Transformation Products of Emerging Contaminants in the Environment : Analysis, Processes, Occurrence, Effects and Risks [1 ed.] 9781118339565, 9781118339558

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Editors

Transformation Products of Emerging Contaminants in the Environment Editors Dimitra A. Lambropoulou

Department of Chemistry, Aristotle University of Thessaloniki, Greece

Leo M. L. Nollet

Emeritus, Faculty of Applied Engineering Sciences, University College Ghent, Belgium

Over the last 15 years, the focus of chemical pollution has shifted from conventional pollutants to so-called “emerging” or “new” unregulated contaminants. These include pharmaceuticals and personal care products, hormones, UV filters, perfluorinated compounds, poylybrominated flame retardants (BFRs), pesticides, plasticizers, artificial sweeteners, illicit drugs, and endocrine disruptor compounds (EDCs). Despite the increasing number of published studies covering emerging contaminants, we know almost nothing about the effects of their transformation products and/or metabolites. This two-volume set provides a unique collection of research on transformation products, their occurrence, fate and risks in the environment. It contains 32 chapters, organised into 7 parts, each with a distinct focus:

• General Considerations • Transformation Processes and Treatment Strategies • Analytical Strategies • Occurrence, Fate and Effects in the Environment • Global Speciality and Environmental Status • Risk Assessment, Management and Regulatory Framework • Outlook Transformation Products of Emerging Contaminants in the Environment is a valuable resource for researchers and industry professionals in environmental chemistry, analytical chemistry, ecotoxicology, environmental sciences, and hydrology, as well as environmental consultants and regulatory bodies.

Lambropoulou

T WO VO LU M E S E T

Nollet

Transformation Products of Emerging Contaminants in the Environment

Analysis, Processes, Occurrence, Effects and Risks

T WO VO LU M E S E T

T WO VO LU M E SET

Transformation Products 201 x 261mm of Emerging Contaminants (for 246 x 189 PPC cover) in the Environment Analysis, Processes, Occurrence, Effects and Risks

Editors Dimitra A. Lambropoulou Leo M. L. Nollet

Transformation Products of Emerging Contaminants in the Environment

Transformation Products of Emerging Contaminants in the Environment Analysis, Processes, Occurrence, Effects and Risks

EDITED BY DIMITRA A. LAMBROPOULOU AND LEO M. L. NOLLET

This edition first published 2014 # 2014 John Wiley and Sons Ltd Registered office John Wiley & Sons Ltd, The Atrium, Southern Gate, Chichester, West Sussex, PO19 8SQ, United Kingdom For details of our global editorial offices, for customer services and for information about how to apply for permission to reuse the copyright material in this book please see our website at www.wiley.com. The right of the author to be identified as the author of this work has been asserted in accordance with the Copyright, Designs and Patents Act 1988. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, except as permitted by the UK Copyright, Designs and Patents Act 1988, without the prior permission of the publisher. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print may not be available in electronic books. Designations used by companies to distinguish their products are often claimed as trademarks. All brand names and product names used in this book are trade names, service marks, trademarks or registered trademarks of their respective owners. The publisher is not associated with any product or vendor mentioned in this book. This publication is designed to provide accurate and authoritative information in regard to the subject matter covered. It is sold on the understanding that the publisher is not engaged in rendering professional services. If professional advice or other expert assistance is required, the services of a competent professional should be sought. The publisher and the author make no representations or warranties with respect to the accuracy or completeness of the contents of this work and specifically disclaim all warranties, including without limitation any implied warranties of fitness for a particular purpose. This work is sold with the understanding that the publisher is not engaged in rendering professional services. The advice and strategies contained herein may not be suitable for every situation. In view of ongoing research, equipment modifications, changes in governmental regulations, and the constant flow of information relating to the use of experimental reagents, equipment, and devices, the reader is urged to review and evaluate the information provided in the package insert or instructions for each chemical, piece of equipment, reagent, or device for, among other things, any changes in the instructions or indication of usage and for added warnings and precautions. The fact that an organization or Website is referred to in this work as a citation and/or a potential source of further information does not mean that the author or the publisher endorses the information the organization or Website may provide or recommendations it may make. Further, readers should be aware that Internet Websites listed in this work may have changed or disappeared between when this work was written and when it is read. No warranty may be created or extended by any promotional statements for this work. Neither the publisher nor the author shall be liable for any damages arising herefrom.

