oxidizers is not an insignificant task because they oppose microorganisms’ tight groups that purge metabolic intermediates, protect them from oxidative pressing factor (Sedlacek et al., 2016), or achieve mutual taking care of capacities, such as cyanate conversion to smelling salts via related nitriteoxidizers (Palatinszky et al., 2015). Alkali oxidizers have long aging times and are sensitive to environmental factors such as substrate concentration, light, pH, temperature, and oxygen (Bollmann et al., 2011). Alkali oxidizers are correspondingly exceptionally fragile to the delicate oxygen class. The commitment of pyruvate in the development mode of AOA was originally believed to withstand mixotrophic digestion; however in was found to detoxify hydrogen peroxide (Kim et al., 2016). Due to the encounters by axenic development, the standard of accessible genome successions for AOB, AOA, and comammox microbes are from metagenomic information or improvement societies to some degree than from segregates.

4.8 Genomics and metabolic models Through the revelation of another microbiological confine, genomic models have produced a standard to contain a genome succession. Admittance to genome succession proof permits portrayal and difference of animating physiological, controlling, and transformative geologies of microorganisms in adding to a stage for building extrapolative genome-scale metabolic models. Genomic list on its individual contributions through creating speculations for invigorating physiologies, for example, corrosive resilience of the AOA detach, “Candidatus Nitrosotalea devanaterra” (Lehtovirta-Morley et al., 2016) and related strains (Herbold et al., 2017). Differentiation of four Ca. Nitrosotalea genome game plans with 23 extra AOA genomes uncovered 743 shared center proteins (Herbold et al., 2017), which is less than the 860 mutual center proteins depicted in an earlier AOA genome contrast (Kerou et al., 2016). All AOA genomes encode the 3-hydroxyproionate/ 4-hydroxybutyrate CO2 fixation pathway, predominant carbon pathways, and proteins for smelling salt oxidation. Nitrosotalea genomes contain an Na+/solute symporter, metal carriers, and a chaperone exact which are proline-rich proteins (Herbold et al., 2017). Soil AOA encodes exclusive highlights, for example, envelope adjustments for biofilm creation, polysaccharide production, and cell-cell grip (Kerou et al., 2016). A previous analysis of N. devanaterra genome predicted 51 potential traits, including low pH substrate transport with high empathy, film

4.9 Terminology of environmental proteomics

impermeability, and different cation carriers (Lehtovirta-Morley et al., 2016); yet, the more vigorous relative examination tracked down that everything except 10 of these qualities are in AOA genomes of nonacidophilus (Herbold et al., 2017). These investigations use the basic for partner across as many genome groups as possible to identify special quality sets that address certain capacities, but they also suggest that quality list alone is an inaccurate indication of distinct routes and capacities. Genome-scale metabolic organization demonstrates collecting layers of practical data from a microorganism into a numerical portrayal. When iteratively complex with trial information, the model can exactly forecast universal metabolic answers of an organism before validation through wet-lab tests, along these lines spreading practical data far outside the genome grouping alone. Metabolic organization models have been utilized to anticipate N2O improvement from nitric oxide (NO)-delivering reactions in AOB consortia through nitrite oxidizing microorganisms (Perez-Garcia et al., 2016; Mellbye et al., 2018) and to characterize the transformation of N. European to anoxic-oxic changes (Yu et al., 2018). The most recent research concluded that exposing N. European to repetitive anoxicoxic cycles resulted in a decrease in N2O formation and that variations in nitrogen oxide metabolite transitions were linked to changes in protein, rather than record, levels (Yu et al., 2018). The hint of this investigation is that N2O discharges from AOB are likely exaggerated in wastewater activity tasks that habitually use anoxic-oxic cycling. With quickly cumulative availability to genome groupings, omics, and physiological data on smelling salt oxidizing separates, genome-scale metabolic organization is rapidly turning into an essential apparatus for deciding residual inquiries on alkali oxidizer digestion, whole genome guidelines, and control of delicate metabolites, similar to NO and N2O.

4.9 Terminology of environmental proteomics Wilmes and Bond (Wilmes et al., 2008) characterized metaproteomics as “the significant portrayal of the entire protein supplement of the natural microbiota at a predefined point in a period.” Briefly, a speedy turn of events and multifold accommodation of high-throughput omics information have been coordinated to numerous new qualities including ecological proteomics, metaproteomics, local area proteomics, or local area proteogenomics. Natural proteomics should be a conventional term, essentially depicting the proteome examination of ecological examples. Metaproteomics incorporates investigations of incredibly complex organic plans that don’t allow the transmission of an enormous number of proteins to the correct species inside the phylotypes. General observation of proteomics reveals that the majority of detected proteins may be linked to specific persons in the region; thus far, such analyses have primarily focused on low- or medium-complexity circumstances. The word proteogenomics, which basically characterizes the use of proteomics to enhance quality, also depicts the valuation of the strain or species varieties and the

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation transformative improvement of the genomics to specific conditions (VerBerkmoes et al., 2009). Moreover, proteogenomics adds to the insight of the genuine quality capacity through connecting perceived protein data to the DNA level.

4.10 Microbial culture proteomic studies techniques Much work has been spent in investigating microbial strategies of late. Proteomic considerations have been gathered, and a few standard procedures have been established, including test selection, protein separation, partition and further fractionation, mass spectrometry evaluation, dataset search, and long-term comprehension of data. Attributable to the presence of different ecological examples, when working with marine and freshwater just like surface examples (Habicht et al., 2011), remarkable techniques are set up for either test handling or protein extraction. Two strategies have been produced for protein partition and recognizable proof. The first is the methodology centered around gel. Blended proteins are typically segregated utilizing possibly one-dimensional or two-dimensional electrophoresis with polyacrylamide gel (2D PAGE). The objective protein spots or groups are then extracted and proteins are processed with trypsin or different chemicals into peptides. Following that, the peptides are subjected to mass spectrometry (MS) or couple (MS/MS) concentration, dataset sweep, and bioinformatic analysis (Wang et al., 2014). The alternative technique is a liquid chromatography (LC)-based interaction, in which the entire proteome is treated with proteases without first partitioning gel proteins into a more complicated peptide combination. Also, using the strong cation trade chromatography or microcapillary turnaround stage, the resultant peptides are secluded. All in all, by utilizing fluid chromatography combined with MS/MS (LC-MS/MS), the separated peptides are analyzed. For protein recognizable proof and bioinformatics examination, the produced MS information is deciphered. The subsequent strategy beats the deficiencies of the gel-based methodology and significantly improves proteome inclusion compared to the gel-based methodology, empowering a large number of proteins to be recognized through high-throughput screening inside a restricted timeframe (VerBerkmoes et al., 2009). It also makes it conceivable to distinguish insoluble film proteins (Wu and Yates, 2003). Subsequently, the LC-based strategy has been the standard of proteomic concentrates in the microbial world, despite the fact that it actually experiences reproducibility, dynamic range, and shortage of datasets. The reliability of metaproteomic investigation is critical in determining if the difference in protein articulation in the microbial population is earth-sized or the product of device failures. Specialized repeatability will be about half across three sets and above 67% between recreations utilizing a comparable MS stage. Specialized reproducibility will generally be near half across sets of three and more than 67% across recreations using a similar MS stage (Elias et al., 2005). Further imitates can build protein acknowledgment, yet reproducibility, especially for organic rehashes, may deteriorate. Notwithstanding a subjective examination in proteomics, the amazingly reproducible and exact presentation of a wide

4.10 Microbial culture proteomic studies techniques

range of quantitative information is valuable, especially for near and quantitative proteomics: their key goal is to assess the varieties in protein articulation across different natural states (for example, guideline versus treatment, stable versus illness, human genotype versus wild t) (e.g., supplement and saltiness angles). Different marking techniques have as of late been created for proteomics, for example, stable isotope naming using amino acids in cell culture, pair mass labels, stable isotopenamed peptides, isotope weakening, isotope-coded fondness labels, and relative and supreme evaluation isobaric labels. However, most name-based assessment procedures are reduced in difficult example preparation, protein improvement, and inadequate stamping, as well as in the measurement of tests. A name-free quantitative proteomic technique has advanced with the production of sufficient computational devices, empowering the profiling of an expansive protein scale with the adaptability of a few distinct examinations. With the development of suitable computational equipment, a quantitative proteomic approach has progressed, allowing for the profile of a large number of proteins with the flexibility of a few different tests. Also, on account of its nonmarking usefulness, it is practical. As an outcome, the mark-free methodology jogged on MS has gotten more normal and has become the overwhelming type of metaproteomics science (Fig. 4.2).

FIG. 4.2 Overview of the functional process of metaproteomics.

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation

4.11 Potential applications of environmental proteomics In their ordinary living space, organisms spend much of the time confronting fast and correctional changes of natural cutoff points such as temperature, moistness, hunters, and supplement accessibility. A shared methodology of microorganisms to defeat these investigations is a change of their protein appearance profiles. Likewise, a basic investigation of the different qualities and their guidelines isn’t sufficient to completely understand microbial transformation approaches. Postgenomics examinations with transcriptomics and proteomics are promptly expected to examine the physiology of the complex microbial relationship at a subatomic level. The ecological proteomics beforehand contains a genuine “money box” of information going from the basic protein listing to relative and quantitative proteomics, investigations of protein limitations, identification of the posttranslational changes that may influence the usefulness of the protein, an investigation of the protein-protein associations, and assurance of amino corrosive successions and their genotypes. Consequently, the most likely use of the higher advances in microbial nature is bountiful and contains a record of new useful qualities, documentation of totally new synergist catalysts or complete metabolic pathways, a record of the practical bio pointer to screen elements, and maintainability of the ecological class (Maron et al., 2007). Moreover, improvement and serious use of the entire arrangement of the “omics” advances will allow us to reevaluate the microbial nature by connecting hereditary and useful verities in a microbial local area, including the ordered and utilitarian variety to the biological system steadiness.

4.12 Enzymology of ammonia-oxidation Because of the toxicity of its intermediates and products, smelling salt oxidation is a very particular digestion across prokaryotes, as is the requirement for specific electron transporters to ferry reductant to the Quinone pool to generate proton thinking process power and ATP. Protected smelling salt monooxygenase (AMO) compounds are used by AOB, AOA, and comammox bacteria to convert alkali to hydroxylamine (Pjevac et al., 2017; Alves et al., 2018). All smelling salt oxidizers correspondingly oxidize hydroxylamine and yield nitrite; however, the enzymology for this piece of the pathway is partly unsettled. In 2017, Caranto et al. determined that the chemical hydroxylamine dehydrogenase (HAO) of AOB oxidizes hydroxylamine to NO marginally more than nitrite (Caranto and Lancaster, 2017). This finding transformed the long-standing marvel of HAO’s direct oxidation of hydroxylamine to nitrite into a practical test to identify nitric oxide oxidoreductases (NOO) (Fig. 4.1). In this manner, two NOO candidates have been suggested for AOB: nitrosocyanin encoded through ncyA, and working copper covering nitrite reductase encoded through nirK (Lancaster et al., 2018). However, the oligotrophic segregate, Nitrosomonas sp. Is79, has a ncyA quality that is exclusive to AOB and begins in all AOB genome arrangements suitably far away (Bollmann et al., 2013). Proteomic investigation of three

4.13 Ammonia-oxidizers in the environment and production of N2O

AOB species presented great articulation of NcyA alongside extra compounds in the alkali oxidation pathway, connoting its part as a vital player (Zorz et al., 2018). Similarly, NirK appearance in AOB isn’t constitutive, and is controlled through openness to high smelling salts (Zorz et al., 2018), NO (Beaumont et al., 2004), and anoxia. NirK is also unconcerned about some AOB genomes (Kozlowski et al., 2016a,b,c). Similarly, in Nitrosomonas europaea, the nirK gene may be removed with no negative consequences for growth, indicating that it is superfluous in this model AOB strain (Kozlowski et al., 2014).

4.13 Ammonia-oxidizers in the environment and production of N2O Since the early 1980s, AOBs have been known as important producers of the greenhouse gas N2O. Undiminished fertilizer use, hypoxia in aquatic and coastal ecosystems subject to nitrate runoff, and mid-20th century wastewater management systems stimulated N2O production through resident AOB through an enzymatic process called “nitrifier denitrification” or from biotic and abiotic transformations of their metabolic intermediates. The main stream of cultivated AOBs encodes and expresses NirK and nitric oxide reductase enzymes (NorB/Y), which in a model should decrease nitrite to NO and N2O. But, the individual NorB enzyme was found important to nitrifier denitrification activity in a gene knock-out study of N. europaea (Kozlowski et al., 2014). Furthermore, oligotrophic strains that lack NorB cannot perform nitrifier denitrification (Kozlowski et al., 2016a,b,c). Also the strain Nitrosomonas lacks NirK, so it retains nitrifier denitrification activity (Kozlowski et al., 2016a,b,c). These studies specify that unidentified nitrite reductases are yet to be discovered in the AOB and, yet again, the physiological role of NirK remains indeterminate. Other mechanisms of N2O generation by AOB include the anaerobic oxidation of hydroxylamine through the enzyme cytochrome P460 (Caranto et al., 2016) and abiotic reactions of metabolic intermediatesNO, hydroxylamine, and nitrite by each supplement, with media components, and with metals (Kozlowski et al., 2016a,b,c; Liu et al., 2017; Terada et al., 2017; Frame et al., 2017; Zhu-Barker et al., 2015). AOAs do not have the physiological capacity to do nitrifier denitrification (Kozlowski et al., 2016a,b,c). Abiotic responses of metabolic intermediates harvest N2O, which can be considerable where AOAs are rich and active, such as marine ecosystems (Trimmer et al., 2016). Abiotic formation of N2O from intermediates of the AOA metabolism has been long-established in soils, and is favored under conditions of low ammonium supply (Hink et al., 2018a,b). These studies established that the manufacture of N2O through nitrifier denitrification in AOB is significantly better than abiotic N2O manufacture through AOA, and is additionally favored in soils with a high ammonium supply (Hink et al., 2017, 2018a,b). Studies are ongoing to define whether comammox bacteria contribute to N2O emissions; however their strict oligotrophic lifestyle would designate that they produce very little, if any.

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CHAPTER 4 Proteomics: A novel tool to study Ammonia-oxidizing bacteria for bioremediation It should be noted, however, that the clarification of exact sources of N2O is an ongoing methodological challenge, complicated through the lack of interlaboratory calibration of standards and instrumentation. Studies attempting to define the microbial populations active in bioreactors are primarily motivated by a desire to better understand the population structures that are typical of entirely common sense and basic frameworks. As a result, it is typical for reactor failure to be predicted before the recognized shifting local area is really hurt or cleaned out of the framework. These approaches may target the whole bacterial local area (Almeida et al., 1998; Eichner et al., 1999) or specific modules such as the β-AOBs. The last technique may have benefits attributable to practical repetition in the more extensive microbial local area, to such an extent that significant changes in local area construction may not impact useful reactor solidness (Fernandez et al., 2000). In any case, displaying such obviously basic conditions can be exceptionally diverse (Painter, 1986), and one should perceive that unequivocal nitrifying lineages might be especially appropriate for a specific reactor arrangement and waste info (Suwa et al. 1997). In this way, the protection of explicit nitrifying populaces might be indispensable to reactor well-being, and difficulties to arrange and display these networks are of extensive importance. Similarly, if reactor action is to be improved through bioaugmentation by nitrifying microscopic organisms, it follows that the inoculum should contain species that are fit to the reactor climate. The comparison must be valid for soil frameworks, assuming that soil attributes control the equilibrium of β-AOB species types. While the essential piece of β-AOBs in biotechnology is in nitrogen changes for wastewater, with soil-based wastewater treatment (rhizoremediation), they additionally hold significant potential for the corruption of numerous unmanageable waste synthetic compounds.

4.14 Remediation of recalcitrant pollutants Because of alkali monooxygenase’s broad substrate specificity, AOBs can potentially follow up on a variety of substrates, some of which can be extremely harmful and wild in the environment. Anthropogenic toxins oxidized through β-AOBs contain trichloroethylene, chloroethane, benzene, various halogenated aliphatics, dimethylether, and characteristic and engineered estrogens (Fernandez et al., 2000). The oxidation of several of these substrates, which are generally regarded inhibitors of autotrophic alkali oxidation, does not provide energy to AOB. Likewise, the redox of the climate can be risky to the bioavailability of numerous contaminations with heavy metals, just as to the movement of extra microorganisms that assume huge parts in the remediation of such mixtures (Painter, 1986). In this manner, the activity of AOB may have both immediate and less-direct effects on ecological clean-up components. The well-being and species game plan of β-AOBs in dirtied conditions and bioreactors can be a significant part of microbial biology.

References

However, lab-based examinations in this space have all decided on N. europaea, a nitrifier not bountiful in various natural settings, and that ecological media disturb the pace of cooxidation of toxins (Painter, 1986).

4.15 Conclusion Our thoughts of smelling salt oxidizing microorganisms have expanded in recent years through the discovery and development of novel organisms; the quick propagation of genome succession and useful omics information; concentrated biochemical work that both delayed the usefulness of recognized proteins and moved us to characterize novel chemicals; and environment examinations that uncovered interconnections of alkali oxidizers by all other and by accomplice microorganisms. We currently distinguish that both enzymology and abiotic associations to air N2O through smelling salt oxidizers and those assorted systems exist. We additionally recognize that NO is the fundamental middle of the road in all smelling salt oxidation pathways. Leftover inquiries incorporate the identification of NOO proteins; the different parts of NirK and cupredoxins underway of NO, the capacity for comammox microorganisms to yield N2O, and genomic-administrative highlights that allow particular capacities and spot inclination among AOB, AOA, and comammox microscopic organisms. The goal of these inquiries will allow a true extenuation for compost overemployment and N2O creation and joining of ammonia oxidizers for modern and ecological judgments. Many reveals remain as significant features of these fascinating microbes, and their selective digesting remains unknown.

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Ambreen Ashara, Muhammad Muneebb, Zeeshan Ahmad Bhuttac, and Muhammad Shoaibd Department of Chemistry, University of Agriculture, Faisalabad, Pakistan bDepartment of Pathology, University of Agriculture, Faisalabad, Pakistan cLaboratory of Biochemistry and Immunology, College of Veterinary Medicine, Chungbuk National University, Cheongju, Chungbuk, Republic of Korea dInstitute of Microbiology, University of Agriculture, Faisalabad, Pakistan a

5.1 Introduction Nitrification is a two-step process that involves the conversion of ammonia to nitrite and then nitrite to nitrate in the nitrogen (N) cycle (Prosser, 1989). Because of its biological relevance in the global N cycle and environmental consequences, ammonia oxidation, the first and rate-limiting phase of nitrification, has been extensively researched (Kowalchuk and Stephen, 2001). Ammonia-oxidizing bacteria (AOB) have been thought to be the major engine of ammonia oxidation for over a century. The finding of ammonia-oxidizing archaea (AOA) has lately cast doubt on this belief (Venter et al., 2004; K€ onneke et al., 2005; Treusch et al., 2005). The first research of AOA’s potential relevance in soils was conducted in a variety of European soils, and it found that AOA were clearly dominating among ammonia oxidizers (Leininger et al., 2006). In a variety of aquatic and terrestrial settings, further investigations have verified the broad dispersion of AOA and their numerical superiority over AOB (e.g., He et al., 2007; Dang et al., 2008). In recent years, the relevance of AOA in ammonia oxidation has garnered a lot of research interest (Nicol and Schleper, 2006; Prosser and Nicol, 2008). To explain the contribution of AOA to ammonia oxidation and therefore comprehend the connection between AOA abundance, diversity, and ecosystem function, soil molecular biology methods such as DNA-based stable isotope probing (DNA-SIP) and transcription analysis have been used (Tourna et al., 2008; Jia and Conrad, 2009; Offre et al., 2009). Environmental factors may shape specific niches of AOA and their contribution to nitrification can be altered in low nutrient, low pH, and sulfide comprising environments (Erguder et al., 2009). The activity and role of AOA in nitrification in low-N input and low pH soils have been verified by SIP studies (Zhang et al., 2010, 2012). The relative significance of AOB and AOA in ammonia oxidation, however, is yet unknown, and their respective contributions to Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00018-8 Copyright # 2022 Elsevier Inc. All rights reserved.

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nitrification may vary depending on soil conditions. China has a vast agricultural terrain with a wide range of soil types and pH levels. Due to substantial increases in the usage of ammonia-based N fertilizers and their related environmental concerns, nitrification research has gotten a lot of interest since the 1980s. The discovery of archaeal ammonia oxidation sparked a surge in interest in Chinese soil ammonia oxidizers. Ammonia monooxygenase (AMO), a multimeric enzyme, transforms ammonia to hydroxylamine, resulting in biological ammonia oxidation. The amoA gene, which has variations in bacteria and archaea, encodes the AMO enzyme’s alpha (amoA) component. In the last five years, new insights into the diversity and distribution of AOB and AOA targeting the amoA genes in various Chinese soils have been rapidly achieved due to the availability of culture-independent molecular ecology techniques, such as denaturing gradient gel electrophoresis (DGGE), terminal restriction fragment length polymorphism (T-RFLP), cloning, and quantitative PCR in China. For example, the quantity and composition of AOB and AOA in acidic red soils (pH 3.7–6.0) were initially studied at a long-term field experiment site in Qiyang (Hunan province, southern China), which had been fertilized continuously for 16 years (He et al., 2007). The copy counts of AOA amoA genes were larger than those of AOB amoA genes in soil samples from all eight fertilization treatments, with AOA to AOB ratios ranging from 1 to 12 (He et al., 2007). These findings corroborated Leininger et al. (2006) finding that AOA outnumbers AOB in agricultural soils. Several studies of alkaline agricultural soil with pH ranging from 8.3 to 8.7 revealed a similar pattern (Shen et al., 2008). Chinese were the first to find AOB and AOA in agricultural soils with different pH levels. Following that, AOB and AOA comparison investigations in China covered a wide geographic range of agricultural soils with varying management regimes (Chen et al., 2008; Wang et al., 2009; Ying et al., 2010; Yao et al., 2011; Zhang et al., 2012). With the exception of two long-term fertilization treatments, where the ratios of AOA to AOB amoA genes were 0.25 and 0.39. A general tendency identified from this research is that the archaeal amoA genes are more common than bacterial amoA genes in most of the soils examined (Wu et al., 2011). China is one of the world’s major rice producers, with roughly a quarter of the world’s rice paddy growing area under its control. Because of the frequent oxic/ anoxic alternations, the paddy environment provides a typical wetland system for studying biogeochemistry and microbiological activities. The porous tissue of rice transports oxygen from the atmosphere into the roots, resulting in an oxic environment capable of sustaining nitrification inside the rice rhizosphere (Arth et al., 1998). Despite the fact that paddy soil nitrification has been extensively researched, the fundamental microbial contributions to the nitrification process remain slightly known. Chen et al. (2008) used microcosm studies to demonstrate the existence of abundant AOA in Chinese paddy soils. This study found that AOA was more abundant than AOB in bulk and rhizosphere soils with and without rice plants, with AOA to AOB ratios ranging from 1.2 to 69.3. A research on acidic field paddy soils with varied

5.2 Emission of nitrous oxide

long-term fertilization treatments (Chen et al., 2011) or on diverse paddy soil types further supported the numerical preponderance of AOA over AOB in paddy soils (Chen et al., 2010). Rice farming increased AOA abundance but not AOB abundance, as measured by the number of copies of the amoA gene (Chen et al., 2008). Empirical data reveals that AOA may tolerate a broad range of oxygen levels (Erguder et al., 2009), and bioinformatic study of AOA genomes supports the possibility of mixotrophic metabolism, i.e., switching between chemoautotrophy (nitrification) and heterotrophic rice root exudate growth (Walker et al., 2010; Pester et al., 2011). The isolation of the first ammonia-oxidizing archaeon (AOA), Nitrosopumilus maritimus SMC1 of the until then mysterious group I.1a archaea, was published shortly after these findings from metagenomics (K€onneke et al., 2005). Members of this lineage have been shown to account for 20%–30% of all marine microorganisms in open ocean and coastal waters. Additional N. maritimus strains have been produced in enrichment cultures in recent years (Park et al., 2010; Wuchter et al., 2006). Furthermore, it was discovered that the uncultivated marine sponge symbiont “Candidatus Cenarchaeum symbiosum” encodes genes required for NH3 oxidation and so became known as an AOA (Hallam et al., 2006). However, until more evidence is available, this organism should be regarded an amoA-encoding archaeon because it has not been proven to catalyze the oxidation of NH3. The genome of group I.1a AOA “Ca. Nitrosoarchaeum limnia” SFB1 was virtually entirely rebuilt using a combination of metagenomics and single-cell sequencing after it was enriched from a low-salinity sediment (Blainey et al., 2011). Enrichment cultures from agricultural soil and estuary sediment were used to acquire the genomes of “Ca. Nitrosoarchaeum koreensis” and “Ca. Nitrosopumilus salaria,” respectively (Mosier et al., 2012). Recently, novel (as-yet-unnamed) AOA species were enriched from freshwater sediment, adding to our understanding of this archaeal lineage’s environmental range (Bollmann et al., 2014).

5.2 Emission of nitrous oxide Nitrous oxide (N2O) is generated as a result of classical and nitrifier denitrification—the reduction of NO2/NO3 to N2 in multiple steps—and partially escapes into the gas phase. Because of its long atmospheric lifespan, N2OH has a 310-fold greater greenhouse warming potential than carbon dioxide (CO2) and accounts for 5%–7% of the known greenhouse effect, making it the third most significant greenhouse gas (Braker and Conrad, 2011). N2O has a negative impact on the earth’s ozone (O3) layer due to its interaction with atomic oxygen (Ravishankara et al., 2009). AOB and perhaps AOA are thought to be responsible for 70% of worldwide N2O emissions when used in conjunction with classical denitrifiers ( Jung et al., 2011).

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5.2.1 Potential sources There are two different sources of this gas in AOB. The chemical breakdown of NH2OH produces small quantities of NO and N2O (Braker and Conrad, 2011). Furthermore, when oxygen is scarce, AOB easily accepts NO2 as a terminal electron acceptor, resulting in the production of nitrogen-containing gases through the action of the NIR and nitric oxide reductase (NOR) enzymes (Braker and Conrad, 2011). The enzymatic synthesis of N2O by AOA has yet to be proven, and reported N2O emissions might be the result of spontaneous chemical events involving metabolic intermediates. There has yet to be discovered a NOR enzyme in any completely sequenced AOA, and NH2OH is not presently considered an intermediary in their AO route. While the processes that lead to the production of N2O appear to differ between AOA and AOB. Investigations using 18O labeling showed that NO2 has the same ratio of oxygen sources (one atom from H2O and one from O2) (Santoro et al., 2011). Furthermore, the site preference for 15N2O produced by AOA enrichments and pure cultures is similar with AOB culture values seen under AO conditions (L€oscher et al., 2012). This indicates that N2O in AOA comes from an intermediate in the AO process rather than AOB-like nitrifier denitrification (at least under the conditions studied) (L€ oscher et al., 2012). If HNO is an intermediary in the AOA AO route, it might be the source of archaeal N2O production (M.G. Klotz, personal communication). HNO molecules dimerize in water to create hyponitrous acid (H2N2O2), which is then dehydrated to N2O—a mechanism that was first suggested for denitrifying bacteria.

5.2.2 Yield While the quantity of N2O released by an enrichment culture containing “Ca. Nitrosoarchaeum koreensis” is much less than that emitted by Nitrosomonas europaea, N. maritimus SCM1 has been shown to produce up to five times more N2O than Nitrosococcus oceani and Nitrosomonas marina (L€oscher et al., 2012). An overestimation of rates due to the use of high-density cultures that may not be typical of the environment might explain this seemingly contradictory outcome ( Jung et al., 2011). Other possibilities include differences in the (relative) reactions of soil and marine ammonia oxidizers to low O2 conditions, or the presence of N2O-removing microorganisms in the enrichment culture ( Jung et al., 2011), which might lead to an underestimate of the rates. N2O emission inversely correlates with O2 concentration in N. maritimus SCM1 cultures, which supports the hypothesis that the gross quantity of N2O produced in studied marine oxygen minimum zones (OMZ) originates from group I.1a archaea (L€ oscher et al., 2012).

5.3 Niche differentiation

5.3 Niche differentiation The widespread presence of AEA/AOA in many natural and man-made systems raises issues about the nature of their cellular and biochemical adaptations, as well as their relative importance in nitrification. This section compares and contrasts the most essential variables that determine the ecology of AOA and AOB. While this chapter provides instances of niche differentiation in distinct AOA populations, it does not purport to be a complete list of habitats where amoA-like gene sequences have been discovered, and it stresses the significance of activity-based analysis in environmental research.

5.3.1 Oligotrophy The substrate concentration is one of the most important factors influencing the relative distribution of AOA vs AOB. In contrast to numerous other AOB studied, N. maritimus SCM1 has been
shown to have extremely high affinities Km  133 nM for joint NH+4 and NH3 and low (10 nM) thresholds for its substrate. Most notably, these figures match the in situ kinetics of oligotrophic open seas, which have been proven to contain significant quantities of AEA/AOA (Martens-Habbena et al., 2009). If the majority of marine N. maritimus-related populations are AOA, they will be accountable for the majority of AO seen in these environments, necessitating a reevaluation of present biogeochemical models (Martens-Habbena et al., 2009). Ca. Nitrososphaera gargensis was the first to show a predilection for low NH3 levels, while Ca. Nitrosotalea devanaterra was the second to do so lately (Lehtovirta-Morley et al., 2011). Furthermore, rising NH3 concentrations were found to lengthen the lag phases of three sedimentary/freshwater AOA enrichments. The soil AOA N. viennensis EN76 and “Ca. Nitrosoarchaeum koreensis,” on the other hand, were found to be adapted to much greater substrate concentrations (Tourna et al., 2011). While the inhibitory NH+4 doses of these AOA are equivalent to the most oligotrophic AOB reported (21.4 mM for N. oligotropha JL21), they are still low when compared to AOB with the highest NH+4 tolerance (50–1000 mM) (Koops et al., 2003). Several environmental studies have also revealed that AEA/AOA prefers low NH+4 concentrations, particularly in soils (Pratscher et al., 2011). This might be explained by greater affinities or densities of archaeal AMO or NH3/NH+4 transporters relative to AOB (134). However, the actual location of AMO’s catalytic site (i.e., whether it faces the periplasm or the cytoplasm), whether the stated transporters are employed for NH3 accumulation, and if archaeal AO is transporter dependent are yet unknown. Group I.1a archaeal NH3 transporters and permeases readings are among the most frequently identified genes and transcripts in marine samples (Hollibaugh et al., 2011).

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5.3.2 pH The finding that AOA prefers low NH3 concentrations is consistent with previous research that found large numbers and activity of AEA/AOA in acidic soils (Gubry-Rangin et al., 2010). The pH value was the most important element regulating AEA community structure (42%) among seven physicochemical characteristics tested (pH; carbon, nitrogen, and organic matter content; C:N ratio; soil moisture; and vegetation). Low pH reduces NH3 availability while increasing the toxicity of NO and N2O, as well as gaseous nitrogen dioxide (NO2). While certain AOB populations have evolved to cope with low pH, cultivated specimens exhibit little or no activity at pH 6.5. Ureases are enzymes that catalyze the conversion of urea to CO2 and NH3 in many, but not all, AOB and AEA/AOA species. The reaction products can subsequently be utilized as carbon and energy sources, as well as perhaps regulating the pH in the cell’s immediate proximity (Pommerening-R€oser and Koops, 2005). AEA/AOA are more transcriptionally active in acidic soils than AOB, and both microbial groups have phylotypes that are suited to low pH (Pester et al., 2012). Several lineages within group I.1b and the I.1a-associated group clearly exhibited adaptation to certain pH regimens in a study targeting amoA-like gene diversity in a wide range of globally distributed soils, and these results were consistent at the global, regional, and local level of sampling sites. The two sublineages of group I.1a-associated amoA-like sequences were shown to have a high association with acidic soils (pH 5) (Pommerening-R€oser and Koops, 2005). These findings support the fact that when urea was supplied as a substrate, AOA closely related to Nitrosotalea and Nitrososphaera thrived in some acidic soils (pH 3.75 and 5.4) when AOB could not be identified (Lu et al., 2012). The number and diversity of archaeal amoA-like genes are directly proportional to soil pH and only one obligately acidophilic AOA has been identified thus far (Lehtovirta-Morley et al., 2011). The widespread presence of the Nitrososphaera cluster and other amoA lineages in acidic soils, on the other hand, implies that additional low-pH-adapted AOA exist (Pester et al., 2012).

5.4 Conclusion Ammonia-oxidizing bacteria (AOB) have been playing a vital role in the ammonia oxidation. These ammonia-oxidizing bacteria convert ammonia to nitrite and then nitrite to nitrate in the nitrogen cycle. These nitrate and ammonia have also role in the pH and fertility of soil. The compensation of AOA and AOB system has also been examined in these studies. Rice farming increased AOA abundance but not AOB abundance, as measured by the number of copies of the amoA gene. In short, ammonia-oxidizing bacteria have a great role in oxidation of ammonia and convert it into by-products. These are part of nitrogen cycle and make a favorable environment for survival.

References

References Arth, I., Frenzel, P., Conrad, R., 1998. Denitrification coupled to nitrification in the rhizosphere of rice. Soil Biol. Biochem. 30, 509–515. Blainey, P.C., Mosier, A.C., Potanina, A., Francis, C.A., Quake, S.R., 2011. Genome of a low-salinity ammonia-oxidizing archaeon determined by single-cell and metagenomic analysis. PLoS One 6, e16626. Bollmann, A., Bullerjahn, G.S., Mckay, R.M., 2014. Abundance and diversity of ammoniaoxidizing archaea and bacteria in sediments of trophic end members of the Laurentian Great Lakes, Erie and Superior. PLoS One 9, e97068. Braker, G., Conrad, R., 2011. Diversity, structure, and size of N2O-producing microbial communities in soils—what matters for their functioning? Adv. Appl. Microbiol. 75, 33–70. Chen, X.P., Zhu, Y.G., Xia, Y., Shen, J.P., He, J.Z., 2008. Ammonia oxidizing archaea: important players in paddy rhizosphere soil. Environ. Microbiol. 10, 1978–1987. Chen, X., Zhang, L.M., Shen, J.P., Xu, Z.H., He, J.Z., 2010. Soil type determines the abundance and community structure of ammonia-oxidizing bacteria and archaea in flooded paddy soils. J. Soils Sediments 10, 1510–1516. Chen, X., Zhang, L.M., Shen, J.P., Wei, W.X., He, J.Z., 2011. Abundance and community structure of ammonia oxidizing archaea and bacteria in an acid paddy soil. Biol. Fertil. Soils 47, 323–331. Dang, H.Y., Zhang, X.X., Sun, J., Li, T.G., Zhang, Z.N., Yang, G.P., 2008. Diversity and spatial distribution of sediment ammonia-oxidizing crenarchaeota in response to estuarine and environmental gradients in the Changjiang Estuary and East China Sea. Microbiology 154, 2084–2095. Erguder, T.H., Boon, N., Wittebolle, L., Marzorati, M., Verstraete, W., 2009. Environmental factors shaping the ecological niches of ammonia-oxidizing archaea. FEMS Microbiol. Rev. 33, 855–869. Gubry-Rangin, C., Nicol, G.W., Prosser, J.I., 2010. Archaea rather than bacteria control nitrification in two agricultural acidic soils. FEMS Microbiol. Ecol. 74, 566–574. Hallam, S.J., Konstantinidis, K.T., Putnam, N., Schleper, C., Watanabe, Y.-I., Sugahara, J., Preston, C., De la Torre, J., Richardson, P.M., Delong, E.F., 2006. Genomic analysis of the uncultivated marine crenarchaeote Cenarchaeum symbiosum. Proc. Natl. Acad. Sci. 103, 18296–18301. He, J.Z., Shen, J.P., Zhang, L.M., Zhu, Y.G., Zheng, Y.M., Xu, M.G., Di, H.J., 2007. Quantitative analyses of the abundance and composition of ammonia-oxidizing bacteria and ammonia-oxidizing archaea of a Chinese upland red soil under long- term fertilization practices. Environ. Microbiol. 9, 2364–2374. Hollibaugh, J.T., Gifford, S., Sharma, S., Bano, N., Moran, M.A., 2011. Metatranscriptomic analysis of ammonia-oxidizing organisms in an estuarine bacterioplankton assemblage. ISME J. 5, 866–878. Jia, Z.J., Conrad, R., 2009. Bacteria rather than archaea dominate microbial ammonia oxidation in an agricultural soil. Environ. Microbiol. 11, 1658–1671. Jung, M.-Y., Park, S.-J., Min, D., Kim, J.-S., Rijpstra, W.I.C., Sinninghe Damste, J.S., Kim, G.-J., Madsen, E.L., Rhee, S.-K., 2011. Enrichment and characterization of an autotrophic ammonia-oxidizing archaeon of mesophilic crenarchaeal group I. 1a from an agricultural soil. Appl. Environ. Microbiol. 77, 8635–8647.

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K€onneke, M., Bernhard, A.E., Jose, R., Walker, C.B., Waterbury, J.B., Stahl, D.A., 2005. Isolation of an autotrophic ammonia-oxidizing marine archaeon. Nature 437, 543–546. Koops, H., Purkhold, U., Pommerening-R€ oser, A., Timmermann, G., Wagner, M., 2003. The Prokaryotes: An Evolving Electronic Resource for the Microbiological Community. Springer-Verlag, New York. Kowalchuk, G.A., Stephen, J.R., 2001. Ammonia-oxidizing bacteria: a model for molecular microbial ecology. Annu. Rev. Microbiol. 55, 485–529. Lehtovirta-Morley, L.E., Stoecker, K., Vilcinskas, A., Prosser, J.I., Nicol, G.W., 2011. Cultivation of an obligate acidophilic ammonia oxidizer from an nitrifying acid soil. Proc. Natl. Acad. Sci. U. S. A. 108, 15892–15897. Leininger, S., Urich, T., Schloter, M., Schwark, L., Qi, J., Nicol, G.W., Prosser, J.I., Schuster, S.C., Schleper, C., 2006. Archaea predominate among ammonia- oxidizing prokaryotes in soils. Nature 442, 806–809. L€ oscher, C.R., Kock, A., K€onneke, M., Laroche, J., Bange, H.W., Schmitz, R.A., 2012. Production of oceanic nitrous oxide by ammonia-oxidizing archaea. Biogeosciences 9, 2419–2429. Lu, L., Han, W., Zhang, J., Wu, Y., Wang, B., Lin, X., Zhu, J., Cai, Z., Jia, Z., 2012. Nitrification of archaeal ammonia oxidizers in acid soils is supported by hydrolysis of urea. ISME J. 6, 1978–1984. Martens-Habbena, W., Berube, P.M., Urakawa, H., Jose, R., Stahl, D.A., 2009. Ammonia oxidation kinetics determine niche separation of nitrifying Archaea and Bacteria. Nature 461, 976–979. Mosier, A.C., Allen, E.E., Kim, M., Ferriera, S., Francis, C.A., 2012. Genome sequence of “Candidatus Nitrosopumilus salaria” BD31, an ammonia-oxidizing archaeon from the San Francisco Bay estuary. Am. Soc. Microbiol. Nicol, G.W., Schleper, C., 2006. Ammonia-oxidising Crenarchaeota: important player sin the nitrogencycle. Trends Microbiol. 14, 207–212. Offre, P., Prosser, J.I., Nicol, G.W., 2009. Growth of ammonia oxidizing archaea in soil microcosms is inhibited by acetylene. FEMS Microbiol. Ecol. 70, 99–108. Park, B.-J., Park, S.-J., Yoon, D.-N., Schouten, S., Sinninghe Damste, J.S., RHEE, S.-K., 2010. Cultivation of autotrophic ammonia-oxidizing archaea from marine sediments in coculture with sulfur-oxidizing bacteria. Appl. Environ. Microbiol. 76, 7575–7587. Pester, M., Schleper, C., Wagner, M., 2011. The Thaumarchaeota: an emerging view of their phylogeny and eco physiology. Curr. Opin. Microbiol. 14, 300–306. Pester, M., Rattei, T., Flechl, S., Gr€ongr€ oft, A., Richter, A., Overmann, J., Reinhold-Hurek, B., Loy, A., Wagner, M., 2012. amoA-based consensus phylogeny of ammonia-oxidizing archaea and deep sequencing of amoA genes from soils of four different geographic regions. Environ. Microbiol. 14, 525–539. Pommerening-R€oser, A., Koops, H.-P., 2005. Environmental pH as an important factor for the distribution of urease positive ammonia-oxidizing bacteria. Microbiol. Res. 160, 27–35. Pratscher, J., Dumont, M.G., Conrad, R., 2011. Ammonia oxidation coupled to CO2 fixation by archaea and bacteria in an agricultural soil. Proc. Natl. Acad. Sci. 108, 4170–4175. Prosser, J.I., 1989. Autotrophicnitri-fication in bacteria. Adv. Microb. Physiol. 30, 125–181. Prosser, J.I., Nicol, G.W., 2008. Relative contributions of archaea and bacteriatoaerobic ammonia oxidation in the environment. Environ. Microbiol. 10, 2931–2941. Ravishankara, A., Daniel, J.S., Portmann, R.W., 2009. Nitrous oxide (N2O): the dominant ozone-depleting substance emitted in the 21st century. Science 326, 123–125.

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Santoro, A.E., Buchwald, C., Mcilvin, M.R., Casciotti, K.L., 2011. Isotopic signature of N2O produced by marine ammonia-oxidizing archaea. Science 333, 1282–1285. Shen, J.P., Zhang, L.M., Zhu, Y.G., Zhang, J.B., He, J.Z., 2008. Abundance and composition of ammonia-oxidizing bacteria and ammonia-oxidizing archaea communities of an alkaline sandyloam. Environ. Microbiol. 10, 1601–1611. Tourna, M., Freitag, T.E., Nicol, G.W., Prosser, J.I., 2008. Growth, activity and temperature responses of ammonia-oxidizing archaea and bacteria in soil microcosms. Environ. Microbiol. 10, 1357–1364. Tourna, M., Stieglmeier, M., Spang, A., K€onneke, M., Schintlmeister, A., Urich, T., Engel, M., Schloter, M., Wagner, M., Richter, A., 2011. Nitrososphaera viennensis, an ammonia oxidizing archaeon from soil. Proc. Natl. Acad. Sci. 108, 8420–8425. Treusch, A.H., Leininger, S., Kletzin, A., Schuster, S.C., Klenk, H.P., Schleper, C., 2005. Novel genes for nitrite reductase and Amo-related proteins indicate a role of uncultivated mesophilic crenarchaeota in nitrogen cycling. Environ. Microbiol. 7 (12), 1985–1995. Venter, J.C., Remington, K., Heidelberg, J.F., Halpern, A.L., Rusch, D., Eisen, J.A., Wu, D., Paulsen, I., Nelson, K.E., Nelson, W., Fouts, D.E., Levy, S., Knap, A.H., Lomas, M.W., Nealson, K., White, O., Peterson, J., Hoffman, J., Parsons, R., Baden-Tillson, H., Pfannkoch, C., Rogers, Y.-H., Smith, H.O., 2004. Environmental genome shotgun sequencing of the Sargasso Sea. Science 304, 66–74. Walker, C.B., de la Torre, J.R., Klotz, M.G., Urakawa, H., Pinel, N., Arp, D.J., BrochierArmanet, C., Chain, P.S.G., Chan, P.P., Gollabgir, A., Hemp, J., H€ ugler, M., Karr, E. A., K€onneke, M., Shin, M., Lawton, T.J., Lowe, T., Martens-Habbena, W., SayavedraSoto, L.A., Lang, D., Sievert, S.M., Rosenzweig, A.C., Manning, G., Stahl, D.A., 2010. Nitrosopumilus maritimus genome reveals unique mechanisms for nitrification and autotrophy in globally distributed marine crenarchaea. Proc. Natl. Acad. Sci. U. S. A. 107, 8818–8823. Wang, Y.A., Ke, X.B., Wu, L.Q., Lu, Y.H., 2009. Community composition of ammoniaoxidizing bacteria and archaea in rice field soil as affected by nitrogen fertilization. Syst. Appl. Microbiol. 32, 27–36. Wu, Y.C., Lu, L., Wang, B.Z., Lin, X.G., Zhu, J.G., Cai, Z.C., Yan, X.Y., Jia, Z.J., 2011. Longterm field fertilization significantly alters community structure of ammonia oxidizing bacteria rather than archaea in a paddy soil. Soil Sci. Soc. Am. J. 75, 1431–1439. Wuchter, C., Abbas, B., Coolen, M.J., Herfort, L., van Bleijswijk, J., Timmers, P., Strous, M., Teira, E., Herndl, G.J., Middelburg, J.J., 2006. Archaeal nitrification in the ocean. Proc. Natl. Acad. Sci. 103, 12317–12322. Yao, H.Y., Gao, Y.M., Nicol, G.W., Campbell, C.D., Prosser, J.I., Zhang, L.M., Han, W.Y., Singh, B.K., 2011. Links between ammonia oxidizer community structure, abundance, and nitrification potential in acidic soils. Appl. Environ. Microbiol. 77, 4618–4625. Ying, J.Y., Zhang, L.M., He, J.Z., 2010. Putative ammonia- oxidizing bacteria and archaea in an acidic red soil with different land utilization patterns. Environ. Microbiol. Rep. 2, 304–312. Zhang, L.M., Offre, P.R., He, J.Z., Verhamme, D.T., Nicol, G.W., Prosser, J.I., 2010. Autotrophic ammonia oxidation by soil thaumarchaea. Proc. Natl. Acad. Sci. U. S. A. 107, 17240–17245. Zhang, L.M., Hu, H.W., Shen, J.P., He, J.Z., 2012. Ammonia oxidizing archaea have more important role than ammonia-oxidizing bacteria in ammonia oxidation of strongly acidic soils. ISME J. 6, 1032–1045.