Library of Congress Cataloging-in-Publication Data Transformation products of emerging contaminants in the environment : analysis, processes, occurrence, effects and risks / edited by Leo M.L. Nollet and Dimitra A. Lambropoulou. pages cm Includes index. ISBN 978-1-118-33959-6 (cloth) 1. Chemicals–Environmental aspects. 2. Speciation (Chemistry) 3. Pollutants–Biodegradation. 4. Environmental chemistry. I. Nollet, Leo M. L., 1948- editor. II. Lambropoulou, Dimitra A., editor. TD196.C45T728 2014 363.7380 4–dc23 2013022173 A catalogue record for this book is available from the British Library. ISBN:9781118339596 (13 digits) Set in 10/12 pt Times by Thomson Digital, Noida, India

I would also like to offer my heartfelt thanks to my co-editor Leo, who generously shared his distinguished expertise with me,while serving as peer respondent during this editorship. I’m really touched by his kind words and thoughts and I feel very grateful for this wonderful collaboration. I am very happy to look forward to working with him again. Dimitra A. Lambropoulou For a fine collaboration during the redaction period of this book I like to thank my co-editor Dimitra. I hope she will continue to be a well appreciated scientist. I hope we will able cooperate in a number of future projects. Leo M.L. Nollet

Contents

Preface

xxi

List of Contributors

xxiii VOLUME 1

PART I

GENERAL CONSIDERATIONS

1

1 Classifying the Transformation Products (TPs) of Emerging Contaminants (ECs) for Prioritizing Research into their Impact on the Environment and Human Health 3 Jacek Namie
snik, Lidia Wolska, Radosław Czernych, Gra_zyna Gałe˛zowska and Monika Cieszy
nska 1.1 Introduction 1.2 Emerging Contaminants – Emerging Problem 1.2.1 Veterinary and Human Antibiotics 1.2.2 Human Drugs 1.2.3 Industrial and Household Wastewater Products 1.2.4 Sex and Steroidal Hormones 1.3 Transformation Products of ECs 1.3.1 Veterinary and Human Antibiotics 1.3.2 Human Drugs 1.3.3 Industrial and Household Wastewater Products 1.3.4 Sex and Steroidal Hormones 1.4 Minimizing Environmental Risk of ECs and their TPs 1.4.1 Designing a Risk Minimization Strategy 1.4.2 Results of the Prioritization Procedure 1.5 Concluding Remarks and Future Perspectives References

3 5 6 6 8 40 41 41 41 42 42 43 43 45 45 49

2 Transformation Products of Emerging Organic Compounds as Future Groundwater and Drinking Water Contaminants Marianne E. Stuart and Dan J. Lapworth

65

2.1 2.2 2.3 2.4

Introduction Sources and Pathways of Emerging Contaminants to Groundwater Persistence in the Groundwater Environment Emerging Contaminants and their Transformation Products in Groundwater 2.4.1 Pesticides 2.4.2 Pharmaceuticals 2.4.3 Personal Care Products and Synthetic Musks

65 66 68 69 69 71 73

viii Contents

2.4.4 Caffeine and Nicotine 2.4.5 Alkylphenols and Other Endocrine Disruptors 2.4.6 Disinfection By-Products 2.4.7 Brominated and Fluorinated Compounds 2.4.8 Triazoles 2.4.9 Naphthenic Acids 2.4.10 Explosive Residues 2.4.11 Algal Toxins 2.5 Toxicity and Risk Assessment 2.6 Conclusions References PART II