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6

Ayesha Kanwala, Zeeshan Ahmad Bhuttab, Moazam Alic, Ambreen Ashard, and Muhammad Shoaibe Institute of Biochemistry, Biotechnology and Bioinformatics, The Islamia University of Bahawalpur, Bahawalpur, Pakistan bLaboratory of Biochemistry and Immunology, College of Veterinary Medicine, Chungbuk National University, Cheongju, Chungbuk, Republic of Korea c Department of Clinical Medicine and Surgery, University of Agriculture, Faisalabad, Pakistan d Department of Chemistry, University of Agriculture, Faisalabad, Pakistan eInstitute of Microbiology, University of Agriculture, Faisalabad, Pakistan a

6.1 Introduction Advancements and modernization in the chemical industry have led to massive releases of environmental pollutants via sewage, especially those containing nitrogen. This is a major factor leading to eutrophication. Both conventional nitrogen detoxification methods–nitrification and denitrification–are now proved to be less functional, demanding some other methods for nitrogen fixation in wastewater. If we consider the microbial conversion of nitrogen, it comprises ammonia and nitrogen oxidation, as mentioned in the equations (NH3-N ! NO2  -N) and (NO2 -N ! NO3 -N), respectively. Unfortunately, rate limitation is a factor that discourages the nitrification process, and leading researchers worldwide to adopt ammonia oxidation as the best biological method for nitrogen removal from wastewater. The oxidation of ammonia via microbes to nitrite in the biogeochemical nitrogen process or cycle is one of the most critical and initial steps. As a result of the nitrification process, the loss of nitrogen leads to environmental destruction in the form of nitrate, a major subject for leaching, and the production of a major greenhouse gas such as nitrous oxide. The direct production of nitrous oxide during ammonia oxidation is further aided by the denitrification process. Yearly 120 Tg of ammoniabased fertilizers are used worldwide but the anthorpogenic ammonia has extraordinary effects on the geochemical cycle. Microbes dealing with ammonia oxidizing

Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00010-3 Copyright # 2022 Elsevier Inc. All rights reserved.

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are the core of the nitrogen fate in the ecosystem, and the three diversified microbial groups are as follows: • • •

Ammonia-oxidizing bacteria (AOB) Ammonia-oxidizing archaea (AOA) Comammox bacteria

Another microbial group that anaerobically oxidized the ammonia is the anammox bacteria, but this group has a separate phylogenetic and biochemical makeup from aerobes and is taken separately in this regard (ammonia-oxidizing process). Schloesing and Muntz (1877) were pioneers in suggesting that the oxidation of ammonia is a biological process instead of a chemical process. Their study stated that the soil-based nitrification process was halted by applying chloroform. The very first cultivates of AOB were provided by Frankland and Frankland in 1890, followed by the isolated Nitrosomonas europaea by Sergei Winogradsky that is still a model AOB in today’s research (Frankland and Frankland, 1890; Winogradsky, 1890). β-proteobacteria (Nitrosomonas and Nitrosospira) and γ-proteobacteria (Nitrosococcus) are two such phylogenic groups confined to the canonical AOB.

6.2 Ammonia-oxidizing bacteria 6.2.1 Ecology The diversified presence of microbes responsible for ammonia oxidation involves the soil; freshwater sources such as lakes, ponds, rivers, marine water beds, and wastewater treatment houses; and even the skin of human beings (Koskinen et al., 2017). The ubiquitous nature of AOBs varies according to the environment, which makes them a central key in regulating the ecosystem as different AOBs respond differently to ecological changes and ultimately varying levels of greenhouse gas emissions. According to Hink et al. (2017), AOBs generate more nitrous oxide gas than AOAs. Similarly, the response of ammonia oxidizers is different to commercial level inhibitors. For example, in allylthiourea used in agriculture management, AOBs show a more sensitive response in contrast to AOA (Shen et al., 2013). Drought is another major factor that has a serious impact on crops and vegetation every year. AOBs are more resistant to drought conditions whereas AOAs have a lesser ability; this makes AOBs more robust to ecological changes (Placella and Firestone, 2013). If we look at the population statistics of theses microbes, then AOAs outstrip AOBs in magnitude in aquatic and soil environments. An estimate stated that the AOA population in oceans is 1  1028 cells, which makes them one of the most abundant living entities on Earth. They comprise up to 40% of the total prokaryote population in marine waters and 1%–5% in the terrestrial environment (Leininger et al., 2006). On the other hand, AOBs dominate AOAs in wastewater treatment plants and intermittently in soils fertilized with a nitrogen base (Bates et al., 2011). Limited data are available regarding Comammox Nitrospira; however, their abundance in the

6.2 Ammonia-oxidizing bacteria

ecosystem is comparatively higher than other ammonia oxidizers in some wastewater treatment plants and soil ecosystems (Pjevac et al., 2017). The nucleotide study of ammonia oxidizers, especially 16S rRNA and the AmoA gene that encodes for the ammonia monooxygenase enzyme, helps scientists gather information regarding their ecology. Ammonia monooxygenase is a fundamental enzyme mutually shared by all three categories of ammonia microbial oxidizers. In a generalized form, the soil ecosystem and engineered ecosystems–wastewater treatment plants are favorites for β-proteobacteria, whereas marine water beds are a hot spot for γ-proteobacteria. Nitrosococcus TAO100, a recently evolved strain, is an exception (Hayatsu et al., 2017). Similarly, γ-proteobacteria is also detected in acidic environments (Fumasoli et al., 2017). An experimental study revealed the cultivation of the TAO100 strain upon acidic soil; however, the genomic and amplicon sequencing data stated the rare availability of γ-proteobacteria in soil. Members of Nitrosospira have the affinity to dominate the ammonia oxidizer communities present in the soil. However, N. europaea originated from a soil culture; similarly, the members of Nitrosomonas have a lesser population in contrast to Nitrosospira. The major clusters of AOAs presented by researchers (Pester et al., 2012) are as follows: (a) (b) (c) (d) (e)

Nitrososphaera Nitrosocosmicus Nitrosocaldus Nitrosotalea Nitrosopumilus

The soil with a pH ranging between the neutral zones subjugated by the Nitrososphaera AOA along with Nitrosoarchaeum and Nitrosocosmicus (Gubry-Rangin et al., 2011); the soil having an acidic pH contains Nitrososphaera and Nitrosotalea whereas the hot springs are a favorite place to collect the Nitrosocaldus and Nitrososphaera devanaterra. Nitrosopumilus and Nitrosopelagicus AOA inhabit marine water habitats while the freshwater residents are Nitrosoarchaeum and Nitrosotenuis (Santoro et al., 2015). It is frequently observed that AOAs thrive in acidic and high temperature surroundings, but both these conditions are inaccessible to AOB. However, the expectation is possible anywhere, Nitrosotalea devanaterra, one of the first obligate acidophilic AOB cultivated from agriculture soil enriched with acidic pH and Nitrosocaldus yellowstoneii (thermophilic in nature) cultivated in a hot spring, provide explanations of nitrification in such environments where AOB growth was not previously possible (Lehtovirta-Morley et al., 2011).

6.2.2 Environmental regulators of ammonia oxidation Several broad-spectrum studies have been executed on the ecological factors that regulate ammonia oxidizer niches and their role in the nitrogen flux to the environment. The two central drivers of ammonia oxidizer diversification and profusion are the environmental pH level and the ammonia concentration. Some other factors also

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involved in AOB dissemination into the ecosystem include the response to organic components, metal exposure, light source, salinity, and temperature (Merbt et al., 2012). Several examples illustrate this behavior, such as higher temperatures boost AOA activity in terrestrial and aquatic ecosystems. Correspondingly, the optimum temperature range of AOAs is relatively higher than AOBs (Zeng et al., 2014; Tourna et al., 2011). Urea favors the growth of few but not all AOBs and AOAs, and cyanate presence aides in AOA N. gargensis growth (Palatinszky et al., 2015). N. europaea, an AOB exhibited and mixotrophic growth on fructose and pyruvate cultures (Hommes et al., 2003), catalase negative AOA growth was aided by pyruvate an organic acid by relieving the oxidation stress (Kim et al., 2016). However, the environmental relevance is not widely explored. The concentration of ammonia in the soil is the key selector that determines the presence of ammonia oxidizer microbes (Bates et al., 2011). It has been frequently reported that soil enriched with higher ammonia concentrations is a favorite place for AOBs in contrast to AOAs, which tend to accumulate in the soil with lower ammonia concentrations. Marine beds have relatively lower ammonia levels up to  10 nM and are a favorite place for AOA growth rather than AOB (Martens-Habbena et al., 2009). In general, the concept has been established that the ecosystems with lower ammonia levels favor AOA growth, except for recent studies in which Nitrosocosmicus-AOA was cultivated in higher ammonia concentrations (Sauder et al., 2017). Comammox Nitrosospira studies showed that AOAs tend to grow in lower ammonia ecosystems, although more findings are needed for the diversification and physiological activities of Comammox (Kits et al., 2017). The bioavailability of ammonia and the pH level of the environment are interlinked as two different forms of ammonia are present in the environment: the NH3 unprotonated type and the NH4 + protonated one having pH-equilibrium (pKa ¼ 9.27). The acidic conditions hinder the availability of NH3. The Suzuki and his team stated that N. europaea and AOB have the capacity to use only NH3 rather than NH4 + substituting the only reason why AOB are impotent to grow in acidic ecosystems. The study done by Suzuki et al. (1974) is still work as a base as it has never been repeated with other ammonia oxidizer strains, although the capacity to utilize the ammonia in lower pH not eradicate the survival and working of AOB; because the AOB has the capacity to grow under acidic environments when urea was given as substrate in biofilms. N. devanaterra overturned the concept of ammonia oxidizers working under lower pH when it was discovered under acidophilic conditions (Lehtovirta-Morley et al., 2011). Initially isolated from soil having acidic conditions, it grows autotrophically in the lab at a 4–5.5 pH range, with NH4Cl as its only energy regulator. Nitrosotalea is the only microbe accomplishing the nitrification process in acidic conditions, but the fact is that it is acidophilic in nature. Besides this, acid-resistant γ- and β-proteobacteria cultivation has also been recorded (Hayatsu et al., 2017). Along with AOB tolerating acidic environments, the genus Nitrososphaera is also frequent in acidic soils worldwide and also grows efficiently in acidic microcosms (Wang et al., 2014). It is quite surprising that little knowledge is available

6.2 Ammonia-oxidizing bacteria

for the nitrification process under basic or alkaline conditions. Only specified ammonia oxidizer lineages do nitrification in basic conditions, but no obligate alkalinophilic AOBs have been discovered yet.

6.2.3 Strategic functional, anatomical, and biological differentiations among ammonia oxidizers The configuration and functionality of ammonia oxidizers is mainly determined by environmental factors, but the pathways involved in this underpinning are not clear. Some strategic functional, anatomical, and biochemical modifications of both AOB and AOA determine their presence, environmental adoptability, and nitrogen turnover rate. Ammonia oxidation kinetics and an affinity for the substrate of ammonia oxidizers explains why a typical form of microbes such as AOB, AOA, and Comammox Nitrospira dominates in distinct environments. Studies have confirmed the elevated half saturation constants (Km) and subordinate substrate affinity for AOB in contrast to AOA and Comammox Nitrospira. The range of Km NH3 of ammonia oxidizers is 6 and 11 μM for AOB; 3.6 and 4.4 μM for AOA; and 49 nM for Comammox Nitrospira inopinata (Kits et al., 2017). There is an inverse relationship between the affinity and substrate concentration, as indicated by molecular ecology surveys where AOA abundance was more in marine and lesser nitrogen content soil while AOBs were frequently located in wastewater plants with relatively higher ammonium levels. The Comammox are oligotrophic. Having a complete understanding of the optimal pathway kinetic theory, it has been suggested that the Comammox yield is higher with a declined growth ratio rather than partial ammonia oxidation, which makes Comammox a viable environment even in lower ammonia concentrations (Costa et al., 2006). It is fascinating that both AOA and Comammox roughly lie in the same half-saturation constant range. This property makes it unclear how the half saturation constant range affects their capability to diversify in the environment. The ammonia oxidizers of AOB, AOA, and Comammox are autotrophic but surprisingly each of them has a different inorganic carbon integration pathway. Studies have found that the atmospheric carbon dioxide fixation by AOB via the Calvin cycle while Nitrospira does it via the reductive tricarboxylic acid cycle and archaea (AOA) fix the HCO3 by the hydroxypropionate-hydroxybutyrate cycle (Berg et al., 2007). The energy requisite of these cycles varies accordingly and the most energyefficient cycle is the archaea hydroxypropionate-hydroxybutyrate cycle for inorganic carbon integration aerobically (K€ onneke et al., 2014). Additionally, carbon dioxide and HCO3 are pH-dependent, and the concentration of HCO3 declines as the pH decreases. It seems like a contradictory presence of acidophilic AOA, meaning that AOAs have a significant affinity toward HCO3 for growing at the lower pH levels. Major study gaps are present to relate the environmental and physiological concerns to the above-mentioned differences. Structural differences are also present, such as all AOBs have widespread intracytoplasmic entities while AOAs and

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Comammox lack such entities and have a plasma membrane; some additional membrane-bound structures were recorded in Nitrosoarchaeum koreensis MY1 and Nitrososphaera viennensis (Stieglmeier et al., 2014). The membrane aggregate is crucial for ammonia oxidation as the enzyme ammonia monooxygenase that triggers ammonia oxidation is membrane-bound. It has been extensively accepted that the widespread membranes in the AOAB are especially for the maximum accommodation of the monooxygenase enzyme. Therefore, it has been postulated that a singular AOB cell might have more ammonia monooxygenase content in contrast to the AOA while Comammox lacks the final confirmation. The Vmax value for ammonia oxidation for AOA and Comammox lies roughly in line, 14.8 μmol N mg protein1 h1, but it is much greater in the AOB N. europaea (Kits et al., 2017). Similarly a 10-fold higher rate per cell is also recorded in AOB Nitrosospira multiformis paralleled to the AOA N. viennensis (Kozlowski et al., 2016). Iron-based respiratory chains are present in AOB while AOA contains the copper-based respiratory chains due to a lack of cytochrome c proteins. Here, the multicopper oxidases and tiny blue-copper enclosing proteins work as substitutes for mislaid cytochromes (Walker et al., 2010).

6.3 Anaerobic ammonium oxidation bacteria In 1977, chemist Engelbert Broda anticipated the existence of formerly unidentified lithotrophic microbes that oxidize ammonia to nitrogen NH4 + to N2 along with NO2  or NO3  as electron acceptors (Broda, 1977). Furthermore, the presence of anammox to nitrogen gas by using NOx  was also concluded from the stoichiometric nitrogen loss from anoxic fjords and basins (Richards, 1965). After a gap of almost three decades, a denitrifying bioreactor confirmed the presence of the anammox process in the Netherlands (Mulder et al., 1995). The bacteria that catalase the anammox process oxidizes the ammonia to nitrogen gas along with nitrous oxide NO2  as an electron receiver. Initial stoichiometric studies by Strous described the anammox process (Strous et al., 1998). This was recently reviewed via a highly enhanced planktonic anammox bacteria cell culture as given in Eq. (6.1) (Lotti et al., 2014): 1NH4 + + 1:146NO2  + 0:071HCO3  + 0:057H+ ! 0:986N2 + 0:161NO3  + 2:002H2 O+ 0:071CH1:74 O0:31 N0:20

(6.1)

The three enzymatic reactions that occur in the anammox process are (Kartal et al., 2013): (a) Reduction of NO2  to nitric oxide (NO). (b) Biosynthesis of hydrazine from NO and NH4 + . (c) Dehydration of hydrazine to N2 gas. Anammoxosomes are the cellular compartments that are membrane-bound where hydrazine dehydration and biosynthesis occur (De Almeida et al., 2015). Anammoxosome comprises some rare lipids, ladderane lipids that have linear

6.3 Anaerobic ammonium oxidation bacteria

concatenated cyclobutane moieties. The affiliation of anammox bacteria to the phylum Planctomycetes was proposed by Strous (Strous et al., 1998) along with five tentative candidate genera based on 16S and 23S rRNA sequencing connections: Candidatus Kuenenia, Candidatus Brocadia, Candidatus Anammoxoglobus, Candidatus Jettenia, and Candidatus Scalindua ( Jetten et al., 2015). Phylogenetic diversified bacterial species are present in anammox that are frequently symbolized by ecological 16S rRNA genetic arrangements: Anammoxoglobus/Jettenia cluster I and Brocadia clusters I, II, III, and IV. The microbiological as well as geochemical outlooks of the anammox bacteria are of great importance as they reflect their environmentalism and normal functioning, as both outlooks have a significant impact on geographical diversification and the geochemical worth of anammox in the ecosystem. In this regard, scientific authors recapitulate the anammox environmentalism and physiology by considering their ecology, geographical diversification, geochemical worth, and environmental regulators besides physiological chattel and conceivable niche disparity in anoxic environments. Recently, genetic studies of anammox bacteria have provided a tremendous opportunity to further explore the physiological chattel of anammox bacteria.

6.3.1 Ecology 6.3.1.1 Geographical distribution The activities and populations of anammox bacteria have been found in various anoxic ecosystems, including marine, brackish, freshwater, and terrestrial environments. In addition, anammox bacterial populations were detected even in harsh environments, including hydrothermal vents, hypersaline basins, sea ice, permafrost soil, and oil-contaminated fields. This indicates that anammox bacteria are ubiquitously distributed in anoxic ecosystems. The geographic distributions of anammox bacteria indicate that they likely have genus-specific and species-specific habitats. For example, anammox bacteria affiliated with the genus Ca. Scalindua were exclusively found in marine environments (Schmid et al., 2007). In addition, the abundance of Ca. Scalindua increased with the increase in salinity of estuarine areas (Dale et al., 2009). The exclusive distribution of Ca. Scalindua in marine environments was further ascertained in the recent work by Sonthiphand et al. (2014), who investigated the relationships between natural habitats and the geographic distribution of anammox bacterial 16S rRNA gene sequences, including next-generation sequencing data. On the other hand, the habitat range is different even within the genus Ca. Scalindua. Phylogenetic analysis revealed that the 16S rRNA gene sequences were clustered together into regionspecific phylogenetic clades: Peru clusters I, II, and III, Namibia, Arabian Sea, and Black Sea clusters. These region-specific distributions suggest the presence of different niches within the genus Ca. Scalindua. In addition to the geographic distribution of Ca. Scalindua, the population shift of anammox bacteria during long-term cultivation also suggests that the bacteria have genus-specific niches, such

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as the population shift from (i) Ca. Brocadia to Ca. Kuenenia, (ii) Ca. Kuenenia to Ca. Brocadia, and (iii) Ca. Brocadia to Ca. Anammoxoglobus.

6.3.1.2 Geochemical importance and important environmental constituents Large amounts of nitrogen are transferred to the atmosphere by anammox bacteria. Geochemical importance determination is done with 15N isotope pairing method and nature sample can be whole and slurry core is incubated under the two conditions such as 15NH4 + and 14NOx , 14NH4 + and 15NOx , and 29N2 gas production is determined by gas chromatography isotope ratio mass spectrometry. The Namibian oxygen minimum zone (OMZ) technique is used by scientists to calculate the total nitrogen loss; almost 79% of nitrogen is determined under marine sediments. In comparison to sea water and brackish regions, the total nitrogen loss measurement by the anammox method was greater than 35% (regionally almost 94%) in the freshwater ponds of riparian regions (Zhu et al., 2013). It was also greater than 13%, 41%, and 80% in the water columns of ponds (Schubert et al., 2006); paddy fields (Nie et al., 2015); and ground water containing ammonium, respectively (Robertson et al., 2012). This anammox activity greatly relies on regional environmental conditions. For example, its activities in the OMZ vary from the water depth in the eastern South Pacific by ranging from greater than 0.1 to 5.8 nmol-N2 L1 h1 (Dalsgaard et al., 2012). However, its activities are not significant in the Arabian OMZ as compared to the Namibian and Peruvian. Dissolved oxygen (DO) largely affects the environmental conditions (De Brabandere et al., 2014). Further, many other factors also affect the large growth of anammox bacteria or its activity (NOx  concentration) (Zhu et al., 2013), sediment activity (Lisa et al., 2014), salinity (Sonthiphand et al., 2014), manganese oxide (MnO) levels (Engstr€om et al., 2005), molar ratio of NH4 + to NOx  (Li et al., 2013), sulfide concentration ( Jensen et al., 2009), and pH (Li et al., 2013). There is also another possibility that these factors can affect those microorganisms that are in competition under the consumption of NH3 or NO2  such as aerobic ammonia and nitrite oxidizers. A study mentioned that there is a lot of competition among microbes such as anammox bacteria and AOA in sea water columns and anammox bacteria and nitrite oxidizers in sediments (Lipsewers et al., 2014). Denitrifiers are in competition with anammox bacteria for NOx , on the same point it also provides the NO2  to anammox procedure, firstly reduce the NO3  to NO2  and then convert it into NO2  that is used by the anammox procedure (Zhou et al., 2014).

6.3.2 Physiology of anammox bacteria The functional properties have been identified with the help of four anammox bacterial genera: Ca. Kuenenia, Ca. Scalindua, Ca. Brocadia, and Ca. Jettenia. These functional properties make it possible for researchers to isolate the anammox bacteria from the niche. Ca. Brocadia likes NH4 + and NO2  rich environments due to its

6.3 Anaerobic ammonium oxidation bacteria

low affinity toward NH4 + and NO2  . On the opposite side, the Ca. Kuenenia stuttgartiensis shows greater affinity toward NH4 + and NO2  while the production rate is very low compared to Ca. Brocadia sinica; these functional properties make clear that Ca. Kuenenia stuttgartiensis and Ca. Brocadia sinica show r and k tacticians, respectively. The suitable position of Ca. Brocadia sinica and Ca. Kuenenia stuttgartiensis is determined by the help of the Monod equation. Ca. Brocadia sinica cells show efficient growth when the concentration of NO2  is higher than 80 μM. In fact, the species members of Ca. Brocadia sinica are observed when they are treated at the higher level of NH4 + and NO2  in bioreactors (Oshiki et al., 2011). The geographic distribution of the genus Ca. Scalindua is determined by the salt concentration parameter (Sonthiphand et al., 2014). The latest physiological studies show that Ca. Scalindua sp. cells lose their anammox activity when the salt concentration is greater than 1.5% w/v, which reveals that they are halophilic microbes (Awata et al., 2013). This physiological characteristic is a deal with the regional distribution of Ca. Scalindua and this bacterial gene’s 16S rRNA abundance increases as the salinity increases. However, this Ca. Scalindua 16S rRNA gene sequence is observed in municipal sewage treatment plants; it gives information that some species of Ca. Scalindua can live in an environment with a shortage of salt concentration. Another bacterium named Ca. Jettenia is known as a freshwater genus as its 16S rRNA gene sequence is obtained from the freshwater environment. The latest functional study also shows that the anammox activity of Ca. Jettenia caeni (early makes as a planctomycete KSU-1) was restrained even at a low salt concentration, that is, a saline IC50 of 68 mM (Ali et al., 2015). On the opposite side, a 513 mM NaCl concentration ensures the anammox activity of Ca. Kuenenia stuttgartiensis; however, its regional distribution is confined to hydrothermal sites and sea sponges. These findings revealed that salt concentration and some other conditions also play important roles in identifying the domination of the genus Ca. Scalindua in the sea environment. Another functional characteristic is temperature, which determines the regional distribution of Ca. Scalindua and Ca. Brocadia. Ca. Scalindua is prosperous at lower temperature ranges compared to other anammox bacteria; however, it can flourish at low temperatures. The maximum activity of Ca. Scalindua was identified at 15°C, 12°C, and 12–18°C by using the sea residual samples from Skagerrak in the Baltic North Sea transition, and the in situ temperature is 6°C, shores of Greenland with greater than 1°C temperature and Arctic fjord with temperature range of 0.2°C to 2.1°C. Lately, sulfide has become known as a crucial component of determining the regional distribution of anammox bacteria in the sea environment ( Jin et al., 2013). Sulfide is a toxic compound for organisms, and there is no anammox activity where the sulfide concentration is higher than that of the micromolar ( Jensen et al., 2009). A repressive sulfide concentration gives confirmation that the enriched culture of the anammox bacteria is found within the range of 10–100 μM. Indeed, 4 μM is a lower concentration of sulfide that completely represses the anammox activity in anoxic water columns of the Arabian Sea. On the opposite side, sulfide-oxidizing

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bacteria associated with the genus Sulfurimonas use sulfide to reduce NO3  by acting as an electron donor. Furthermore, the sulfide-oxidizing bacteria associated with the genus Thioploca cause respiratory ammonification to reduce the NO3  to NH4 + through NO2  (i.e., nitrate reduced to ammonium dissimilatory). Sulfide-oxidizing bacteria increase the bionomic existence of anammox bacteria as they lower the poisonous sulfide while giving NO2  and NH4 +. In fact, sulfide-dependent denitrification triggers the anammox activity of estuary sediments (Lisa et al., 2014); sea anammox bacteria and sulfide-oxidizing bacteria coexistence shows in situ anoxic sea sediments and science laboratory-enriched cultures. Anammox bacteria also take part either in oxidation or reduction reactions rather than anammox activities. For example, the anammox bacteria can oxidize ferrous ion, propionate, formate, methylamine, acetate, and hydrogen for respiratory ammonification, although some specific actions such as NO3  reduction in general are lower compared to anammox activity (Oshiki et al., 2013). Furthermore to NO, anammox bacteria could reduce manganese oxides and ferric ion along with the formate, acetate, and propionate (Zhou et al., 2014). Especially acetate, the formate-oxidizing activity, and propionate activity of anammox bacteria were clearly different from the other bacterial strains. For example, the Ca. Brocadia fulgida cells had more metabolic potential to oxidize acetate. They fully compete with other bacteria that live in coexistence with anammox bacteria in bioreactors where acetate was applied as a supplement in addition to NH4 + , NO2  , and NO3  ( Jenni et al., 2014). Thus, the metabolic activity of the respiratory ammonification affects the genus-specific and species-specific classification of anammox bacteria. New research is increasing our knowledge about anammox bacterial functionality. Also, new genomic sequences have been identified from different species of Ca. Kuenenia and Ca. Scalindua. When the genomic sequence of these species are compared, this shows that 75% of the genes are commonly shared. The genes they share are involved in anammox processes such as hydrazine dehydrogenase (hdt) or hydroxylamine dehydrogenase (hao), acetyl CoA synthase (acsAB), nitrite or nitrate oxidoreductase (nxrAB), and hydrazine synthase (hzs). An orthologue sequence of nirS encoded cytochrome cd1-type NO generating nitrate reductase was identified in the Ca. Scalindua profunda and Ca. Kuenenia stuttgartiensis. On the other hand, nirK encoded copper containing NO forming nitrate reductase was observed in the Ca. Jettenia caeni genome. But the most interesting thing is that these nirS and nirK orthologues were absent in the Ca. Brocadia sinica and Ca. Brocadia fulgida genomes (Oshiki et al., 2015). This insufficiency demonstrates that multiple kinds of nitrite reductase are working within the anammox bacterial classes. Genes required for the TCA cycle are observed only as species-specific, consisting of citrate synthase and citrate lyase found in Ca. Scalindua brode and Ca. Scalindua profunda (Speth et al., 2015). Some interesting functionality information of anammox bacteria is inferred from the genome while the rest of the information is obtained from physiological and biochemical experiments, such as iron or manganese oxidation and reduction, hydrazine synthesis, NO2  reduced to NO, the existence of peptidoglycan, the use of anammox bacterial NirK in NO2  reduction, and the S-layer membrane.

6.4 Microbial interactions

This information indicates that the anammox bacterial genome sequence data are perfect to observe the physiological properties without culturing. This information increased the examination of anammox bacterial physiology and makes possible the investigation of biochemical and genetic procedures that are being undertaken in the niche of anammox bacteria.

6.4 Microbial interactions and their contribution to enhanced nitrogen removal There are a lot of interactions among the microbes and anammox bacteria, which are involved in the nitrogen removal system. Usually, in an oxygenated atmosphere, AOB gives nitrite for anammox bacteria, although NOB may fight for nitrite. It has been confirmed that there is a collaboration between the AOB and anammox bacteria. In anaerobic environments, nitrite is produced by the nitrate-reducing bacteria for the anammox bacteria in the presence of electron donor limited conditions, although denitrifying bacteria also fight for nitrite when enough electron donors are available. Anammox bacteria with sufficient ammonium are available by the application of DNRA. In addition, a few heterotrophs also fight with denitrifiers by digesting organic compounds as fast as the denitrifiers, which plays an important role in the support of anammox bacteria (Tao et al., 2019). AOB can survive in an anammox reactor, but it shows a very slow growth rate (Ye et al., 2018). For instance, the Nitrosomonas eutropha anaerobic activity is 50 times slower than that of the anaerobic ammonium oxidizer Ca. Brocadia anammoxidans while it is 200-fold slower than that of the aerobic activity of N. eutropha. Under some properties, Nitrospira does not fight with the anammox bacteria, although it helps anammox bacteria by giving nitrite over canonical ammonia oxidation (Ciesielski et al., 2018). To maintain the balance of the anammox activity, it is necessary to wash out NOB from the autotrophic nitrogen removal system. Divisions between the flocs and biofilms is a good way to clean the NOB from the system. Laureni et al. (2019) stated that floc removal is an efficient method to clean NOB from the system while the use of a biofilm system to control NOB in mainstream nitration and anammox (PN/A) applications is far better than the biofilm system. By keeping NOB and anammox bacteria separated in flocs and biofilm, the competition between them for NO2 is seen easily and their different actual growth rates, that is, mNOB < mAOB in flocs and let the selective NOB to remove on a larger scale of simulated SRTs (6.8–24.5d) (Laureni et al., 2019). In anammox systems, microorganism communities are divided between flocs and biofilm. Flocs have denitrification activity that protects them from anammox activity in biofilm, which produces a large number of organic loadings while a loss of ammonia may happen (Yang et al., 2019). This process is being conducted in the plug flow granular sludge-based scale vector (4 m3) at 19  1°C, and it was continuously supplied with the actual effluent of the A stage of a wastewater treatment plant (WWTP). Under these conditions, the anammox bacteria can grow

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in the mainstream and the newly produced granules remain safe in the system. On the other hand, the heterotrophic biomass grows in the flocs and is removed from the system effectively. Different microorganisms interact with other processes to increase the removal of nitrogen. Xie et al. (2018) used a unique technology integrating anammox denitrifying anaerobic methane oxidation (DAMO) reactions in a membrane biofilm reactor (MBfR); an effluent total nitrogen (TN) concentration below 3.0 mg N/L and a TN removal rate of 0.28 kg N/m3 d could be achieved. In this procedure, 30%–60% of the nitrate produced during anammox was converted back to nitrite by DAMO archaea, and then the nitrite wase washed out by the anammox and DAMO bacteria with contributions of >90% and 90

N.R.

15

74

>90

100

62

40

Thiobacillus denitrificans N.R.

15

800

16

Paraccocus sp.

80

4000

100

Thiobacillus sp.

278

3.250

100

50

240

96

Thiobacillus denitrificans N.R.

250

286

95

Pleomorphomonas

Sulfur compound

S2O 3 FeS2 FeS2 Fluidized bed reactor

S0 S0 S2O3

Membrane reactor

S0 S0

GAC, granular activated carbon; NLR, nitrogen loading rate; NRE, nitrogen removal efficiency.

Ref. Zeng et al. (2021) Wang et al. (2021) Sahinkaya and Dursun (2015) Di Capua et al. (2020) Tong et al. (2017) Christianson et al. (2015) Zhang et al. (2015) Di Capua et al. (2017) Sahinkaya and Dursun (2015) Ucar et al. (2020)

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CHAPTER 8 Diversity of nitrogen-removing microorganisms

Unlike fixed bed reactors, fluidized bed reactors apply a flow with an upward velocity that allows expansion and fluidization of the supporting material, thus eliminating the problems of clogging and channeling caused by the excessive growth of the biofilm. On the other hand, the expansion of the bed makes it possible to obtain a greater surface for bacterial adhesion, thus increasing the reactive sites to carry out the process efficiently (Papirio et al., 2013). The fluidized bed reactor has a hydrodynamics that resembles that of a fully mixed reactor. Thanks to this, in this type of reactor an efficient contact between the biomass and the substrates is achieved, the accumulation of intermediate compounds like nitrite is reduced, a high treatment capacity is achieved and backwashing is not necessary (Di Capua et al., 2015). On the other hand, a disadvantage of this technology is that the start-up times are high, mainly due to the fact that it requires time for the biofilm to establish on the support, otherwise when operating at high up-flow velocity the biomass would end up being completely washed out of the system. Finally, membrane reactors are mainly based on the use of filter media that retain both the denitrifying biomass and the substrates, in order to maximize mass transfer and thus increase the efficiency of the process. Biological contamination is the main problem that affects the performance of this type of reactors together with the excess growth of biofilms and the formation of mineral precipitates that negatively affect the useful life of the membranes. The latter and high capital and operating costs of these types of reactors represent a considerable limit to the widespread application of this technology (Gede Wenten et al., 2020). Regarding the performance offered by each technology, as can be seen in Table 8.4, in general, the fluidized bed reactor is capable of treating higher nitrogen loading rates compared to other types of reactors. For example, when elemental sulfur is used as electron donor it is possible to appreciate that the fluidized bed reactor is capable of treating almost twice the nitrogen load compared to the membrane reactor and more than 11-fold in comparison with the fixed bed reactor.

8.2.6 Relevant operating conditions in the SDAD process to treat wastewaters The SDAD as well as other bioprocesses performed at full-scale need specific operational conditions that ensure the proper efficiency of the process. Because ADSOB is the functional core of the process, the operating conditions applied to reactors need to be in the range of physiological capacity of these bacteria. As seen previously, pH is a relevant parameter that directly affects the SDAD process, because of this, pH needs to be maintained in the range of 7.5 and 8. However, apart from this variable and in order to obtain a high activity level of both sulfur oxidizing and denitrifying bacteria, it is also important to apply the optimum nitrogen to sulfur molar ratio (N/ S). N/S is important not only for maintenance of conditions with maximum denitrification rate, but also due to the fact that at values above the optimum an accumulation of nitrites is observed and can inhibit the activity of bacterial community in the reactor (Marconi et al., 2017). Dolejs et al. (2015) reported that sulfide oxidation

8.2 Microorganisms using sulfur compounds as electron donor

efficiency reached 100% at a N/S ratio of 0.6 with no presence of intermediary compounds. In contrast, at lower N/S values with excessive amount of sulfides, reduced intermediate products like sulfur are formed, that is, the oxidation process is not completely carried out. Can-Dogan et al. (2010) accomplished a series of experiments with real wastewater containing both nitrates and sulfides, which was obtained from wastewater treatment plant treating industrial wastewater (Can-Dogan et al., 2010). Activated sludge adapted for 52 days to nitrates and sulfides was used as inoculum. The presence of nitrates ensured the oxidation of 90% of sulfides. However, the molar ratio 2 NO had a significant impact on the nitrate removal efficiency. Given the molar 3 /S 2 ratio NO ¼ 4.5, nitrogen removal efficiency was 25% at nitrate loading rate of 3 /S 3 1 2 1.039 kg m d ) and decreasing the molar ratio NO to 0.7 increased the effi3 /S ciency of nitrate removal up to 98% while maintaining the sulfide removal efficiency without accumulation of nitrites. In this case, the final product of sulfides oxidation was elemental sulfur. If sulfide oxidation was not limited by lack of nitrates and both nitrates and sulfides were in stoichiometric ratio, complete oxidation of sulfides to sulfates occurred.

8.2.7 Projections of using the SDAD process to remove nitrogen in wastewaters Although the advantages of sulfur-based autotrophic denitrification in wastewater treatment are well recognized, there are still several unsolved questions related with the reaction mechanism and the microbial community involved, which make difficult to massively implement this process at full scale. Some of these questions include how different operating conditions affect autotrophic denitrification, including the dynamics and activity of microorganisms present in reactors and the availability and stability of various sulfur substrates, and how the final products or intermediates affect the overall process performance (Fajardo et al., 2014). As discussed before, among many factors, pH has been found to be crucial in affecting the efficiency of sulfur-driven autotrophic denitrification. For the treatment of municipal wastewater, pH is not a big concern because this is quite stable and near the optimal range to perform SDAD process, thereby this technology is envisioned as an attractive treatment option for effluents with low organic carbon, for example, those produced in activated sludge process (as a tertiary treatment) or in anaerobic digesters (Zhu et al., 2019). On the other hand, the estimation of real possibility to carry out SDAD process in industrial wastewaters could be challenging, since many of these wastewaters that possesses considerable amount of nitrogen compounds are acidic, specifically, those generated in mine industry, ammunition industries/labs, or pharmaceutical industries, and until now there are very limited studies focusing on how SDAD processes perform under acidic conditions and how this could be improved without impairing the economic feasibility of implementing this process at full scale (Huang et al., 2019).

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CHAPTER 8 Diversity of nitrogen-removing microorganisms

8.3 Nitrogen removal by microorganisms that use hydrogen as electron donor: Hydrogenotrophic denitrification 8.3.1 Nitrate removal pathway and hydrogen as electron donor During hydrogenotrophic denitrification, hydrogen is used as an alternative energy source to organic matter to perform the nitrate reduction to gaseous nitrogen (N2) and water, using CO2 or H2CO3 as carbon source (Belmonte et al., 2017; Karanasios et al., 2010). Eq. (8.5) represents the sum of reduction reactions catalyzed by four reductases, producing the following intermediates nitrate ! nitrite ! nitric oxide ! nitrous oxide ! nitrogen gas (Albina et al., 2019).  2NO 3 + 5H2 ! N2 + 4H2 O+ 2OH

(8.5)

In this process nitrate respiration involves the electron transfer from hydrogen to nitrate as the last electron acceptor in the respiratory chain. This generates a proton gradient across the bacterial cell membrane, allowing the production of energy in the form of ATP by ATP synthase (Albina et al., 2019). Hydrogen is included in the denitrification respiratory chain through hydrogenases, which oxidize hydrogen to produce electrons and protons. These electrons are used by bacteria to regenerate coenzymes, such as NAD+ and then, the electrons are transferred through the respiratory chain. Here, three types of electron transporters are typically used: (i) Coenzyme Q or ubiquinone, (ii) Cytochrome bc1 complex, and (iii) Cytochrome c. Finally, these transporters are responsible for interacting with the reductases that transform nitrate to nitrogen gas (Albina et al., 2019). Hydrogenases are redox enzymes involved in both the production and consumption of hydrogen. These enzymes have an iron ion in the active center and may also be associated with a second iron or nickel ion: [Fe], [NiFe], [FeFe]. The latter two are the most commonly found in bacterial species, while [Fe]-hydrogenase has only been identified in methanogenic archaea (Albina et al., 2019; Winkler et al., 2013). Hydrogenotrophic bacteria assimilate inorganic carbon for growth through carboxylase enzymes from CO2 or H2CO3. Independently of the pathway used by the microorganisms, this assimilation has an energy expense in the form of NADH or ATP. Six microbial pathways have been identified to date: (i) reductive pentose phosphate (Calvin-Benson) cycle, (ii) reductive acetyl-CoA (Wood-Ljungdahl) pathway, (iii) reductive citric acid cycle, (iv) the 3-hydroxypropionate bicycle, (v) dicarboxylate/4-hydroxybutyrate cycle, and (vi) 3-hydroxypropionate/4hydroxybutyrate cycle (Albina et al., 2019; Ghafari et al., 2009).

8.3.2 Microorganisms and microbial community involved in the process Denitrifying bacteria are usually facultative anaerobes and in the presence of oxygen denitrification is inhibited. Bacteria capable of hydrogen oxidation are not abundant in nature and generally belong to phylum of Proteobacteria (Karanasios et al., 2010).

8.3 Microorganisms using hydrogen as electron donor

Simple and complex microbial cultures have been used to carry out hydrogenotrophic denitrification. For example, pure culture of Alcaligenes eutrophus (Chih et al., 1999) and Rhodocyclus sp. (Smith et al., 2005). However, working with a mixed microbial community, such as activated sludge or anaerobic sludge, has been much more widely used, but there are few studies focusing on species identification in microbial communities. In general, the limited environmental conditions of no oxygen and no organic matter together with the availability of molecular hydrogen allow the hydrogenotrophic species selection (Szekeres et al., 2002; Rezania et al., 2005; Karanasios et al., 2011). In this context, the literature has reported the selection of different microbial species depending on the systems used, operational parameters, reactor configuration, and inoculum source. Genera commonly found in hydrogenotrophic reactors include Paracoccus, Thauera, Hydrogenophaga, Rhodocyclus, Acinetobacter, Aeromonas, Marinobacter, and Pseudomonas (Karanasios et al., 2010; Komori and Sakakibara, 2008). For instance, Szekeres et al., 2002 reported that initially the microbial community was dominated by Ochrobactrum anthropic, Pseudomonas stutzeri, Paracoccus panthotrophus, and Paracoccus denitrifican. But, the first three were dominant at the end of the operation and their relative abundance was correlated with their specific denitrifying activity (Szekeres et al., 2002). Wang et al., 2015 reported that the microbial community was dominated by Proteiniclasticum, Gemmata, Planctomyces, and Hyphomicrobium, after hydrogenotrophic denitrification using anaerobic activated sludge as inoculum source (Wang et al., 2015). Zhu et al., 2017 reported the enrichment of the genera Simplicispira, Thauera, Thermomonas, Azoarcus, and Ottowia, in the cathodic chamber of a hydrogen-producing bioelectrochemical reactor (Zhu et al., 2017). Whereas in another study focusing on the combined removal of nitrate and perchlorate, the microbial community was dominated by Marinobacter hydrocarbonoclasticus (Van Ginkel et al., 2010).

8.3.3 Basis of operational conditions The main factors affecting hydrogenotrophic denitrification include nitrate concentration, hydrogen concentration, and system pH (Belmonte et al., 2017). These operating parameters must be properly adjusted to avoid accumulation of intermediate products, which could negatively impact microbial growth and process performance (Tian and Yu, 2020). The impact of nitrate concentration has not been systematically studied and apparently contradictory results have been reported (Belmonte et al., 2017). However, working with concentrations lower than 130 mg NO3 L1 would avoid inhibition of the process and nitrite accumulation (Zhou et al., 2007). High nitrate concentrations have been reported to cause nitrite accumulation, which inhibits denitrification and can even be toxic to bacteria (Albina et al., 2019). Stoichiometrically, 80 mg H2 are required to reduce 1 g nitrate, but experimentally up to 15% more H2 may be required (Belmonte et al., 2017; Albina et al., 2019). As a result, hydrogen concentration in the reactor should be kept higher than

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0.2 mg H2 L1 to avoid inhibition of the enzymes in charge of nitrate reduction, while values between 0.4 and 0.8 mg H2 L1 are considered optimal (Belmonte et al., 2017; Karanasios et al., 2011). Maintaining optimal hydrogen concentrations in the medium could limit denitrification due to its low solubility (0.74 mM; 1 atm, 30 °C) (Albina et al., 2019). To overcome this, different reactor configurations have been evaluated and will be discussed later in this chapter. The pH affects the functioning of all enzymes and consequently could generate an imbalance in the activity of the four reductases involved in denitrification (Albina et al., 2019). Thus the optimum range is considered to be between 7.0 and 8.6 (Ghafari et al., 2010), since lower values could promote the activity of homoacetogenic bacteria, which use H2 and CO2 to produce acetate (Tian and Yu, 2020). In contrast, higher pH values could lead to an increase in nitrite accumulation and decrease the efficient reduction of nitrate to N2 (Rezania et al., 2005). Some strategies have been successfully used to keep the pH of the system constant by buffering the released OH ions, such as the application of a phosphate buffer or CO2 bubbling (Xing et al., 2020).