TRANSFORMATION PROCESSES AND TREATMENT STRATEGIES

3 Phototransformation Processes of Emerging Contaminants in Surface Water Davide Vione and Serge Chiron 3.1 Introduction 3.2 Direct Photolysis and Sensitised Reactions in the Transformation of Emerging Contaminants 3.2.1 Direct Photolysis 3.2.2 Reaction with OH 3.2.3 Reaction with CO3  3.2.4 Reaction with 3 CDOM  3.2.5 Reaction with 1 O2 3.3 The Case of Photonitration 3.4 Towards the Modelling of Phototransformation Kinetics in Surface Water 3.4.1 Surface-Water Absorption Spectrum 3.4.2 Reaction with OH 3.4.3 Direct Photolysis 3.4.4 Reaction with CO3  3.4.5 Reaction with 1 O2 3.4.6 Reaction with 3 CDOM  3.4.7 Photochemical Transformation of Organic Pollutants 3.4.8 Photo-Transformation of Intermediates References 4 Transformation Products of Emerging Contaminants upon Reaction with Conventional Water Disinfection Oxidants Jos
e Benito Quintana, Rosario Rodil and Isaac Rodr
ıguez 4.1 Introduction 4.2 Analytical Methodology for Transformation Products Identification 4.2.1 GC-MS-Based Approaches 4.2.2 LC-MS-Based Approaches 4.3 Factors Influencing the Kinetics of Chlorination 4.4 Overview of Typical Reaction Mechanisms During Free Chlorine Treatments

73 73 74 75 75 76 76 76 76 78 79 87

89 89 90 90 96 100 102 103 104 106 108 108 111 112 113 114 114 117 118 123 123 124 125 128 131 135

Contents ix

4.5 Review of Current Knowledge of Emerging Pollutant Reactions with Free Chlorine 4.5.1 Pharmaceuticals 4.5.2 Androgenic and Estrogenic Steroidal Compounds 4.5.3 Substances of Abuse 4.5.4 Bisphenol A and Nonylphenol 4.5.5 Bactericides: Triclosan and Parabens 4.5.6 UV Filters 4.5.7 Antioxidants 4.5.8 Cyanotoxins 4.6 Other Disinfection Agents 4.6.1 Chlorine Dioxide 4.6.2 Chloramination 4.6.3 Permanganate and Ferrate 4.7 Conclusions and Outlook References 5 Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants: Ongoing Research and Recommendations for Future Work Sixto Malato, Pilar Fern
andez-Ib
a~ nez, Isabel Oller, Lucia Prieto-Rodriguez, Sara Miralles-Cuevas and Alejandro Cabrera-Reina 5.1 5.2 5.3 5.4 5.5

Introduction Ozonation Membrane Processes Membrane Bioreactors (MBR) AOPs Including Solar AOPs 5.5.1 Solar Driven Advanced Oxidation Processes 5.5.2 Different Approaches for Treating ECs by Solar AOPs References 6 Transformation Products of Emerging Contaminants Formed during Advanced Oxidation Processes Ioannis K. Konstantinou, Maria Antonopoulou and Dimitra A. Lambropoulou 6.1 Introduction 6.2 Pesticides 6.2.1 Organophosphorus Insecticides 6.2.2 Anilide Herbicides (ANHs) 6.2.3 Phenylurea Herbicides (PUHs) 6.2.4 Neonicotinoid Insecticides (NCIs) 6.2.5 Glyphosate Herbicide 6.3 Phthalate Esters 6.4 Pharmaceutical Compounds 6.4.1 Fibrates 6.4.2 b-Blockers 6.5 Others 6.5.1 Bisphenol A 6.5.2 Triclosan

138 141 146 147 147 147 148 148 149 150 150 152 153 155 155

161

161 163 165 167 169 170 172 175 179 179 180 180 183 187 198 203 203 204 205 211 215 215 216

x Contents

6.6 Conclusions Acknowledgments References 7 Enzymatic Reactors Applied for the Biotransformation of Endocrine Disrupting Chemicals Juan M. Lema Rodicio, Ma Teresa Moreira, Gemma Eibes, Thelmo A. Lu-Chau, Luc
ıa Lloret, Roberto Taboada, Adriana Arca-Ramos and Gumersindo Feijoo 7.1 Endocrine Disrupting Chemicals 7.1.1 Sources and Fate 7.1.2 Physicochemical Properties and Ecotoxicity of Endocrine Disrupting Chemicals 7.1.3 Estrogenic Activity 7.1.4 Methods for the Removal of EDCs 7.2 White-Rot Fungi and Their Lignin Modifying Enzymes 7.2.1 Characteristics of the White-Rot Fungi 7.2.2 Lignin-Modifying Enzymes and Their Mediators 7.3 Enzymatic Reactors 7.3.1 Free Enzymes 7.3.2 Immobilized Enzymes 7.4 Determination of Transformation Products from the Enzymatic Treatment of EDCs 7.4.1 Analytical Techniques 7.4.2 Transformation Products Detected in Enzymatic Treatments References 8 Biologically Mediated Chiral Inversion of Emerging Contaminants Stuart J. Khan 8.1 Introduction 8.1.1 Nomenclature 8.1.2 Enantiomeric Fraction 8.1.3 Chiral Emerging Contaminants 8.2 Enantiospecific Analytical Methods 8.3 Changes in Enantiomeric Composition During Biological Transformation Processes 8.4 Evidence for Biologically Mediated Chiral Inversion 8.5 Implications and Priorities for Future Research References PART III

ANALYTICAL STRATEGIES

9 Quality Issues in Water Sampling, Sample Pre-Treatment and Monitoring Sara Bogialli, Stefano Polesello and Sara Valsecchi 9.1 9.2 9.3 9.4

Introduction Monitoring of Transformation Products in Water Bodies Sample Representativeness and Stability Issues Identification of Transformation Products and Legislative Requirements

217 218 218

229

229 230 231 231 234 234 234 234 238 238 243 248 248 249 255 261 261 261 262 262 264 268 271 274 274 281 283 283 284 287 292

Contents xi

9.4.1 Basic Principles 9.4.2 Quality Control in Qualitative Analysis of TPs 9.4.3 Applications 9.5 Conclusions References 10 Transformation Products of Emerging Contaminants: Analytical Challenges and Future Needs Bozo Zonja, Jaume Ace~ na, Aleksandra Jelic, Mira Petrovic, Sandra P
erez Solsona and Dami a Barcel
o 10.1 Introduction 10.2 Generation, Detection and Identification of Transformation Products at Lab Scale: An Analytical Challenge 10.2.1 Detection of Transformation Products with GC-MS and LC-MS 10.2.2 Identification of Transformation Products 10.3 Quantitative Analysis of TPs in the Environment 10.3.1 Sample Preparation 10.3.2 Determination of TPs in the Environment with High and Low Resolution Mass Spectrometric Techniques 10.4 Evaluation of the Toxicity of TPs 10.5 Conclusions and Future Needs Acknowledgments References 11 Advanced Mass Spectrometry-Based Techniques for the Identification and Structure Elucidation of Transformation Products of Emerging Contaminants Paola Calza and Debora Fabbri 11.1 Introduction 11.2 Potential and Differences Among the Different MS Systems for Determining Unknown Compounds 11.2.1 GC-MS versus LC-MS 11.2.2 Capability and Potential of Instrumentation: The Right Analyzer 11.3 How to Proceed in the Structural Attribution 11.3.1 Kind of Analysis to be Performed 11.3.2 Strategies of Accurate Mass Screening for (non)Target Compounds and Unknowns: Elementary Composition Assignment 11.3.3 Structural Attribution by MSn 11.3.4 Limitation 11.4 Accurate Mass Screening and Identification of Emerging Contaminants in Environmental Samples: Some Cases Studied 11.4.1 The Spectra Library: GC-MS Approach 11.4.2 LC-MS2 (QqQ) Identification of Fluoxetine Transformation Products 11.4.3 Identification of Unknown with LTQ Orbitrap 11.4.4 Focus on the Same Pollutant Studied with Different Analyzers: The Case of Diclofenac 11.4.5 The Use of FTICR-MS for Photodecomposition Transformation Products of Two Pesticides

292 294 296 297 298 303

303 305 306 309 311 312 313 318 319 320 320

325 325 326 326 327 330 330 333 333 334 334 334 335 336 340 340

xii Contents

11.4.6 Combined Approach 11.5 Conclusions References

344 345 346

12 Applications of NMR Techniques for the Identification and Structure Elucidation of Emerging Organic and Other Xenobiotic Organic Contaminants Alfred Preiss and Markus Godejohann 12.1 Introduction 12.2 Basic Techniques 12.2.1 Sample Enrichment and Clean-up 12.2.2 Tube NMR 12.2.3 Hyphenated NMR Techniques 12.3 Applications 12.3.1 EOCs and XOCs: Identification of the Transformation Products 12.3.2 Identification of Disinfection By-Products 12.3.3 Direct Analysis of XOCs and EOCs in Environmental Samples 12.3.4 Other Topics 12.4 Conclusions List of Abbreviations References