8.3.4 Possibilities and available technologies for large-scale application Hydrogenotrophic denitrification has attracted attention for application in the production of drinking water from groundwater as well as for the treatment of wastewater with low organic matter content and high nitrate concentrations. This is because the process is clean with low biomass production and no generation of organic carbon residues (Karanasios et al., 2010). In this context, subway deposits of radioactive waste could cause nitrate permeation, generating water contamination with high nitrate concentrations ranging from 10 mM to 1 M. While other industries could generate nitrate-rich wastewater such as stainless steel plants (80 mM) and explosives production (500 mM) (Albina et al., 2019). This technology has been studied using mainly two configurations with respect to biomass growth: suspended cell and forming biofilms. These latter are preferred because they allow efficient use of autotrophic bacteria due to their low growth rate (Karanasios et al., 2010). Biofilm support materials offer systems with a high specific surface area, generating high biomass concentrations and allowing working with higher nitrate concentrations (Di Capua et al., 2015). In fact, several studies have reported higher nitrate removal rates in biofilm reactors compared to suspended biomass reactors (Ghafari et al., 2009; Rezania et al., 2005; Ghafari et al., 2010). In addition, systems with biofilms have been reported to require lower investment and operating costs (Di Capua et al., 2015; Karanasios et al., 2010). In this context, four types of configurations have been mainly used: packed bed reactor (PBR), fluidized bed reactor (FBR), membrane biofilm reactor (MBR), and biofilm electrode reactor (BER). The PBR considers biofilm growth on a fixed support material, where different materials such as polypropylene (Gros et al., 1998), polyurethane, granulated

8.3 Microorganisms using hydrogen as electron donor

activated carbon, sand, iron mixed with sand (Westerhoff and James, 2003), steel wool mixed with sand, silicic gravel (Karanasios et al., 2011; Vasiliadou et al., 2009), and bio-ceramsite have been successfully used (Di Capua et al., 2015; Karanasios et al., 2011; Chen et al., 2014). In these cases, hydrogen has been supplied mainly from an external storage tank into the reactor by bubbling or through gas permeable membranes (Lu et al., 2009). Another strategy applied to supply hydrogen has been its direct production inside the reactor through electrolysis of water and anoxic corrosion of metallic iron (Di Capua et al., 2015; Smith et al., 2005; Szekeres et al., 2002; Karanasios et al., 2011; Sunger and Bose, 2009). Regarding the reported operating conditions for PBR, volumes vary between 0.27 and 7 L, HRTs between 0.16 and 374 h, but most investigations have been between 1 and 3 h, while temperature has generally been ambient conditions (18–27°C). Research has worked with nitrate concentrations between 66.4 and 190.4 mg NO3 L1 and denitrification rates between 0.89 and 1.51 kg NO3 m3d have been obtained (Di Capua et al., 2015). In particular, a higher nitrate removal rate was obtained by increasing both hydrogen feed flow rate (from 20 to 100 mL min1) and HRT (from 10 to 120 min), reaching a removal of 10.7 kg NO3 m3 d (Lee et al., 2010), as presented in Table 8.5. The FBR and in contrast to PBR considers biofilm growth on a fluidized support material. The operation is comparable to a stirred tank reactor, increasing the contact efficiency between biomass and substrate, increasing the mass transfer of nitrate through the biofilm which helps to control the biofilm thickness (Di Capua et al., 2015). Limited applied research on hydrogenotrophic denitrification using FBR has been reported, where sand, polyacrylamide-alginate copolymer (Chih et al., 1999), and polyvinylalcohol porous cubes have been used as support for biomass growth. Strategies for supplying hydrogen have included direct bubbling in the Table 8.5 Case examples of hydrogenotrophic denitrification application using different reactor technologies. Type of reactor— volume

Nitrate feed (mg NO3 L21)

Operational conditions: HRT— temperature

Denitrification rate (kg NO3 m23 d)

Packed bed reactor—2.5 L Fluidized bed reactor—2.2 L

97.4

2 h—23  1°C

10.7

93–420.6

1 h—22°C

95.6

Membrane biofilm reactor—14 L Biofilm electrode reactor—0.6 L

59.3

0.06 h— 27.7  1.9°C

20.8

66.4

0.33 h—ambient

1.7

Ref. Lee et al. (2010) Komori and Sakakibara (2008) Tang et al. (2010) Prosnansky et al. (2002)

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reactor, through the recirculation line and direct production through electrolysis of water (Chih et al., 1999; Komori and Sakakibara, 2008). Regarding the reported operating conditions the volumes used are from 0.72 to 2.2 L, HRTs between 0.88 and 6.7 h, and the temperature between 22°C and 30°C. Research has worked with nitrate concentrations between 93 and 420.6 mg NO3 L1 and denitrification rates between 0.58 and 95.6 kg NO3 m3 d have been obtained (Di Capua et al., 2015). In MBR, hydrogen flows at high pressure through the membrane lumen and diffuses through the inner wall of a hydrogen-permeable membrane into the biofilm formed on the outer surface of the membrane. Most commonly used membranes have been hollow fiber membranes manufactured from materials such as polyethylene/ polyurethane, polyethylene, and polyester. Hollow fiber membranes can be operated as dead-end or flow-through, depending on whether the distal end of the membrane is closed or open, respectively (Di Capua et al., 2015; Martin and Nerenberg, 2012). Dead-end operation allows full hydrogen utilization, but its operation can cause nitrogen gas produced by the external biofilm to diffuse into the membrane lumen, decreasing the hydrogen concentration and affecting process performance. This challenge can be solved by a flow-through operation, but this involves higher energy costs and hydrogen consumption (Martin and Nerenberg, 2012). A major drawback of this type of reactor is the biofilm thickness, since increasing the thickness also increases the limitation of nitrate mass transfer (Martin and Nerenberg, 2012). Regarding the reported operating conditions, the volumes used are from 0.02 to 14 L and HRTs between 0.06 and 194.4 h. Research has worked with nitrate concentrations between 22.1 and 885.4 mg NO3 L1 and denitrification rates between 0.09 and 31.9 kg NO3 m3 d have been obtained (Di Capua et al., 2015). During BER operation, hydrogen is produced at the cathode by electrolysis of water. At the same time, this hydrogen is directly consumed by the hydrogenotrophic denitrifiers while forming a biofilm on the electrode surface. To promote biofilm formation on the cathode surface, the electrode material should preferably be carbon based (Di Capua et al., 2015; Zhou et al., 2007). However, studies using electrodes based on metallic materials such as stainless steel have also been successfully applied by denitrification mainly dominated by planktonic bacteria (Di Capua et al., 2015). As for the current intensity required for hydrogen production, it has been reported that between 5 and 200 mA is adequate to completely remove nitrate. Within this range, it has been observed that a higher current intensity increases nitrate removal (Wang et al., 2020). Whereas current intensities >200 mA lead to an overproduction of hydrogen and consequently an inhibition of denitrification. Additionally, the current intensity is strongly influenced by the configuration of each BER, where a larger cathode surface area also increases denitrification rates (Di Capua et al., 2015). The main disadvantage of these systems is the control of biofilm growth on the cathode surface as it limits the production and diffusion of hydrogen (Di Capua et al., 2015). Regarding the reported operating conditions the volumes used are from 0.2 to 19 L, HRTs between 0.33 and 50 h, and the temperature between 20°C and 40°C. Research has worked with nitrate concentrations between 44.3 and 221.4 mg NO3 L1 and denitrification rates between 0.044 and 1.7 kg NO3 m3 d have been obtained (Di Capua et al., 2015).

8.3 Anaerobic nitrate-dependent methanotrophic microorganisms

8.4 Nitrogen removal by anaerobic nitrate-dependent methanotrophic microorganisms 8.4.1 Nitrogen removal pathways and ecosystem distribution of the different types of microorganisms Anaerobic oxidation of methane (AOM) was first observed in the marine sediments and linked to simultaneous sulfate reduction (Knittel and Boetius, 2009). In the marine environments sulfate-dependent AOM contributes significantly to the control of the methane flux into the atmosphere. Approximately 90% of the methane emitted from the ocean bed is consumed. The responsible microorganisms were identified as archaea, often performing sulfate-dependent AOM in a consortia with sulfatereducing bacteria (SRB) (Hinrichs et al., 1999). These anaerobic methanotrophic archaea (ANME) are prevalent not only in the marine environment where S-AOM occurs, but as well in freshwater environments such as rivers, lakes, rice fields, and wastewater treatment plants (Vaksmaa et al., 2017a). However, as sulfate concentrations in the freshwater environment are much lower, different clades of AOM archaea have shown the ability to use alternative electron acceptors, such as nitrate or nitrite and also metals (including Fe(III) and Mn(IV)) (see Eqs. 8.6–8.8) (Haroon et al., 2013; Cai et al., 2018). Electron acceptors such as nitrate, nitrite, and iron mainly drive the AOM process in freshwater environments. Nitrate- or nitrite-dependent anaerobic methane oxidation (nitrite-AOM or nitrate-AOM, commonly referred to as N-AOM) is a microbial process performed by two groups of microorganisms such as bacteria and archaea. Nitrite-dependent AOM was the first process identified by enrichment of bacterium of NC10 phylum, named “Candidatus Methylomirabilis oxyfera” (N-damo bacteria) (Ettwig et al., 2008). This bacterium couples anaerobic methane oxidation to nitrite reduction to nitrogen gas. Furthermore, this bacterium has anaerobic and aerobic metabolic pathways (Ettwig et al., 2009). The genome harbored the complete aerobic pathway to oxidize methane, including the pmoCAB operon encoding the particulate methane monooxygenase (pMMO) complex. In anoxic conditions, it was postulated that this bacterium performs an intraaerobic process in which oxygen is produced via a putative nitric oxide dismutase enzyme (NO decomposition, 2NO ¼ N2 + O2) and uses the methane monooxygenase complex for the methane oxidation pathway via the particulate methane monooxygenase complex (pMMO) producing methanol and water. “Candidatus Methylomirabilis oxyfera” do not require a syntrophic partner to carry out nitrite-dependent AOM (Ettwig et al., 2009). Phylogenetically NC10 phylum is divided into five clades (group A, B, C, D, and E) (He et al., 2015; Wang et al., 2018a; Wang et al., 2018b). To date, mainly representatives of clade A of the NC10 phyla have been enriched in lab-scale experiments and possessed the ability to carry out nitrite-AOM process, and group B, although closely related to group A, has mainly been detected in environmental studies (Wang et al., 2018b; Zhang et al., 2020; Zhu et al., 2012). Whereas group A and B have been mainly detected in freshwater habitats, the C, D, and E groups have been mainly detected in saline

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ecosystems (Zhang et al., 2020). Since the discovery of the Methylomirabilis genus, several species have been characterized in laboratory-scale enrichments, such as Methylomirabilis sinica (Zhou et al., 2019), Methylomirabilis limnetica (Graf et al., 2018), and Methylomirabilis lanthanidiphila (Versantvoort et al., 2018). Nitrate-dependent anaerobic methane oxidation was first confirmed with the discovery of an archaeon of ANME-2d clade, enriched in a bioreactor supplied with methane, ammonium, and nitrate (Haroon et al., 2013). This archaeon, named “Candidatus Methanoperedens nitroreducens” (N_damo archaea) harbors pathways for the complete reverse methanogenesis and nitrate partial reduction, resulting in methane conversion to carbon dioxide and nitrate reduction to nitrite and ammonia (see Eqs. 8.6, 8.7 involved) (Haroon et al., 2013). + 5CH4 + 8NO 3 + 8H ! 4N2 + 5CO2 + 14H2 O Nitrate  dependentAOM

(8.6)

+ 3CH4 + 8NO 2 + 8H ! 4N2 + 3CO2 + 10H2 O

(8.7)

Nitrite  dependentAOM

2+ CH4 + 8FeðOHÞ3 + 15H+ ! HCO 3 + 8Fe + 21H2 O

Fe  dependent AOM

(8.8)

“Candidatus Methanoperedens nitroreducens” (anaerobic methanotroph, ANME2d clade, N-damo archaea) is an archaeon of Methanosarcinales order that is distantly related to marine methanotrophic archaea. This archaeon can couple anaerobic methane oxidation with nitrate reduction via “reverse methanogenesis” pathway confirmed by metagenomic and metatranscriptomic analyses and genome harboring respective genes for these reactions. Thus methane is activated by the activity of the enzyme methyl-CoM reductase and subsequently undergoes full oxidation to carbon dioxide via reverse methanogenesis (Arshad et al., 2015) and nitrate reduction to nitrite is catalyzed by an unusual Nar-like protein complex (Arshad et al., 2015). However, the subsequent denitrification steps are not carried out, nor have respective genes found in the genome, indicating that the N-damo archaea cannot reduce the produced nitrite further to NO, N2O, or N2 (Guerrero Cruz et al., 2018; Lu et al., 2019). Generally, Methanoperedens nitroreducens, using nitrate, typically forms a syntrophic relationship with Methylomirabilis oxyfera (see Fig. 8.2) (Raghoebarsing et al., 2006; Vaksmaa et al., 2017a), which is responsible for nitrite removal. The syntrophic relationship provides Methylomirabilis oxyfera with necessary electron acceptor and simultaneously removes it, as nitrite is a potential inhibitor of Methanoperedens nitroreducens (Costa et al., 2019; Haroon et al., 2013; Ettwig et al., 2009; Guerrero Cruz et al., 2018; Raghoebarsing et al., 2006). Besides, Ndamo archaea have shown the ability to use versatile electron acceptors for methane oxidation. Methanoperedens nitroreducens could switch from nitrate to using iron and manganese in short-term incubations (Ettwig et al., 2016). Revealing extracellular electron transfer strategies facilitated by multiheme c-type cytochromes commonly encoded by Methanoperedens nitroreducens (Haroon et al., 2013; Arshad et al., 2015). Furthermore, novel species Methanoperedens ferrireducens, Methanoperedens manganireducens, and Methanoperedens manganicus only carrying out metal-dependent AOM have been characterized.

8.3 Anaerobic nitrate-dependent methanotrophic microorganisms

FIG. 8.2 N-damo mechanisms, carrying out nitrite and nitrate-dependent AOM and the syntrophic relationship between Methylomirabilis oxyfera (N-damo bacteria) and Methanoperedens nitroreducens (N-damo archaea).

In nature and man-made ecosystems, these organisms have been found in diverse environments such as freshwater environments where nitrate is available in anoxic zones, such as in paddy soil, wetlands, lake sediments, anammox sludge, and wastewater sludges (between 50 and 100 cm depth) where water nitrate concentration is higher than 10 mg N L1 in the presence of methane at 30–40 mg CH4 L1 in the sediment (Lu et al., 2019; Islas-Lima et al., 2004; Wang et al., 2012) and recently the presence of NC10 phylum bacteria was found in a saline environment (Wang et al., 2021; Zhang et al., 2020; Zhou et al., 2019). Deutzmann et al. (2014) have suggested that N-AOM process can be the major methane sink in stable, nitrate-rich environments.

8.4.2 Activity and factors affecting the enrichment of these microorganisms Enrichment of nitrate and nitrite-dependent AOM microorganisms Methylomirabilis oxyfera and Methanoperedens nitroreducens is a notoriously slow process directly linked to the doubling times of these microorganisms. Whereas aerobic methanotrophs may have doubling times of hours minutes even, AOM microorganisms have it in the range of days (Guerrero Cruz et al., 2018) to several weeks. Obtaining a highly enriched N-damo culture (generally over 10% with respect to all microorganisms) depends on the inoculum source for enrichment, bioreactor operation, specifically media composition and trace elements and enrichment time (Hu et al., 2009). Reports ranging from few months up to years of operation demo in laboratory exist (Vaksmaa et al., 2017a; Ettwig et al., 2009; Guerrero Cruz et al., 2018; Raghoebarsing et al., 2006; Hu et al., 2009). Commonly in these reports where higher enrichments have been achieved, the inoculum originates from an existing enrichment of more than a year and not from an environmental source (Ettwig et al., 2009; Guerrero Cruz et al., 2018; Lu et al., 2019; Hu et al., 2014; Ding and Zeng, 2021). To date very few enrichments above 50% have been reported for Methanoperedens nitroreducens (Guerrero Cruz et al., 2018; Ding and Zeng, 2021) with to

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date no pure culture available to perform psychological experiments without the interference of N-damo bacteria or Anammox bacteria (the common partners in N-damo enrichments). He et al. (2016) isolated Methylomirabilis sinica belonging to group A of the NC10 phylum through the microcolony technique, highlighting the possibility of isolating Methylomirabilis bacteria via alternative method than enriching it. Ding et al. (2017) studied the possibility of uncoupling N-damo archaea from N-damo bacteria and they observed an increase of N-damo archaea from 24.29% to 65.77% and decrease of bacteria from 24.39% to 2.07% using a microbial fuel cell (MFC) with methane as the electron donor during 50 days of operation. However to apply N-damo microorganisms in WWTPs or biotechnological applications, optimal environmental condition information of pure culture is crucial. Environmental factors driving the variability of these microorganism are pH, carbon contents, salinity, dissolved oxygen, temperature, CH4/N ratio, ammonium and nitrite nitrogen, and other macro and micro nutrients. These factors have been reported to play important roles in structuring the community (Zhu et al., 2012). Supplying the bioreactors with nitrite will lead to enrichment of Methylomirabilis bacteria (Ettwig et al., 2009; Raghoebarsing et al., 2006). These have shown to operate at temperatures up to 30°C; however, at temperatures over 35°C damo activity is reduced. Also, it was observed that amount of Methanosarcinales archaeon decreases in enrichments over time, when nitrite is the sole electron acceptor provided. Reports exist showing that in subsequent upscaled enrichment culture these archaea are no longer detectable after incubation for 15 months (Ettwig et al., 2009). In case nitrate is provided as nitrogen source, Methanoperedens archaea would be enriched (Haroon et al., 2013; Vaksmaa et al., 2017a). However, as the intermediate products of nitrite and ammonia are produced, Methylomirabilis bacteria would be coenriched as well. Depending on the reactor operation, advantage can be given to Methylomirabilis bacteria or Anammox, although to obtain a pure culture is more laborious. pH is another significant factor for enrichments, as the Methanoperedens archaea activity first increases when pH is lower than 7.5, and then decreases when pH is higher than 7.5 (Ding and Zeng, 2021). Xu et al. (2018), reported that the optimal pH is 7.6, but the process could also exist at higher pH environment (pH 9.24). Bioreactors for enrichments of Methanoperedens have been operated at pH 7 as well as 7.5 (Vaksmaa et al., 2017a; Guerrero Cruz et al., 2018; Lu et al., 2019; Yu et al., 2017) and stable cocultures of Methanoperedens archaea, Methylomirabilis bacteria, and Anammox at pH 7.3 (Stultiens et al., 2019). In turn, to enrich an aqueous culture, it is reported that the stimulation by nitrite is more significant than the one by nitrate (growth rate of 0.002 gcells g1 COD methane) (Ettwig et al., 2009; He et al., 2015; Raghoebarsing et al., 2006; Hu et al., 2009; Yu et al., 2017; Shen et al., 2017). He et al. (2013) reported a nitrite inhibition constant of 57.4  7 mg N L1 for N-damo bacteria at 30°C. Besides, iron content is confirmed to be a key factor for the competition between the Anammox and N-AOM processes (Lu et al., 2019). For Methanoperedens archaea, containing high number of cytochromes, increasing iron levels in the media has shown to yield higher enrichment levels in comparison to original enrichment (Guerrero Cruz et al., 2018). Potential

8.3 Anaerobic nitrate-dependent methanotrophic microorganisms

application of either Methanoperedens archaea or Methylomirabilis bacteria depends on their oxygen tolerance as well. Although Methanoperedens archaea have been shown to be sensitive to oxygen exposure (Wang et al., 2018b; Guerrero Cruz et al., 2018), their ability to recover after oxygen stress is to date unknown. About methane, affinity constant of Methylomirabilis oxyfera for methane is estimated to be 2.6 μM by measuring methane consumption by membrane inlet mass spectrometry (MIMS), which is comparable to the affinity constant of aerobic methane oxidizers (0.2–6 μM) (Guerrero Cruz et al., 2019). However, other authors have also measured the methane affinity rate by measuring methane in the gas phase of bioreactors obtaining higher values (92 μmol L1) (He et al., 2013). Therefore the effects of other environmental factors such as dissolved oxygen, COD, and methane pressure on microbes performing N-AOM should be studied in future (Ding and Zeng, 2021; Lee et al., 2018).

8.4.3 Molecular tools for assessing microbial diversity Authors have shown that nitrite and nitrate are key factors regulating the distribution of N-AOM microorganisms using specific PCR primers for their identification (16S rRNA and functional genes) (Wang et al., 2018c). Commonly environmental detection is based on 16S rRNA gene primers for both Methylomirabilis oxyfera and Methanoperedens nitroreducens. However, functional gene-based markers have been developed for both. Methylomirabilis oxyfera has a methane monooxygenase. (pMMO), which is a key gene transcribed and expressed during methane oxidation process. It has a highly conserved a-subunit encoded by the pmoA. Luesken et al. (2011) developed pmoA-based primers for Methylomirabilis oxyfera, as pmoA primers designed to capture aerobic methanotrophs contain several mismatches to the pmoA of Methylomirabilis oxyfera (Luesken et al., 2011). Methanoperedens nitroreducens contain the mcrA gene encoding the alpha subunit of the MCR enzyme (Ettwig et al., 2009; Guerrero Cruz et al., 2018) and this has successfully applied to detect N-damo archaea in enrichment cultures and environmental samples. McrA-based primers were developed for Methanoperedens nitroreducens (Vaksmaa et al., 2017b) and further modified and applied for mcrA-based amplicon sequencing to investigate environmental diversity of the Methanoperedens archaea (Xu et al., 2018). However, overall, further studies of the physiology and enzymes other than MCR of Methanoperdens archaea are still necessary (Ding and Zeng, 2021).

8.4.4 Application possibilities in sewage and industrial wastewater treatment plants—Main operating conditions description WWTPs are recognized as contributors to greenhouse gas emissions by releasing N2O and CH4 (Daelman et al., 2012), thus finding economically feasible and sustainable ways to treat wastewater are urgently required. The nitrite/nitrate-dependent anaerobic methane oxidation process provides a new way for simultaneous methane

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and nitrogen removal under anoxic conditions. WWTP are rich in carbon and nitrogen compounds. Methane in wastewater treatment plants is naturally produced by methanogenesis. This process is carried out by methanogenic archaea, using a plethora of available substrates such as hydrogen, acetate, CO2, or methylated compounds, and termed hydrogenotrophic, aceticlastic, and methylotrophic methanogenesis, respectively, depending on the electron donor used in the reduction (Liu et al., 2014). In methanogenesis, methane is produced: H2, acetate, or other methylated compounds, respectively. Methanoperedens archaea, Methylomirabilis bacteria, and Anammox bacteria have been reported to be present at anoxic compartments of the wastewater treatment systems (Xu et al., 2018; Stultiens et al., 2019). The most suitable application of Methanoperedens and Methylomirabilis microorganisms would be in cooperation with anaerobic ammonia-oxidizing bacteria, resulting in the simultaneous removal of nitrogen compounds (ammonium, nitrite, and nitrate) and methane. Anammox produces nitrogen gas and nitrate from ammonium and use nitrite generally from aerobic partial nitrification. At the same time, the produced nitrate can then be used by Methanoperedens archaea to convert it to nitrite and ammonia, feeding both Methylomirabilis bacteria and anammox bacteria with substrate. However, the applicability of a coupled Anammox and N-AOM system for wastewater treatment needs further research to understand how these two groups of microorganisms compete for substrate (Stultiens et al., 2019). As potential mediators, the applicability of these microorganisms in the same system to remove the nitrogen compounds and simultaneously remove methane has mainly been investigated in small-scale reactor systems (Arshad et al., 2015; Stultiens et al., 2019). Although theoretically feasible, the applicability of these microorganisms in large-scale WWTPs faces several challenges. These microorganisms have slow doubling times, as enrichments take from months to years. Specific systems to compensate the low biomass yield were used in laboratory-scale enrichments, e.g., reactors with biomass retention such as fixed bed, fluidized bed, and biofilm membrane reactors or fed-batch reactor systems have been used for enrichment (Ghafari et al., 2010; Ettwig et al., 2009; Raghoebarsing et al., 2006; Shen et al., 2017; van Kessel et al., 2018). From an engineering perspective, methane concentration is a critical factor in future applications. Methane has a very low solubility and would rapidly escape to the atmosphere (Nguyen et al., 2020). In bioreactors it has been demonstrated that a continuous supply of methane is essential for cultivation of these anaerobic methane oxidizers (Haroon et al., 2013; Ettwig et al., 2009; Guerrero Cruz et al., 2018; Stultiens et al., 2019). Thus, in WWTP, for application, a high methane concentration or sufficient partial pressure has to be secured. As long as a high methane concentration (or partial pressure) can be provided, there would be efficient nitrogen removal. The maximal methane concentration that can be achieved under normal pressure is around 20 mg CH4 L1, thus limiting the actual application in most initial studies with low nitrogen removal rate (Lu et al., 2019). On the availability of methane, in anaerobic wastewater treatment systems, methane will be present as dissolved gas in the bulk liquid and the improvement in this availability could be done using methane supply via a membrane system (Zhang et al., 2018).