351 351 353 353 354 356 359 359 367 373 376 377 377 378

VOLUME 2 PART IV

OCCURRENCE, FATE AND EFFECTS IN THE ENVIRONMENT: AN OVERVIEW OF MAJOR CLASSES

13 Transformation Products of Pesticides in the Environment: Analysis and Occurrence Ana Ag€ uera L
opez, Mar
ıa del Mar G
omez and Amadeo R. Fern
andez-Alba 13.1 Introduction 13.2 Transformation of Pesticides in the Environment 13.3 Analytical Techniques Used in the Identification and Analysis of TPs 13.3.1 Sample Preparation and Preconcentration 13.3.2 Analytical Determination 13.4 Occurrence of Pesticide TPs in the Environment 13.5 Concluding Remarks Acknowledgments References 14 Metabolites and Transformation Products of Pharmaceuticals in the Aquatic Environment as Contaminants of Emerging Concern Irene Michael, Marlen Ines Vasquez, Evroula Hapeshi, Tarek Haddad, Ewelina Baginska, Klaus K€ ummerer and Despo Fatta-Kassinos 14.1 Introduction

385

387 387 388 397 397 403 407 408 408 409

413

413

Contents xiii

14.2 Human Metabolites in the Aquatic Environment 14.3 Biotransformation Products in the Aquatic Environment 14.4 Transformation of Pharmaceuticals During Photolysis and Advanced Oxidation Processes 14.4.1 Photolysis 14.4.2 Advanced Oxidation Processes (AOPs) 14.5 Conclusions and Outlook Acknowledgments References

415 418 425 425 439 446 447 447

15 Transformation Products of Personal Care Products: UV Filters Case Studies Kristina Pestotnik, Tina Kosjek and Ester Heath 15.1 Introduction 15.2 Main Physico-Chemical Parameters of UV Filters and their Influence on Environmental Behaviour 15.3 Occurrence of UV Filter Residues 15.3.1 UV Filters 15.3.2 UV Filter Transformation Products 15.4 Fate of UV Filter Residues 15.4.1 Abiotic Processes 15.4.2 Biotic Processes 15.5 Analytical Methods for Identification of Transformation Products 15.6 Effects and Toxicity of UV Filters and their Transformation Products in the Environment 15.6.1 Ecotoxicity 15.6.2 Estrogenic Activity 15.6.3 Toxicity of Transformation Products 15.7 Conclusions and Future Strategies Acknowledgments Abbreviations References

459

16 Transformation Products of Illicit Drugs Dimitra A. Lambropoulou and Eleni Evgenidou

493

16.1 Introduction 16.1.1 What are “Illicit Drugs?” 16.2 Fate and Treatment of IDs and Their Metabolites/TPs 16.2.1 Environmental Fate of IDs and Their Metabolites/TPs 16.2.2 Treatment Studies of IDs and Their Metabolites/TPs 16.3 Analytical Methods and Detection 16.3.1 Sampling and Storage 16.3.2 Analytical Methods and Detection 16.4 Occurrence of IDs and their Metabolites/TPs in the Environment 16.4.1 Wastewaters 16.4.2 Surface Waters 16.4.3 Groundwater and Drinking Water 16.4.4 Air

459 461 465 465 467 467 467 475 478 483 483 483 485 486 487 487 489

493 494 495 495 497 503 503 504 507 507 512 512 513

xiv Contents

16.5 Ecotoxicity of IDs and Their Metabolites/TPs 16.5.1 Sewage Epidemiology 16.6 Concluding Remarks References 17 Transformation Products of Artificial Sweeteners Marco Scheurer, Heinz-J€ urgen Brauch and Frank Thomas Lange 17.1 Introduction 17.2 Processes Leading to the Formation of Artificial Sweetener Transformation Products 17.2.1 Metabolism in Mammals Including Man 17.2.2 Abiotic Transformation 17.2.3 Biodegradation 17.2.4 Saccharin and its Transformation Products from Agricultural Sources 17.3 Summary and Conclusions References 18 Transformation Products of Brominated Flame Retardants (BFRs) Alin C. Dirtu, Alin C. Ionas, Govindan Malarvannan and Adrian Covaci 18.1 Introduction 18.2 Transformation Products of PBDEs 18.2.1 Degradation of PBDEs in Abiotic Matrices 18.2.2 Biotransformation Pathways for PBDEs 18.3 Transformation Products of HBCDs 18.3.1 Degradation of HBCDs in Abiotic Matrices 18.3.2 Biotransformation Pathways for HBCDs 18.4 Transformation Products of TBBPA 18.4.1 Degradation of TBBPA in Abiotic Matrices 18.4.2 Biotransformation Pathways for TBBPA 18.5 Transformation Products of NBFRs 18.5.1 Decabromodiphenyl Ethane (DBDPE) 18.5.2 1,2-Bis(2,4,6-tribromophenoxy)ethane (BTBPE) 18.5.3 2-Ethylhexyl2,3,4,5-tetrabromobenzoate (TBB) and Di(2-ethylhexyl)tetrabromophthalate (TBPH) 18.6 Concluding Remarks and Future Perspectives Acknowledgments References

513 514 515 515 525 525 527 527 531 536 537 539 540 545 545 546 546 553 557 557 559 561 561 563 566 566 566 567 568 568 569

19 Transformation Products of Alkylphenols 577 Montserrat Cortina-Puig, Gabino Bol
ıvar-Subirats, Carlos Barata and Silvia Lacorte 19.1 Alkylphenols: Types, Properties and Uses 19.2 Transformation of Alkylphenols and Identification of Transformation Products 19.2.1 Biodegradation 19.2.2 Photodegradation 19.2.3 Sonolysis 19.3 Occurrence of Alkylphenol Transformation Products in the Environment

577 580 580 587 590 591

Contents xv

19.3.1 Water 19.3.2 Air 19.3.3 Soil/Sediments 19.3.4 Biota 19.4 Risks and Effects of Alkylphenols and their Transformation Products in the Environment 19.4.1 Toxicological Effects 19.4.2 Risk Assessment 19.5 Conclusions Acknowledgments References 20 Biotic and Abiotic Transformation Processes of Benzotriazoles: Possible Pathways and Products Dimitra Voutsa 20.1 Introduction 20.2 Biotic Degradation Processes 20.3 Abiotic Transformation Processes 20.3.1 Photochemical Transformation 20.3.2 Chemical Oxidation Processes 20.4 Future Research Needs References 21 Identification (Quantitative Determination and Detection) and Fate of Transformation Products of Rocket Fuel 1,1-Dimethylhydrazine Bulat Kenessov and Lars Carlsen 21.1 Introduction/Background 21.2 Identification of Transformation Products of 1,1-Dimethylhydrazine 21.2.1 Laboratory Experiments 21.2.2 Examination of Fall Sites 21.2.3 Possible Mechanisms of Formation of the Main Transformation Products 21.3 Distribution and Fate of Transformation Products of 1,1-Dimethylhydrazine in Soil at Fall Sites 21.4 Analytical Methods Applied in the Monitoring 21.4.1 Methods of Quantitative Determination 21.4.2 Systematic Approach to the Control of TPs of 1,1-Dimethylhydrazine 21.5 Conclusion References 22 Assessment of the Occurrence and Fate of Transformation Products of Endocrine Disrupting Compounds EDCs in the Environment Vasiliki Boti, Vasilios Sakkas and Triantafyllos Albanis 22.1 Introduction 22.2 Endocrine Disrupting Compounds (EDCs) of Concern 22.2.1 Definitions and Regulatory Issues 22.2.2 Mechanisms of Endocrine Disruption