8.3 Anaerobic nitrate-dependent methanotrophic microorganisms

The reactor configuration is another factor. B. Hu et al. (2014) compared different reactor configurations (stirred gas lift reactor (MSGLR), sequencing batch reactor (SBR), and continuously stirred tank reactor (CSTR)). They obtained maximum volumetric nitrogen removal rate of 76.9 mg N L1 d1 with MSGLR, as this kind of reactor enhances the mass transfer of gas-liquid phases and the mixing of liquid-solid phases compared to other reactor types. However, with an upflow continuous reactors and batch cultures and hydraulic retention time of 2.1 h, Hatamoto et al. (2014) obtained similar nitrogen removal rate (70.4  3.4 mg NN L1 d1). Ding and Zeng (2021) studied N-AOM process in an microbial fuel cell (MFC) using carbon fiber felt as anode and cathode and methane as the fuel, where the growth of Methanoperedens archaea was favored unlike damo bacteria. It demonstrated weak electrogenic capability with around 25 mV output, so further studies are still required. Another aspect is competition for substrate; anammox bacteria compete with Methylomirabilis oxyfera for available nitrite and have a higher affinity for nitrite than the N-damo bacteria. This may result in removal via outcompeting the Ndamo bacteria (van Kessel et al., 2018; Wang et al., 2017). Recently, the coculture system N-damo microorganisms and Anammox bacteria in a hollow-fiber membrane biofilm reactor (HfMBR) with high rate of methane mass transfer were investigated. The maximal removal rates of nitrate and ammonium were 78–85.33 mg N L1 d and 26–37.95 mg N L1 d1, respectively (Ding and Zeng, 2021; Fu et al., 2017). The highest nitrogen removal rates were achieved when the coculture formed granules (Fu et al., 2017). However, Shi et al. (2013) reported on Anammox bacteria and N-AOM microorganisms combination processes, nitrate and ammonium removal rates to be higher (190 and 60 mg N L1 d1, respectively). In comparison, Stultiens et al. (2019) with a coculture of Methanoperedens archaea, Methylomirabilis bacteria, and Anammox bacteria obtained in a sequencing batch reactor fed with methane, ammonium, and nitrite reported nitrite and ammonium removal rates of up to 455 mg N-NO2 L1 d1 and 228 mg N-NH4 L1 d1 and a nitrogen removal efficiency of 97.5%. In addition, coexistence of Methylomirabilis bacteria and methanogens in an UASB reactor fed with inorganic carbon (CO2/H2) was reported. The nitrite removal was 75% with nitrogen removal rates of 5.63  0.01 mg Ng VSS1 d1 (Ma et al., 2017). The configuration of the WWTP and compartmentalization will affect not only the above-mentioned microorganisms present. WWTPs contain a complex microbial community, which through next generation sequencing only recently has received more attention to characterize it (Guerrrero Cruz et al., 2021). However, for real applicability in WWTPs, several aspects of the AOM process and responsible microbes need to be still revealed. As no pure culture has been obtained for Methanoperedens archaea, the physiology has been described based on coenrichments. For Methylomirabilis bacteria, their ability to compete for nitrite should be better understood in a full-scale WWTP. Both organisms have shown stress reaction after oxygen exposure, which may occur in WWTP, however recovery after exposure is to date unknown.

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Acknowledgments O.F. is grateful for the financial support of Fondecyt-ANID postdoctoral No. 3210326. J.T.-A. is grateful for the financial support of Fondecyt-ANID postdoctoral No. 3210456. E.T.-V. is grateful for the financial support of Fondecyt-ANID project No. 11200211.

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CHAPTER

An overview of the anammox process

9

Yan Guo and Yu-You Li Department of Civil and Environmental Engineering, Graduate School of Engineering, Tohoku University, Sendai, Miyagi, Japan

9.1 Introduction In the face of the dual set of challenges posed by global warming and resource depletion, many changes are required for the sustainable development of societies around the world. Known for their high levels of energy consumption and for being resource rich, wastewater treatment plants (WWTPs) all over the world need to be improved technically to save energy and recover resources. The past 40 years have witnessed the development and expansion of the anammox-based processes both at lab-scale and in full-scale operations. There is no doubt that anammox-based processes can provide WWTPs with the option of being energy autarkic as well as many other remarkable advantages (Reino et al., 2018). The learning curve of the anammoxbased processes includes many different aspects, including microbial metabolism, process configuration, and engineering applications. However, since problems remain at both the cognition level and the technical level, a large proportion of wastewater treatment installations around the world continue to use the traditional nitrification and denitrification (TND) process or other high-cost processes. In recent years, anammox-based research has ushered in a spurt of development with significant findings reported from a large number of investigations. Therefore it is a matter of urgency that the results of the latest research are collated and processed to provide an overview of state-of-the-art knowledge for further advances and, ultimately, the popularization of this process (Guo et al., 2022). Since the discovery of the anammox phenomenon several decades ago, much effort has been dedicated to pursuing the ultimate goal of the smooth and reliable application of this technology over a broader scope (Guo et al., 2020a). The theoretical knowledge obtained from the research work on this process is extensive, from a chemical perspective, microbial perspective, engineering perspective, and process application perspective. From the perspective of actual applications, a large number of installations adopting anammox-based processes have been constructed and operated in an attempt to achieve a reliable energy-saving and operation cost-reducing system for widespread Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00005-X Copyright # 2022 Elsevier Inc. All rights reserved.

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CHAPTER 9 An overview of the anammox process

utilization in the wastewater treatment sector (Lackner et al., 2014). However, applications have largely been limited to the treatment of high ammonium-strength wastewater (HSWW), like biomass digestion effluent, landfill leachate, and the advantages of this process have been shown to be considerable indeed. On the other hand, in the treatment of low-strength wastewater (LSWW), performance of this process has been poor and unstable (Cao et al., 2017; Chen et al., 2019). Therefore the main challenge in actual applications is the treatment of LSWW, and new ways to overcome the bottlenecks are being vigorously explored (Chen et al., 2019). Even in the HSWW treatment field, the problems facing the anammox-based process, like long start-up time, the limited nitrogen removal efficiency (NRE), and the fragile resistance ability of microbes to external environment fluctuation are considered insurmountable barriers in the effective implementation of this process. These problems are caused partly by such extrinsic factors as the difficulty of operation and control and complex feedback conditions (Guo et al., 2020b), and also by such intrinsic factors as the low growth rate of anammox bacteria, the low anammox specific activity at low temperature, the easily inhibition of anammox bacteria at extreme environment (van de Graaf et al., 1996). It is vital to ensure the stable and reliable acquisition of both nitrite and ammonium, both of which are reactants in the anammox reaction in the anammox-based process (Ma et al., 2017). The intense focus on the combination of the nitriteproduced step and the anammox step has led to the development of various configurations and strategies (Deng et al., 2019; Zhang et al., 2018b). In actual applications, the nitrogen species in wastewater differ according to the source of the wastewater (Guo et al., 2020a). In practice, in the biomass digestion effluent and the landfill leachate, the main nitrogen species is ammonium, while in the fertilizer in runoff, and in explosive production wastewater, the main nitrogen species is nitrate. Therefore it is necessary to deal with the actual nitrogen species in the specific wastewater, and the pretreatment of influent is needed to balance the nitrite and ammonium supply in the application of the anammox process. Many anammox-based processes have been developed in combination with pretreatment processes: the partial nitritation/anammox (PNA) process (Qian et al., 2018), the autotrophic desulfurization/ anammox process, the denitratation/anammox process, the nitrate-dependent methane-dependent denitrification (DAMO)/anammox process (Shi et al., 2013), and the dissimilatory nitrate reduction by anammox bacteria/anammox process. Thus, with a focus on several of the aspects mentioned before, and with explicit consideration of the needs in the application of the anammox-based process, in this chapter, the three major anammox-based processes in actual application are reviewed. The related research achievements in recent years are also summarized.

9.2 The evolution of anammox reaction stoichiometry The correct knowledge of chemical stoichiometry of the biological reaction is crucial not only in scientific research but also in process design and operation control (Lotti et al., 2014). There have been several clear steps in the evolution of knowledge of the

9.2 The evolution of anammox reaction stoichiometry

chemical stoichiometry of the anammox reaction. In this section, the proposed reaction stoichiometries from the relevant research are listed, and the similarities and differences among them are discussed. As presented in Table 9.1, the first notable event in the development of anammox technology was the predicted existence of lithotrophic bacteria capable of catalyzing the anammox reaction in 1977 by Broda (1977). Based on thermodynamic calculations, and considering evolutionary changes and thermodynamical principles, Broda predicted chemosynthetic bacteria could oxidize ammonium to nitrogen with oxygen or nitrate as an oxidant. About two decades later, the current widely accepted anammox stoichiometry was proposed by Strous et al. (1998). In the research by Strous et al., the stable conditions of the sequencing batch (SBR) reactor used made it possible to carry out mass balance calculations under defined conditions. However, just 90% of the biomass retained with residual washout was used in the calculations of the composition of elements. It should also be noted that a high standard deviation was included in the set of data of dry weight measurements for the carbon balance calculation. The electron balance of the measured conversion rates used to calculate the stoichiometric equation also had an error of 15%. While the parameters given were probably adequate for rough anammox-based research and applications, it is also fair to assume that they would also have resulted in systematic inaccuracy in competition studies (Lotti et al., 2014). According to the chemical stoichiometry proposed by Strous et al., as presented in Table 9.1, the nitrate produced in the anammox reaction will have a stoichiometric NO3
-N/Nr (Nr referred to removed NH4 +-N and NO2
-N) ratio of about 11%. This nitrate formation was associated with the utilization of nitrite as the reductant by anammox bacteria for cell synthesis from carbon dioxide (Prevot et al., 2012; Strous et al., 1998). However, in some experiments with a long solid retention time (SRT), an NO3
-N/Nr ratio of lower than 11% has been reported. When using a membrane biological (MBR) reactor with an SRT of 500 days, Sun et al. (2018) achieved an average nitrogen removal rate (NRR) of 0.98 kg-N/m3/day with a 2.9% NO3
-N/Nr ratio. The authors proposed that a long SRT resulted in a decrease in the production of nitrate due to denitrification. The reduced net cell yield that results from long SRT is well known as a basic concept in environmental biotechnology. Further investigation is required to determine whether the low cell yield also contributes to this low ratio in these long SRT reactors. Lotti et al. (2014) revised the stoichiometry of the anammox reaction by performing kinetic experiments and analyzing the constituent elements in an MBR reactor with a highly enriched anammox microbial community. An almost pure free cells suspension of highly active anammox bacteria was used in a detailed kinetic and stoichiometric analysis of the anammox process. The chemical reaction stoichiometry was identified from the yield and the elemental composition of the biomass was validated by long-term reactor operations. The data-set used in the study by Lotti et al. to calculate the anammox stoichiometry through reconciliation was more consistent than that used by Strous et al. Notably, the error in the electron balance was smaller, and the carbon balance was considerably more accurate. Furthermore, instead of

167

Table 9.1 Summary of reported anammox reaction stoichiometries.

Reactor type

Reaction stoichiometry

-N ratio

NO3 2 N/ Nr ratio (%)

Thermodynamic calculations SBR

1NH+4 + 1NO
2 ! 1N2 + 2H2O

1.0

–

–

–

1.32

11.2%

87.9

0.049

1.146

7.5%

91.9

0.051

1.133

7.5%

91.9

0.103

1.30

11.4%

88.7

0.108

NO2 2 N/NH4 +

MBR UASB


1NH+4 + 1.32NO
2 + 0.066HCO3 + 0.13H+ ! 1.02N2 + 0.26NO
3 + 0.066CH2O0.5N0.15 + 2.03H2O +
1NH+4 + 1.146NO
2 + 0.071HCO3 + 0.057H ! 0.986N2 + 0.161NO
+ 0.071CH O 3 1.74 0.31N0.20 + 2.002H2O +
+ 0.092HCO + 0.038H 1NH+4 + 1.133NO
2 3
! 0.980N2 + 0.161NO3 + 0.092CH2.26O1.07N0.14 + 1.961H2O (NLR ¼ 5.0kg/m3/day) +
1NH+4 + 1.300NO
2 + 0.121HCO3 + 0.367H !
1.020N2 + 0.242NO3 + 0.121CH1.74O0.81N0.15 + 2.139H2O (NLR ¼ 50.0kg/m3/day)

Maximum NRE (%)

Biomass/ Nr ratio

Biomass/Nr: the mole mass of produced organic matter to the mole mass of removed ammonium and nitrite ratio according to reaction stoichiometry.

Reference Broda (1977) Strous et al. (1998) Lotti et al. (2014) Zhang et al. (2018a)

9.3 Existing problems and countermeasures for anammox process

using independent averaged rates, the data-set consisting of the original measurements obtained during long-term steady operations were even cross correlated during error propagation. According to the stoichiometry proposed by Lotti et al., the amount of nitrite removed relative to the amount of ammonium removed (NO2
-N/NH4 + -N) ratio is 1.146. Meanwhile, recent research further demonstrated that a molar ratio of around 1.0 resulted in good anammox behavior in the MBR reactor (Nawi and Stuckey, 2018). There are several recent reports which also showed close values of both the NO2
-N/NH4 + -N ratio and NO3
-N/Nr ratio in MBR bioreactor to that of the stoichiometry proposed by Lotti et al. Later, Zhang et al. (2018a) proposed a varied stoichiometry of the anammox process according to increases in the nitrogen loading rate (NLR) from 5.0 to 60.0 kg-N/ m3/day using an upflow anaerobic sludge blanket (UASB) reactor, and concluded that anammox stoichiometric equations changed as the loading rate increased. Their results indicated that a higher NLR was a key factor in increasing the cell yield, before an increase in the nitrate occurred. According to the stoichiometric equations proposed by Zhang et al. for the anammox process under the low NLR condition (5.0 kg-N/m3/day), the stoichiometric equation was similar to that proposed by Lotti et al., while under the high NLR condition (50.0 kg-N/m3/day), the stoichiometric equation was similar to that proposed by Strous et al. As suggested by the results cited before, the actual NO2
-N/NH4 + -N ratio and NO3
-N/Nr ratio of anammox reaction in anammox reactors varied according to the experimental configuration, operation condition, and time. Lotti et al. (2014) also proposed that the values of the parameters reported in the literature were often hampered by limitations of mass transfer and by the presence of a significant side population. Therefore the existing stoichiometric parameter values are meaningful only as a rough guide to the operation of the anammox process operation. However, in the case of a high NRE and from the perspective of complete depletion of ammonium and nitrite in actual applications, the proper adjustment of the influent NO2
-N/ NH4 +-N ratio according to the nitrogen species distribution in the effluent is a practical strategy.

9.3 The existing problems and countermeasures for anammox process application 9.3.1 The rapid start-up and recovery of anammox-based process With the proper conditions, including the dissolved oxygen (DO), temperature, pH, and a well-maintained residual ammonium concentration, the desired suppression of NOB and heterotrophic bacteria can be achieved. It has been confirmed that these conditions are crucial factors for creating a suitable environment for anammox sludge enrichment during the start-up of anammox process. However, the intrinsic low growth rate of anammox bacteria limits the rapid start-up of the anammox-based

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process. Besides, because of the frailty of anammox bacteria, its metabolism is easily inhibited when conditions are not optimal. It highlights the importance of efficient and reliable recovery strategies when the inhibition of anammox bacteria occurs in the anammox-based process application. Of course, if a large amount of anammox sludge is available, the startup of the anammox process can be accelerated. However, without the source, or with a limited source, it is clear that the enrichment of anammox sludge is inevitably a lengthy process. Many researchers have explored methods to accelerate the growth of anammox bacteria and enrichment from different aspects. Biomass concentrations as high as 73.2  8.7 g-VSS/L have been achieved in the high loading UASB reactor (Zhang et al., 2018a). Therefore, to a certain extent, a high loading rate anammox reactor can also function as an anammox sludge enrichment installation to inoculate pilot-scale or full-scale installations. Meanwhile, Guo et al. (2018) found that low nitrogen strength was beneficial to increase anammox biomass yield and nitrogen removal. Lotti et al. (2015) suggested that the maximum specific growth rate was not an intrinsic property and could be increased by optimizing the growth conditions (through increasing the biomass specific electron transfer capacity). Also, a startup with an MBR appears to be more promising for the enrichment of anammox bacteria due to the high sludge retention property of the MBR configuration. The appropriate feeding of ferrous iron was shown to be beneficial to the increase of heme c synthesis, the enhancement of hydrazine dehydrogenase activity, and the growth of anammox bacteria (Qiao et al., 2008). There are also several efficient countermeasures for the recovery of anammox bacteria from inhibition. Nitrite accumulation is the crucial and common factor resulting in the inhibition of anammox bacteria. For the sudden inhibition of anammox, the timely washing of sludge has been deemed an effective strategy to recover the activity of anammox sludge, and this method has often been adopted in our laboratory (Zhang et al., 2018a). A short duration dose of nitrate to anammox bioreactor for the rapid activity restoration of nitrite-stressed anammox bacteria was also reported. Appropriate amounts of nanoscale zero valent iron (short-term dosing) have been shown to rapidly recover anammox activity after the anammox bacteria were accidentally exposed to oxygen and also could be useful in facilitating the robust operation of anammox-based processes (Yan et al., 2019).

9.3.2 The retention of anammox sludge in the reactor Excellent biomass retention and a short hydraulic retention time (HRT) are both required to achieve a sufficiently high NRR. As stated earlier, the anammox UASB reactor can achieve a super high NRR with the sludge granules, and in the one-stage PNA process, with well-formed biofilm and granules, the desired NRR can be achieved. However, because both the biofilm formation and granule formation are time consuming, the startup time of the anammox-based process is prolonged (Ding et al., 2018). Therefore strategies for maintaining a high biomass concentration and good sludge retention rates in the reactor need further innovation. On the

9.4 The status of several main anammox-related processes

other hand, it has been reported that a biofilm that is too thick or granules that are too large are not well suited to the effective activity of anammox bacteria: this indicates the status of the biomass in the reactor is also important for improving the NRR of the anammox-based process. In recent years, biological-induced phosphate formation in anammox sludge has been proposed as a method for enhancing the settleability of anammox sludge to facilitate a high biomass concentration in the reactor, and a good performance was achieved (Guo and Li, 2020; Ma et al., 2018b; Zhang et al., 2018b). In our lab, even in the one-stage PNA reactor, through the biological-induced phosphate formation with the biomass, the highly dispersed suspended sludge with particle size below 1.0 mm was formed. Meanwhile, a good NRR was achieved for low-strength wastewater with this suspended sludge.

9.3.3 The further improvement of NRE Two factors affect the NRE of anammox-based processes. One is the deviations of the influent NO2
-N/NH4 + -N ratio in the anammox reactor, as stated earlier. This leads to the incomplete consumption of ammonium or nitrite, which makes it impossible to achieve the desired NRE close to the theoretical NRE of approximately 90%. Besides, because subtle excess nitrite results in the inhibition of anammox bacteria, the depletion of nitrite is a higher priority than ammonium depletion in an anammoxbased reactor operation. That also explains why ammonium and nitrate were the main nitrogen species in the effluent in the results of many investigations (Abma et al., 2007). Another factor is that nitrate is produced as the oxidation product for anammox cell synthesis from carbon dioxide (Prevot et al., 2012; Strous et al., 1998). Therefore two points emerge for the further improvement of the NRE: one is that the nitritation step needs to be carefully monitored for more precise control of the NO2
-N/NH4 +-N ratio based on the real-time monitoring of nitrogen species distribution in the effluent, and the other is nitrate removal (Abma et al., 2007). Some researchers have proposed the solution of setting predenitrification unit ahead of the anammox-based unit (Li et al., 2018): the effluent of anammox-based unit with a proper recirculation rate was returned to the predenitrification unit where the nitrate was removed by denitrification bacteria (DB) utilizing the COD or DAMO archaea using methane (Liu et al., 2020) in original wastewater. This configuration can realize nitrate reduction without external COD addition, amounting to significant savings in the treatment cost.

9.4 The status of several main anammox-related processes In most research, anammox reactors were operated without the nitrite-producing step, using both ammonium and nitrite-containing synthetic wastewater (Zhang et al., 2018b), or actual wastewater amended with nitrite, or even ammonium. This kind of process, referred to as “the pure anammox process” (Guo et al., 2020a), has played an essential role in exploring anammox technology.

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CHAPTER 9 An overview of the anammox process

For the ammonium wastewater, the nitritation process is reliable for nitrite production from part of the influent. The produced nitrite and the ammonium in the residual influent are prepared for the anammox step. Since the balance between nitrite and ammonium can be controlled, it is theoretically feasible to use a solo source of ammonium wastewater when adopting anammox-based process for nitrogen removal. In the case of nitrate wastewater, the denitratation process is also a reliable process for nitrite production. However, access to ammonium is a problem. In some instances with ammonium wastewater and nitrate wastewater in one location, the joint treatment of two kinds of wastewater seems feasible: for the nitrate wastewater, the nitrate can be converted into the nitrite, then the nitrite wastewater and the ammonium can converge, and finally the nitrogen species can be removed through the anammox process. In instances where there is only nitrate wastewater, the application of anammox process is limited because of the lack of access to ammonium. From a different perspective, these pretreatment processes can be divided into different categories. From the nitrite supply route, the pretreatment processes can be separated into the nitritation process and denitratation process (Ma et al., 2020a). From a conversion mechanism perspective, the pretreatment processes can be separated into a nitritation process, an organic matter-based denitratation process (Ma et al., 2017), a thiosulfate-based denitratation process, and a methane-based denitratation process (Shi et al., 2013). With the combination of anammox and these above-mentioned pretreatment processes, many anammox-based processes have been developed toward actual application. Based on the core mechanism and actual feasibility, three comprehensive and practical anammox-based processes have garnered most of the research attention: the PNA process (Chen et al., 2019), which targets at the HSWW and LSWW; the denitratation/anammox process, which targets at the simultaneous treatment of high-strength nitrate and HSWW (Ma et al., 2017), HSWW and LSWW; and the DAMO/anammox process (Shi et al., 2013), which targets at the simultaneous treatment of nitrate wastewater and HSWW, HSWW (Liu et al., 2019) and LSWW (Liu et al., 2020). In the DAMO/anammox process, because there is always DAMO bacteria in the reactor besides DAMO archaea and anammox bacteria, part of the nitrogen removed can be attributed to the DAMO bacteria (Xie et al., 2017). Among these, the PNA process was seen as the most presentative and practical since, among different types of actual wastewater, ammonium wastewater has the largest share of nitrogen species (Guo et al., 2020a). Since the performance of the nitratation/ anammox process and the DAMO/anammox process has improved considerably with further improvements in the total nitrogen removal efficiency in recent years, this process has been given more and more attention. The mechanisms of the earlier mentioned three process-based processes are shown in Fig. 9.1. A comparison of the three anammox-based processes is provided in Table 9.2. Since there are so many anammox-based processes, a summary of each process is necessary as a guide for configuration design and actual application. For this reason, these processes are reviewed in this section.

9.4 The status of several main anammox-related processes

FIG. 9.1 The mechanism of three anammox-based processes.

Table 9.2 A comparison of main anammox-based processes. Process

Substrate +

PNA

NH4 -N

DNA

BOD NO3
-N NH4 + -N CH4 NO3
-N NH4 + -N

DAMOA

Reaction

Microbe

Environment

Products

Nitritation anammox Denitratation anammox

AOB anammox DB anammox

Aerobic Anaerobic Anaerobic

DAMO anammox

DAMOB anammox

Anaerobic

N2 NO3
-N N2 CO2 NO3
-N N2 CO2 NO3
-N

9.4.1 Nitritation process The rapid startup and stable operation of the nitritation process are crucial in most biological nitrogen removal processes, including the TND process, the nitrite shunt process, and the two-stage PNA process. Nitritation is conducted by autotrophic ammonium-oxidizing bacteria (AOB) with the reaction, expressed by Eq. (9.1). The anammox reaction is conducted by autotrophic bacteria by either the reaction shown in Eq. (9.2) (Strous et al., 1998) or Eq. (9.3) (Lotti et al., 2014): The nitritation reaction: NH4 + 1:238O2 + 0:04HCO
+ 0:161CO2 ! 0:96NO
2 + 0:04C5 H7 NO2 + 0:919H2 O + 0:919H+ (9.1)
+
NH+4 + 1:32NO
2 + 0:066HCO3 + 0:13H ! 1:02N2 + 0:26NO3 + 0:066CH2 O0:5 N0:15 + 2:03H2 O

(9.2)
+
1NH+4 + 1:146NO
2 + 0:071HCO3 + 0:057H ! 0:986N2 + 0:161NO3 + 0:071CH1:74 O0:31 N0:20 + 2:002H2 O

(9.3)

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The nitritation process can be achieved by selecting the appropriate temperature, pH, substrate availability, and free ammonia (FA) and free nitrous acid (FNA) inhibition levels required to wash nitrite-oxidizing bacteria (NOB) from the system (Van Hulle et al., 2010). As presented in Table 9.3, the highest nitrite accumulation rate (NAR) reported was 9.0–12.0 kg-N/m3/day for the treatment of rare earth ammonium wastewater (Wenjie et al., 2016). However, in most of the relevant literature, the maximum NAR achieved was 5.2 kg-N/m3/day (Tora` et al., 2014). In the case of LSWW and the low-temperature condition, the NAR was considerably lower. Besides the limit to the maximum NAR, it was shown to be considerably difficult to obtain an effluent with a stable NO2
-N/NH4 +-N ratio in the case of the two-stage PNA process. Due to these two points, the nitritation step is considered the rate-limiting step for nitrogen removal in the PNA process. This will be discussed in more detail later (Liu et al., 2018b).

9.4.2 Pure anammox process In pure anammox research, like that in Table 9.4, it is possible to achieve an extremely high NRR. The highest value reported, at 76.7 kg-N/m3/day, was achieved for synthetic HSWW (Tang et al., 2011), while the achieved maximum NRR for actual HSWW was only 10.7 kg-N/m3/day (Yokota et al., 2018). In other words, there is a substantial gap in the performance of the anammox process between synthetic HSWW and actual HSWW. On the other hand, in the case of LSWW, regardless of whether the wastewater was synthetic or actual wastewater, the NRR was considerably lower than that for HSWW. Thus further investigation is required to develop strategies for improving the NRR of the anammox process for LSWW, since much of the actual wastewater, including the municipal wastewater, tends to be LSWW. In almost all of the studies which achieved a relatively high NRR, the UASB configuration was adopted due to the formation of anammox granules and the accumulation of anammox bacteria in the reactor (Chen et al., 2014; Ma et al., 2018b; Zhang et al., 2018b). For the anammox process conducted in the MBR reactor, it was reported that a high volatile suspended solids (VSS) concentration could be achieved by maintaining a long SRT, and then a high NRR followed. Sun et al. (2018) achieved a VSS of 5.6 g/L with an SRT of 500 d, and an average NRR of 0.98 kgN/m3/day at 25°C. Also, the benefits of low nitrate production were noted due to denitrification, as stated in Section 9.3. Even at a low temperature of 13°C, Wu et al. (2018) achieved an NRR of higher than 0.9 kg-N/m3/day when treating amended domestic sewage. Guo et al. (2018) achieved a high NRR of 1.51 kg-N/ m3/day with a low influent nitrogen concentration of 50 mg NO2
-N/L and 50 mg NH4 +-N/L. These results highlight the advantages of the MBR anammox reactor configuration in the treatment of LSWW at low temperatures. As stated in Section 9.4.1, the NAR range tended to be under 5.2 kg-N/m3/day for most nitritation processes. It is apparent that in the PNA process, an extremely high

Table 9.3 Research reported on the nitritation process.

Reactor MBBR SBR SBR ALR ALR

ALR

Sludge type Biofilm sludge Suspended sludge Suspended sludge Suspended sludge Suspended sludge

Suspended sludge

NH4 + -N-inf

DO

Wastewater

(mg/L)

(mg/L)

Synwastewater Synwastewater Reject water

900

2 mg/L (d) pH
6.8–8.0 (e) Alkalinity requirement
7.14 mgCaCO3 per mg of N. (f) Long HRT (g) Nitrifiers are sensitive to toxic substances, some organic and inorganic substances. (h) Separate sludge systems for BOD and Ammonia removal protect slow-growing nitrifiers and optimize the system. AOBs are Nitrosomonas, Nitrosospira, Nitrosococcus, and Nitrosolobus; and NOBs (that convert nitrites (toxic to plants) to nitrates) are Nitrobacter, Nitrospina, and Nitrococcus.

Denitrification (a) Wastewater, i.e., C10H19O3N (US EPA, 1993)C10H19O3N + 10NO 3 ! 5 N2 + 10CO2 + 3H2O + NH3 + 10OH /or (b) Methanol5CH3OH + 6NO 3 ! 3 N2 + 5CO2 + 7H2O + 6OH (a) D.O. < 0.1 mg/L (b) Sufficient availability of READILY BIODEGRADABLE COD. (c) Denitrification utilizes
2.86 mg CBOD per mg of nitrate denitrified.

Commonly found denitrifiers that can use oxidized nitrogen include Pseudomonas, Bacillus, Zooglea, Alcaligenes, Accumulibacter, and Paracoccus. When they decompose NO3 to metabolize oxygen, nitrogen gas is released.

Benefits of combining both the processes: (a) Denitrification reduces 2.86 mg of carbonaceous BOD demand per mg of nitrate and no addition of air. (b) Reduces 3.57 times of alkalinity requirement per mg of nitrate denitrified.

17.3 Advantages of SND

17.3.4 Nitritation–ANAMMOX The Anaerobic Ammonium Oxidation process is abbreviated as Anammox. Anaerobic ammonium-oxidizing (ANAMMOX) process is used to purify high strength wastewater containing large quantity of ammonia and is a technology having great extent for lesser oxygen requirement. In this method, aerobic ammonia-oxidizing bacteria (AOB) oxidize nearly 50% ammonia to nitrite, used by ANAMMOX bacteria (Candidatus Brocadia fulgida) as an electron acceptor for oxidizing the remaining
50% ammonia to N2 gas. Anaerobic ammonium-oxidizing bacteria (AnAOB) are autotrophic bacteria. Currently, five genera of AnAOB have been accounted for, such as Candidatus Brocadia, Candidatus Kuenenia, Candidatus Scalindua, Candidatus Anammoxoglobus, and Candidatus Jettenia, but the most widespread were Brocadia Anammoxidans and Kuenenia Stuttgartiensis (Strous et al., 1999; Schmid et al., 2000). The process came into the picture at the beginning of the 1990s and has outstanding potential for eliminating ammonia nitrogen in wastewater. The liable bacteria convert ammonium (NH+4 ) and nitrogen dioxide (NO2) into nitrogen gas (N2) and water (H2O). It reduces the cost (lesser energy) for aeration without external carbon sources like methanol or sludge recirculation. During the past 20years, the Anammox process is being widely used in research. Lesser CO2 emissions and sludge production are its strength. In 2007 the first full-scale Anammox process-based bioreactor was constructed in Rotterdam, which showed its excellent potential. But, there are some disadvantages of using that process, such as high construction costs and feasibility. If the Anammox process replaces the conventional nitrification/denitrification in treatment plants, it would be less feasible for large-scale treatment systems. Also, there is a lack of knowledge available (i.e., skilled operation and maintenance required) regarding this process, and the growth rate of anammox is relatively slow, and it takes a long start-up time for sufficient biomass generation for operation ( Jetten et al., 1999; Mozumder and Hossain, 2020).

17.3.5 CANON process Partial nitrification combined with anammox refers to combining anammox reaction with partial nitrification in one reactor titled CANON (completely autotrophic nitrogen removal over nitrite) was one of the potent alternatives for removing ammonium nitrogen. In this method, nitrifiers oxidize 50% of ammonium to nitrite under limited oxygen conditions, which consequently gets ahead through anoxic conditions for anammox reaction. The produced nitrite was utilized with the remaining ammonium by anammox bacteria and converted into nitrogen gas. It was observed that in a CANON process, Nitrosomonas perform as AOB and Planctomycetes as anammox bacteria and perform two sequential reactions simultaneously under oxygen-limited conditions. CANON process takes place in a single reactor in which AOB and anammox coexist in a biofilm or form compact granules (Sliekers et al., 2003; Mozumder and Hossain, 2020; Shah, 2020). Batch experiments and microbial analysis showed that nitrite was at the external biofilm layer, managed by the diffusion of oxygen. The remaining ammonium and nitrite diffuse into the profound part of the

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CHAPTER 17 Role of ammonia oxidizers in SND

N/D over nitrate

N/D over nitrite Advantages: - Reduction of ~25 % aeration cost - ~40% less external carbon source is required - Less CO2 emissions and sludge production Disadvantages: Require external organic carbon source and oxygen

Advantages: -Easy operation -Moderate cost -~96 % Ammonia removal Disadvantages: - Low COD/N ratio is preferable - High external carbon source and oxygen is required -Applicable for low nitrogen containing wastewater

CANON Advantages: - Up to ~62.5 % less oxygen consumption and no requirement of organic carbon - Low process cost and single stage reactor is possible to use -Effective in both low and high COD/N ratio and adverse temperature and toxic shocks cab be mitigated. Disadvantages: Sensitive operational characteristics

Biological Nitrogen Removal processes

Anammox Advantages: -100 % less organic carbon consumption - 50 % less oxygen consumption - Less aeration cost Disadvantages: - Existence of organic matter has detrimental effect -Slow growth rate of anammox organism

SND Advantages: - Low operating time and Operation in a single reactor with identical operating condition - ~95 % total nitrogen removal -Reduction of ~40 % energy consumption, ~300 % less biomass production and faster consumption of Ammonia, Nitrite and Nitrate Disadvantages: - Requires external carbon source and needs optimum and proper control of DO operation

FIG. 17.3 Advantages and disadvantages of various biological nitrogen removal processes (N/D refers to nitrification–denitrification) (Mozumder and Hossain, 2020).

biofilm where anoxic conditions sustain, and nitrite acts as an electron acceptor, which reacts with the remaining ammonium to form nitrogen gas. However, it has been observed that compared to nitrification/denitrification, the N-removal rate of CANON is lower (Sliekers et al., 2003). The merits and demerits of the process have been shown in Fig. 17.3. According to literature and practical results, the SND process has been observed as more economical than other processes because faster consumption of ammonia, nitrite, and nitrate occurs (Walters et al., 2009). Efficient handling reduced near about
40% energy consumption with increasing
63% in denitrification. If no nitrite inhibition was generated,
300% less biomass and
75% oxygen consumption were only needed for nitritation than to complete oxidation to nitrate (Abeling and Seyfried, 1992; Mozumder and Hossain, 2020).

17.4 Types and characteristics of different microbes promoting SND

17.4 Types and characteristics of different ammonia oxidizers and nitrate reducers encouraging SND mechanism prevailing in these systems Nitrification is a considerable process in nitrogen (N) cycling, consisting of the two steps, the oxidation of ammonia to nitrite and consequently nitrite to nitrate (Prosser, 1989). There is a high significance of ammonia oxidation in terms of the global nitrogen cycle and environmental applications as it is the first and rate-controlling step of nitrification (Kowalchuk and Stephen, 2001). For over a hundred years, ammoniaoxidizing bacteria (AOB) were believed to be the only responsible bacteria for ammonia oxidation. However, different ammonia oxidizers can be AmmoniaOxidizing Bacteria, Ammonia-Oxidizing Archaea, Comammox clade A and Comammox clade B. The latter two are responsible for complete ammonia oxidization in salt marshes (Wang et al., 2021). Ammonia-oxidizing bacteria are essential for the fate of nitrogen in the environment. It is known that nitrification is a two-step process of oxidation of ammonia to nitrite and then nitrite to nitrate. It is a fundamental process for the biological removal of nitrogen from municipal and industrial wastewaters. Ammoniaoxidizing bacteria have a vital role in the conversion of ammonia to nitrite in wastewater treatment plants. However, the first step governed by these bacteria is rate limiting as these bacteria are slow growing and quite sensitive to environmental conditions. The characterization of the AOB community in such systems requires the use of genomic methods like Fluorescence in situ hybridization (FISH), realtime polymerase chain reactions (qPCR), PCR followed by denaturing gradient gel electrophoresis (PCR-DGGE). AOB community structure is closely correlated with nitrification performance. Biological ammonia oxidation takes place when the multimeric enzyme ammonia monooxygenase (AMO) transforms ammonia to hydroxylamine. The alpha (A) subunit of the AMO enzyme is predetermined by the amoA gene, modifications generally found in both bacteria and archaea. Using phylogenetic analysis of the 16S rRNA gene and amoA gene, it has been displayed that every AOB found in wastewater is in proper connection with the β-proteobacteria. Associations of the N. oligotropha, N. europaea/Nc. Mobilis, and N. communis lineage usually discovered; however, the expansion of Nitrosospira spp. can be preferential in particular circumstances, like either during typically low pH and low temperature or in industrial WWTPs (Bellucci and Curtis, 2011). Due to the availability of culture-independent molecular biology techniques such as denaturing gradient gel electrophoresis (DGGE), terminal restriction fragment length polymorphism (T-RFLP), sequencing, cloning, and quantitative PCR, new insights into the diversity and circulation of AOB and AOA focusing the amoA genes have developed (Bellucci and Curtis, 2011). Quantitative PCR (qPCR) based microbiological study on the 16SrRNA and amoA gene is frequently applied for the AOB quantification in biofilm and activated sludge (Harms et al., 2003; Wells et al., 2009). Some characteristic species responsible for SND have been presented in

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Table 17.2 (observed using different molecular biological techniques like FISH, qPCR, DGGE, and metagenomic sequencing).

17.5 Operational parameters/factors that control the diversity of nitrifiers (ammonia and nitrite oxidizers) and denitrifiers (nitrate reducers) during the SND mechanism in advanced SBR plants 17.5.1 Carbon source: Readily biodegradable COD, soluble COD, soluble BOD5 Understanding the share of organic contaminants susceptible and resistant to biodegradation is very important in the design of biological removal of nitrogen and phosphorus, as it influences the dynamics of the wastewater treatment process, e.g., the oxygen demand, maintaining a constant sludge age, and kinetic parameters of the biological reactor operation. Płuciennik-Koropczuk et al. (2017) have shown that the Readily Biodegradable COD fraction acts as an essential parameter for influencing SND or denitrification rate in BNR plants. It has been observed in the SND process that readily available substrates present in the wastewater like rbCOD, soluble COD, and soluble BOD (sCOD and sBOD) play a significant role as denitrifiers need a carbon source that can fulfill their requirement of electron donors during denitrification process. So, in advanced SBR plants, anoxic selector compartments before the aeration tanks take up the influent sewage and return activated sludge containing nitrates which perform some denitrification and proliferation of floc formers. It benefits in making large sludge floc sizes up to >500 μm and ultimately promotes SND (Srivastava and Kazmi, 2021).

17.5.2 C/N ratio Denitrifying bacteria need to fight for a carbon source with other heterotrophs, a low C/N ratio in the influent results in a rapid carbon deficit, causing unbalanced simultaneous nitrification and denitrification (SNdN) (Zhao et al., 2008). The rbCOD/TN and BOD5/ TKN ratios also affect denitrification rates in the WWTPs ( Jimenez et al., 2010). Denitrification needs carbon and energy source for bacteria. Complete denitrification is achieved at COD: TKN ratio of 7 (Barnard, 1992), whereas according to Goronszy (1992), generally, a low value of 9 is required to attain biological nutrient removal. Issacs and Henze (1994) proposed that 1.5–2.5 g COD/ g P is used for the phosphate removal, whereas the COD: TN ratio for denitrification is in the range of 3.5–4.5 g COD/g N. This is near to the theoretical obligation for denitrification devoid of COD wastage due to aerobic practices. Acetic acid or methanol is the best carbon source for obtaining higher denitrification rates in comparison to other organic acids

17.5 Operational parameters

(the denitrification rates obtained by them are 1/3rd of the rates produced by acetic acid/methanol) (Tam et al., 1992; Henze, 1989; Gerber et al., 1986).

17.5.3 Floc size and PHB storage The possibility for PHB (poly-β-hydroxybutyrate) to be used as the electron donor for efficient SND was investigated in a SBR plant using a mixed culture and acetate as the organic substrate (Third et al., 2003). Enough carbon prerequisite for denitrification is accomplished through the carbon storage (biosorption—PHB) mechanism and the proportional DO requirement control, which reduces the utilization of substrate carbon by oxic metabolization. The recording by microelectrodes for the occurrence of O2 concentration gradients in individual activated sludge flocs was performed by Satoh et al. (2003). The research concluded that when the O2 concentration in the bulk liquid became less than 45 μM, the anoxic zones have been identified within the sludge flocs with a bigger diameter (approximately 3000 μm). The depth of O2 penetration (diffusion) in the floc was observed to be reliant on the O2 concentration in the bulk liquid. The SND process was observed in the aerated activated sludge when O2 concentration varied from 10 μM to 35 μM (Satoh et al., 2003). Similarly, Daigger et al. (2007) observed that for larger floc particles (2-mm diameter and greater), the dissolved oxygen concentration reached near-zero values at depths depending on process operating conditions. Implications of these results regarding the occurrence of simultaneous nitrification and denitrification include consideration of the factors that affect floc size and distribution in the wastewater treatment plants.

17.5.4 Dissolved oxygen control The other factor is ‘Dissolved oxygen concentration’ control. Augmented DO in the reactor’s bulk liquid negatively affects SND. However, the correlation is not directly obvious, with denitrification and nitrification rates increasing at DO levels up to 0.8 mg/L. SND is a decreasing function of DO concentration at DO > 0.8 mg/L (Pochana and Keller, 1999). SND facilities, examined in the studies, have been demonstrated to remove 80%–96% nitrogen without additional carbon and alkalinity ( Jimenez et al., 2010). Carbon source and bulk dissolved oxygen (DO) concentrations were important process parameters for SND activity. Data suggest that an optimum bulk DO concentration ranging from approximately 0.3 mg per liter (mg/L) to 0.7 mg/L appears to maximize SND activity ( Jimenez et al., 2010). However, this low DO condition needed for SND makes an environment more susceptible. Therefore, in the aeration tanks of SBR plants, DO concentration ranges between 0.0 and 2.5 mg/L with proper control. The

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diffusion rate of nitrate is approximately ten times that of DO within the bigger sludge flocs. A detailed overview of nitrate concentrations and dissolved oxygen flow inside a floc can be seen in the model of Schramm et al. (1999). Under aerobic situations, there is characteristically no nitrate inadequacy in the inner zone of the floc.

17.5.5 ORP The oxidation–reduction potential is an essential parameter regarding nitrification and denitrification. For nitrification, it has been observed that the ORP should lie between +100mV and +330mV in the aerobic zone. For denitrification, it should lie between 50mV and +50mV in the anoxic zone or during the settling phase of SBR aeration tanks (Goronszy, 1992).

17.5.6 pH Typically, the pH for biological nitrogen removal should lie between 6.0 and 8.5. The pH has a significant effect on removing NH+4 -N and TN removal efficiency (Wang et al., 2020). It has been observed that at an average influent pH of 6.5, the removal of NH+4 -N and TN is around 60%–65%. If it gets increased to approximately 7.5, it leads to an increase in NH+4 -N and TN removal efficiency to 80%. When the pH is increased to about 8.0, the removal of NH+4 -N and TN reaches their peak of >90% and >80%, respectively (Wang et al., 2020). Afterward, the nitrification efficiency remained stable up to pH 8.5. However, there is a slight variation in SND efficiency and reaches >90% with varying pH, but instability occurs at sudden changes in pH. It can be concluded that the pH is not the most significant factor influencing SND, but it is an obligatory factor that needed to be considered all through the research (Wang et al., 2020).

17.5.7 Temperature The temperature should be optimal between 85°F and 95°F (30–35°C). However, AOB and NOB are capable of growing between 50°F and 100°F (10–38°C). In temperatures below 41°F (5°C) and above 115°F (45°C), little nitrification can be anticipated. It has been observed that the optimum temperature range is 35–45°C for ammonium oxidizers and 38°C for nitrite oxidizers (Van Hulle et al., 2007; Zhang et al., 2009; Mozumder and Hossain, 2020).

17.5.8 HRT and SRT Long HRT and SRT are needed for the nitrification process in wastewater treatment plants. A representation of nitrogen removal profiles (ammonia and nitrate) concerning ORP and DO in different phases of an SBR cycle is shown in Fig. 17.4.

17.6 Effect of free ammonia, nitrates and metals on AOBs

FIG. 17.4 General pictorial representation of nitrogen removal profiles and DO and ORP during SND process in advanced SBR plants (with selectors).

17.6 Effect of free ammonia (FA), nitrate concentrations, and some metals on AOBs Nitrite accumulation resulted from higher activities of AOB than NOB is unwanted as NOB is more sensitive to free ammonia and nitrous acid. It has been observed that ammonium and nitrite oxidation gets inhibited by free ammonia; where Nitrobacter hinders nitrite oxidation at free ammonia concentration of 0.1–1.0 mg /L (selective inhibition of nitrite oxidation occurs at free ammonia concentration of 1.0–10 mg/L), while Nitrosomonas inhibits ammonium oxidation at free ammonia concentration of 10–150 mg /L. Conclusively, at high pH (>8), free ammonia turns out to be the primary inhibitor for the nitrification process, and at low pH (99%), present in an ordered cluster of amoCamoAamoB in 2–3 copies (except, Nitrosococcus oceani (a γ-proteobacterium) where it is present in only one copy) (Fig. 18.1) which encode three different

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polypeptides, i.e. AmoA, AmoB, and AmoC (Arp and Stein, 2003). The AMO operon consists of amoCAB genes which are present in multiple copies along the genome of bacterial ammonia oxidizers except for N. inopinata, which had been predicted to acquire gene by horizontal gene transfer. In AOB, transcription starts at 166 and 103 bp upstream for amoC start codon and 114 bp upstream for the amoA start codon with an identical σ70 promoter sequences associated with them (Arp et al., 2002; Prosser and Nicol, 2008a). The amoCAB gene cluster codes for three different polypeptides (AmoA, AmoB, and AmoC) which were identified through analogy with particulate Methane Monooxygenase (pMMO). AmoA is a 27 kDa polypeptide that covalently binds 14C2H2, AmoB is encoded by amoB gene with a mass of 44.26 kDa formed by cleavage of a periplasmic signal peptide (Arp et al., 2002). Findings state the presence of 2–3 copies of the identical gene cluster in the amo gene present in the β proteobacterial (beta-AOB) AOB genomes. This cluster is coded by two ORFs (orf4, orf5), except for amo-hao supercluster in Nitrosospira multiformis. However, in γ proteobacterial AOB (gamma-AOB), the terminator for amoB gene sequence is coded by only orf5 gene. This gene (orf5), however, remains conserved in all the AOB at DNA and protein sequence level. This attributes to the ability of AOB cell regulation, through which it rectifies differences in the sequences between multiple copies within an organism (Arp et al., 2002; El Sheikh et al., 2008). Disruptions of individual copies of amoA or amoB (having identical promoter sequence) do not result in differential expression due to the presence of an open reading frame (ORF 4) downstream from both the copies which encode a protein product of 23.53 kDa involved in regulating oxidation of ammonia. In N. europaea, transcripts encoding amoCAB gene cluster along with encoding amoC (with unknown function) and amoAB were observed which possibly are derived from the full-length transcript or are transcribed independently (Arp and Stein, 2003). However, transcripts encoding amoA genes in Nitrosospira sp. were observed by heterologous expression of amoA in Escherichia coli which indicated the presence of at least two separate promoters, controlling amoC and amoAB gene present within the amoCAB operon. The operon is always upregulated by the presence of NH4+ ions, which served dual purposes: firstly, as a signaling molecule and secondly, as an energy source that determines the continuous transcription of amoCAB and amoAB genes even presence of AMO inhibitors (Arp and Stein, 2003). New approaches such as in silico and expression studies analysis had provided more new insights into the biology of ammonia oxidation (El Sheikh et al., 2008). Such studies had revealed the presence of 5 genes in amo gene cluster in the order amoRCABD in Nitrosococcus with overlapping operons (Stein and Klotz, 2011). The two additional genes from the conventional amoCAB gene cluster code for ammonia synthesis (amoR gene present in Nitrosococcus oceani but absent in N. halophilus and N. watsonii) and the other one is found tandemly with a duplicated orthologue downstream of the amoCAB genes in β-AOB (amoD gene). This amoD orthologue encodes particulate methane monooxygenase (pMMO) (Alphaproteobacteria) or encodes copper-blue oxidases (Gammaproteobacteria) despite the presence of amoE gene either downstream of the gene cluster or nearby of a gene tandem

18.5 Economic importance of AOBs

repeats. This gene encoding copper-blue oxidase was conserved in γAOB (upstream of amo gene cluster) along with N. eutropha (Stein and Klotz, 2011). Hydroxylamine Oxidoreductase (HAO) is another important enzyme that catalyzes the oxidation of hydroxylamine to nitrite (Fig. 18.1). This enzyme is unique to autotrophic AOB due to its structure and primary sequence. The gene encoding HAO (hao) is monocistronic and present in multiple copies. The enzyme is encoded by a gene sequence of about 1710 bp in length which codes for 60 kDa polypeptide along with 18–24 amino acid leader sequence, which is removed during the processing and maturation of HAO (Pacheco et al., 2011). The amino acid sequence of HAO is unique among the members of nitrifiers and contained three identical sequences of hao which differed in only one nucleotide gene copy. Two identical nucleotide sequences are present 160 bp upstream of two hao gene copies, hao1 and hao2 whereas a third sequence is present 15 bp upstream of the start codon diverging from the other two corresponds to hao3. Transcription start sites are located 71 bp upstream of the start codon for hao1 and hao2 and 54 bp upstream of the start codon for hao3 with a σ70 promoter sequence associated with all the transcript start sites (Arp et al., 2002). Another important component in the process of ammonia oxidation is the cytochromes that are the direct acceptor during electron transport from HAO (Fig. 18.1). The process is mediated by cytochrome c554, c552 (c-type tetraheme cytochrome), coded by a gene sequence (hcy/cyc) located 1162 bp intervening sequence with no identifiable genes downstream of each hao copy. A σ70 promoter sequence was associated with the transcription start site which is located 97 bp upstream of the start codon with two out of three copies are identical in gene sequence (Hooper et al., 1997). The capacity of AOB to aerobically catabolize ammonia requires both specialized protein complexes, AMO and HAO along with cytochromes c554 and c M552, which relay the electrons to the quinone pool. The three-subunit AMO protein complex starts ammonia catabolism by oxidizing ammonia to hydroxylamine when supplied with reductant from the quinone pool (Badger and Bek, 2008). In addition to amo, hao, and cyc, RubisCO genes also appeared to elicit a role in ammonia oxidation. The presence of RubisCO gene sequence in N. europaea along with other AOB (β-subgroup of the Proteobacteria) imparts a pivotal role in ammonia oxidation. Another copper-containing dissimilatory nitrite reductase (NirK) gene plays a pivotal role in ammonia oxidation. Shreds of evidence had shown that N. europaea cells become more sensitive to NO2 due to the inactivation of nirK gene (Beaumont et al., 2002).

18.5 Economic importance of AOBs Ammonia-oxidizing bacteria play a pivotal role in the proper management of the ecosystem through the natural cycling of nitrogen. With the rise in the nitrogenous wastes from different nitrogen producing industries, animal husbandry, and various anthropogenic activities, the need for the proper management of these emerging

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contaminants is extremely important. Nitrogen-containing pollutants are the important factors causing eutrophication in wastewater (Yin et al., 2018). The wastewater generated in the manufacturing companies consists of a large number of pollutants including metallic compounds, oil, grease, and other suspended solids which need immediate removal. The release of these untreated nitrogenous wastes causes eutrophication of the surrounding polluted areas which poses a threat to the health, including oxygen depletion and toxicity to the ecosystem. Biological treatment of these nitrogenous wastes using bioreactors is an efficient method of removing these wastes from the environment. During biological wastewater treatment, the nitrifying microorganisms undergo nitrification of ammonia and are converted into nitrite and nitrate. Ammonia-oxidizing organisms utilize these wastes as the source of energy. Bacterial genera Nitrosomonas, Nitrosococcus (γ-Proteobacteria), and Nitrosospira (β-Proteobacteria) are reported as the nitrifying bacteria involved in nitrification in wastewater treatment plants (Koops and Pommerening-Roser, 2001). AOB, particularly the Nitrosomonas and Nitrosospira genera are considered as the most potent nitrifying bacteria in the activated sludge (Park et al., 2006). These bacteria possess the enzyme nitrite oxidoreductase (NOR) which catalyzes the formation of nitrate to nitrite. Park et al. (2006) studied the diversity of AOB in activated sludge collected from nine wastewater treatment plants (WWTP) and found that the ammonia monooxygenase (AMO) enzyme complex showed broad specificity to substrates and helps in the oxidation of a wide range of contaminants including aliphatic, aromatic, and recalcitrant halogen (Hooper et al., 1997). They act upon various anthropogenic pollutants such as chloroethane, trichloroethane, dimethyl ether, and estrogens (Hyman et al., 1988). The nitrification enzymes also contribute to the removal of micropollutants such as trace elements and pharmaceuticals from the surface water (Forrez et al., 2011). The application of AOB is also found in biofilter systems which are generally used for the elimination of odors of composting, poultry house exhaust, and waste treatment (Shah et al., 2011). In a biofilter system, AOB interacts with the microbial consortium that utilizes the influent containing ammonia as the substrates to maintain constant biomass and produce contaminant-free effluent (Kowalchuk and Stephen, 2001). In this process, the influent such as ammonia or CO2 is consumed by the microorganisms and thus the nutrients are recycled and a balance is maintained. Moreover, AOB also plays an important role in the removal of hydrocarbon contaminants. For example, N. europaea has been reported to carry out hydroxylation of a range of hydrocarbons such as straight-chain alkane and alkenes and produce primary and secondary alcohol as the by-product (Hyman et al., 1988). Thus ammonia-oxidizing microorganisms, particularly AOBs, have received great interest in the field of wastewater removal plants and monitoring of the environment for sustainable development.

18.6 Conclusion and prospects Nitrification, the microbial oxidation of ammonia to nitrate via nitrite, plays a pivotal role in the biogeochemical nitrogen cycle. Agricultural ecosystems throughout the world receive approximately 25% of the global nitrogen input in the form of

References

synthetic nitrogen fertilizers and thus contribute a common practice that significantly increases crop productivity to feed the growing human population. This fertilizer, mainly (ammonia), has to be oxidized at least once by ammonia-oxidizing prokaryotes to complete the nitrogen cycle. Increases in available nitrogen can alter ecosystems by increasing primary productivity and impacting carbon storage. Because of the importance of nitrogen in all ecosystems and the significant impact on human activities, nitrogen and its transformations have received a great deal of attention from ecologists. Earlier ammonia-oxidizing bacteria were believed to be the only players in the conversion of ammonium to nitrite but accelerated flow of information about ammonia-oxidizing microorganisms due to the application of modern techniques during the recent years, however, has enabled the discovery of AOA and unexpected discovery of complete ammonia oxidizers within the bacterial Nitrospira genus (comammox). This indicated niche specialization and differentiation of comammox organisms with other recognized nitrifying prokaryotes in soil microcosms. Hence, a detailed and systematic understanding of the environmental distribution and activity of these ammonia oxidizers, particularly bacterial contribution and their coordination with recently discovered ammonia-oxidizing communities in terms of ecology, physiology, and biochemistry is required. Functional genomics and allied ‘omics’-based techniques will not only expand our knowledge on the functionality of known enzymes but also help in the identification and understanding of novel enzymes and their metabolic interconnections in AOB and related communities (Stein, 2019). A better understanding of their linkage/contribution to nitrogen losses and greenhouse gas production throughout different environments should be addressed with future research. This will be critical for enhancing our ability to develop better strategies for nitrogen cycle management and to improve the nitrogen use efficiency of crops (Amoo and Babalola, 2017), while at the same time minimizing adverse environmental impacts.

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CHAPTER

Anaerobic ammonia oxidation: From key physiology to full-scale applications

19

Sumira Malika,∗, Shristi Kishorea,∗, Shradha A. Kumara, and Vinay Kumarb Amity Institute of Biotechnology, Amity University Jharkhand, Ranchi, Jharkhand, India bModern College of Arts, Science and Commerce, Pune, India

a

19.1 Introduction In this developing modern world, extensive domestic and industrial activities have led to an increase in the rate of production of wastewater all over the world. If this wastewater is discharged into the water bodies without being treated, it can affect the aquatic life adversely and also cause eutrophication. In order to discharge wastewater into water bodies, it must be decontaminated from the wastes and harmful chemicals. Among all the pollutants present in the wastewater, nitrogen-containing pollutants are one of the most common contaminants present in different types of wastewaters. Since a very long time, various physical, chemical, and biological methods have been employed for the treatment of nitrogen-containing wastewater. For the removal of nitrogen from wastewater, biological treatment is considered efficient as well as eco-friendly than other treatment procedures. The traditional biological treatment system generally involves two steps: nitrification and denitrification. Firstly, the ammonia is biologically oxidized into nitrate (nitrification), and then the oxidation product, i.e., nitrate is reduced subsequently to dinitrogen gas (denitrification). Oxidation of ammonia is the rate-limiting step in the process of nitrification. Thus, it plays a key role in the overall biological nitrogen removal process. Since anaerobic ammonia oxidation (anammox) is carried out by anaerobic microorganisms (anammox bacteria), it does not need an external carbon source and addition of oxygen in the wastewater. This makes anammox treatment methods cost effective and energy efficient over conventional biological wastewater treatments. Also, the ∗

Both the authors contributed equally.

Development in Wastewater Treatment Research and Processes. https://doi.org/10.1016/B978-0-323-91901-2.00008-5 Copyright # 2022 Elsevier Inc. All rights reserved.

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process is environmental friendly as it reduces the emission of greenhouse gases such as CO2 and N2O in the atmosphere. First discovered in 1994 in a denitrifying-fluidized-bed reactor, anammox is a process in which removal of ammonia and nitrite is done simultaneously (Zhang et al., 2008). Due to an absence of knowledge on the pathway involved in the process, anammox received a little attention from researchers. After studying the biological mechanisms and the intermediates involved, it was concluded that anammox is a chemolithotrophic process in which nitrite readily oxidizes one of mole of ammonia to produce dinitrogen gas in anaerobic conditions by the activities of bacteria belonging to Planctomycetes, which are now known as anammox bacteria (Strous et al., 1998). The process can be shown as follows: NH4 + + 1:32NO2
+ 0:066HCO3
+ 0:13H+ ! 1:02N2 + 0:26NO3
+ 0:066CH2 O0:5 N0:15 + 2:03H2 O

The current chapter aims to discuss the key physiology and mechanism of anammox. The chapter further explicitly describes the laboratory and full-scale applications of the anammox process in detail.

19.2 Anammox bacteria: Diversity and cell biology Anammox bacteria are considered as the sole drivers of anaerobic ammonia oxidation. They are known to play a key role in the nitrogen cycle. Being important elements in the anammox process, they are of utmost importance and important area of study. In order to improve the anammox wastewater treatment plants’ efficiencies, the diversity and physiology of anammox bacteria need to be studied well. Anammox bacteria are assigned in the phylum Planctomycetes under domain Bacteria. To date, seven Candidatus genera are described under the phylum Planctomycetes by different literature (Table 19.1). The term Candidatus is generally used to signify those microbes which are yet to be isolated and cultured by traditional microbiological approaches such as culturing on agar plates. These Candidatus bacterial genera include Ca. Brocadia, Ca. Kuenenia, Ca. Jettenia, Ca. Brasilis, Ca. Anammoxoglobus, Ca. Scalindua, and Ca. Anammoximicrobium. Except Ca. Scalindua and Ca. Anammoximicrobium, all the rest five genera are found enormously in the activated wastewater sludge whereas the species of Ca. Scalindua are found more in the natural ecosystems such as marine sediments and oxygen minimum zones and the species of Ca. Anammoximicrobium are isolated from the freshwater (Table 19.1). All these seven bacterial genera belong to a monophyletic order Brocadiales in the phylum Planctomycetes. The anammox bacteria show great diversity and distribution in different locations. For example, in the sediments of Eastern Indian ocean, the most dominant species observed were Ca. Scalindua sp. (Qian et al., 2018) whereas in a mainstream municipal wastewater stream, a dominance of Ca. Brocadia fulgida and Ca. Anammoxoglobus propionicus was seen (Nejidat et al., 2018).

19.2 Anammox bacteria: Diversity and cell biology

Table 19.1 Diversity of anammox bacteria described in the literature. Genus

Species

Sources

References

Candidatus Kuenenia

Ca. Kuenenia stuttgartiensis

Schmid et al. (2000)

Candidatus Brocadia

Ca. Brocadia anammoxidans

Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Freshwater

Ca. Brocadia fulgida

Ca. Brocadia sinica

Ca. Brocadia brasiliensis

Ca. Brocadia caroliniensis Candidatus Anammoxoglobus

Ca. Anammoxoglobus propionicus Ca. Anammoxoglobus sulfate

Candidatus Jettenia

Ca. Jettenia asiatica Ca. Jettenia caeni

Ca. Jettenia moscovienalis Candidatus Brasilis

Ca. Brasilis concordiensis

Candidatus Scalindua

Ca. Scalindua brodae and Ca. Scalindua wagneri Ca. Scalindua sorokonii and Ca. Scalindua arabica Ca. Scalindua sinooifield

Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Wastewater treatment systems Oxygen minimum zones High temperature oil reservoirs

Strous et al. (1999) Kartal et al. (2008) Hu et al. (2010) Araujo et al. (2011) Vanotti et al. (2011) Kartal et al. (2007) Liu et al. (2008) Quan et al. (2008) Ali et al. (2014) Nikolaev et al. (2015) Viancelli et al. (2011) Schmid et al. (2003) Woebken et al. (2008) Li et al. (2010)

Continued

395

396

CHAPTER 19 Anaerobic ammonia oxidation

Table 19.1 Diversity of anammox bacteria described in the literature.—cont’d Genus

Species

Sources

References

Ca. Scalindua zhenghei

Deep sea subsurface sediments Marine sediments Suboxics marine zones Marine sediments

Hong et al. (2011)

Ca. Scalindua marina Ca. Scalindua richardsii Ca. Scalindua profunda

Candidatus Anammoximicrobium

Ca. Anammoximicrobium moscowii

Freshwater

Brandsma et al. (2011) Fuchsman et al. (2012) van de Vossenberg et al. (2012) Khramenkov et al. (2013)

Anammox are coccoid shaped bacteria belonging to phylum Planctomycetes having an average diameter generally lying in the range of 800–1100 nm (van Niftrik et al., 2007). The cell of an anammox bacterium is surrounded by a surface protein layer, also called as an S-layer and has three distinct compartments: periplasm, cytoplasm, and the anammoxosome. Although most of the prokaryotic cells generally do not contain membrane-bound organelles, anammox bacteria contain an intracellular membrane-bound organelle known as anammoxosome. Anammoxosome is the only compartmental structure present inside the cytoplasm of the anammox bacterium. It is the center of the energy metabolism and an absence of DNA and ribosomes is observed here. In most of the cases, it has been noticed that the configuration of the membrane of this intracytoplasmic compartment is in a curved manner (van Niftrik et al., 2008). The presence of this curved anammoxosome membrane provides an advantage of maximized surface area to the anammox bacteria, contributing to enhanced metabolic activities. Anammoxosome membrane contains the enzyme ATPase that clarifies its role in the energy metabolism (Karlsson et al., 2014). Furthermore, some key metabolic enzymes such as hydrazine synthase, hydrazine dehydrogenase, hydroxylamine oxidase, nitrite reductase, nitrite oxidoreductase, etc. have been reported to be localized in the anammoxosome (Kartal et al., 2013). It has been hypothesized that the anammoxosome membrane is associated with the coupling of the anammox reaction and the electron transport chain (ETS) that brings about a proton motive force (PMF). Subsequently the ATP is synthesized by the ATPase bounded to the membrane of the anammoxosome. This hypothesis of energy metabolism in anammoxosome was at first based on the presence of an enzyme similar to hydroxylamine oxidoreductase (HOA) in the anammoxosomes of Ca. K. stuttgartiensis and Ca. B. anammoxidans ( Jetten et al., 2009). Although it has been confirmed that anammoxosome is associated with the anammox reaction and the metabolic proteins required for the process, the

19.3 Physiological parameters and the metabolic pathway in anammox

connection of the anammoxosome membrane with the ETS, PMF, and ATP synthesis still needs to be validated with experimental proofs. Other than having a curved membrane, some iron-containing electron-dense particles and tubular structures have also been found to be present in the anammoxosome (van Niftrik et al., 2008). Although the functions of these iron-containing electron-dense particles are not very clear, they are assumed to help in the iron respiration and act as iron sources for the cell as well the iron-containing or heme proteins. The tubular structures found inside the anammoxosome are assumed to be analogous to cytoskeletal elements in eukaryotes. They might help in the division of the anammoxosome or maintain the curvature of its membrane (van Niftrik and Jetten, 2012). An S-layer has been found to be the outermost layer in the anammox bacterium Ca. K. stuttgartiensis. In laboratory, it has often been found that the bacterium loses its S-layer over time. However, S-layer has been reported to be retained by Ca. K. stuttgartiensis kept over 10 years in the laboratory. This suggests that S-layer plays an important function in Ca. K. stuttgartiensis, probably by protecting the integrity of the cell by acting as an exoskeleton (Peeters and van Niftrik, 2019).

19.3 Physiological parameters and the metabolic pathway involved in anammox To date, the genome of Ca. K. stuttgartiensis, Ca. B. fulgida, Ca. J. Asiatica, Ca. S. profunda, Ca. B. sinica, Ca. B. sp. 40, and Ca. B. caroliniensis has been sequenced (Pereira et al., 2017). The genome sequencing and physiological characterization helps in the niche partitioning of the anaerobic ammonia-oxidizing bacteria. Let us consider the example of environmental conditions required by the different anammox species. NH+4 and NO
2 flushed environments are preferred by Ca. B. sinica due to its low affinity for NH+4 and NO
2 . On the other hand, the affinity constant of Ca. K stuttgartiensis is higher than that of Ca. B. sinica, but still it has a slower rate of growth than Ca. B. sinica. These physiological characteristics suggest that Ca. B. sinica has adopted the r-strategy whereas Ca. K stuttgartiensis has adopted the k-strategy (Oshiki et al., 2016). Table 19.2 summarizes the physiological characteristics of different anammox species. The metabolic pathway for the anammox reaction was first proposed on the basis of 15N studies in a fluidized-based reactor. The dominant species kept in the reactor was Ca. B. anammoxidans. In that study, the intermediate considered for the reduction of nitrite was hydroxylamine (van de Graaf et al., 1997). In another study, a different but related mechanism was put forward by the genomic investigation of Ca. K. stuttgartiensis that considered NO as the intermediate of nitrite reduction rather than hydroxylamine (Strous et al., 2006). In both these studies, hydrazine was considered to be a principal intermediate in the overall process. For the formation of hydrazine, firstly ammonia is oxidized by hydrazine hydrolase, and then HOA or converts it into N2. The metabolic pathway for the anammox reaction is further explained in the next section.

397

Table 19.2 Physiological characterization of different anammox bacteria. Dissolved oxygen tolerance (μM)

Salinity tolerance (mM)

0.0026– 0.0035

0–200

50–200

34  21

0.0041