591 595 596 597 598 598 600 603 604 604 613 613 615 619 619 620 622 622 627 627 628 628 630 631 634 635 636 641 644 645 649 649 650 650 652

xvi Contents

22.3 Environmental Fate and Transformation of EDCs 22.4 Analytical Methodology 22.5 Occurrence and Endocrine Disruption Effects of the TPs of Selected EDCs 22.5.1 Pesticides 22.5.2 Industrial Chemicals 22.5.3 Synthetic and Natural Steroids 22.6 Future Needs –Recommendations References 23 Transformation Products of Hazardous Cyanobacterial Metabolites in Water Anastasia Hiskia, Theodoros M. Triantis, Maria G. Antoniou, Armah A. de la Cruz, Kevin O’Shea, Weihua Song, Theodora Fotiou, Triantafyllos Kaloudis, Xuexiang He, Joel Andersen and Dionysios D. Dionysiou 23.1 Introduction 23.2 Cyanobacterial Secondary Metabolites 23.2.1 Hazardous Cyanobacterial Metabolites: Cyanotoxins 23.2.2 Taste and Odor Compounds: Geosmin and 2-Methylisoborneol 23.3 Transformation Products of Cyanobacterial Metabolites in Water 23.3.1 Fate of Cyanobacterial Metabolites in the Environment 23.3.2 Chemical Oxidation 23.3.3 Advanced Oxidation Processes 23.4 Research Gaps, Recent Trends and Future Needs References PART V

GLOBAL SPACIALITY AND ENVIRONMENTAL STATUS OF TRANSFORMATION PRODUCTS IN THE ENVIRONMENT

24 Occurrence of Transformation Products of Emerging Contaminants in Water Resources Carlos GonSc alves, Maria A.D. Sousa and Maria de F
atima Alpendurada 24.1 Brief Introduction on the Sources of Transformation Products of Emerging Contaminants 24.2 Transformation Products in Natural Waters: From Contamination Sources to Drinking Water Production 24.3 Wastewaters as a Major Source of Transformation Products 24.4 Origin and Presence of Transformation Products in Drinking Water 24.5 Ubiquity and Regio-Specificity of Transformation Products 24.6 Transformation Products of Emerging Contaminants: Fate and Behavior 24.7 Conclusions References

652 654 660 660 664 667 668 669 675

675 676 676 681 682 682 687 692 698 699 709 711

711 713 732 738 740 741 744 746

25 Occurrence of Transformation Products of Emerging Contaminants in Water Resources of the United States 751 Imma Ferrer and E. Michael Thurman 25.1 Introduction: Emerging Contaminants 25.2 State-of-the-Art Techniques for the Identification of Emerging Contaminants and Their Transformation Products

751 752

Contents xvii

25.2.1 Liquid Chromatography/Tandem mass Spectrometry (LC/MS-MS) for the Analysis of Target Compounds. EPA Method 1694 25.2.2 Liquid Chromatography/Time-of-Flight/Mass Spectrometry (LC/TOF-MS) for the Analysis of Non-target Compounds 25.2.3 Liquid Chromatography/Quadrupole-Time-of-Flight/Mass Spectrometry (LC/Q-TOF-MS) for Structural Elucidation of Unknown Compounds and Transformation Products 25.3 Use of Accurate Mass Tools for the Identification of Emerging Contaminants 25.3.1 Molecular Features 25.3.2 Accurate Mass Databases 25.3.3 Accurate Mass Filters and Isotopic Mass Defect 25.3.4 Accurate Mass Profiling 25.4 Occurrence of Transformation Products in Environmental Waters in the US References 26 Spatial Modeling for Elucidation of Perfluorinated Compound Sources and Fate in a Watershed Yasuyuki Zushi and Shigeki Masunaga 26.1 Introduction 26.1.1 Transformation Products of PFCs 26.1.2 History, Regulation and Pollution of PFCs 26.2 Source Identification of PFCs Using GIS 26.2.1 Study Area and Dataset 26.2.2 Method of GIS-based Source Identification 26.2.3 Results and Discussion 26.3 Spatial Distribution of PFOS and PFOA Contributed by Nonpoint Sources 26.3.1 Method for Spatial Prediction with Fine Scale; Use of the Digital Elevation Model (DEM) and Land-Use Regression (LUR) Model 26.3.2 Results and Discussion 26.4 Conclusion Acknowledgments References 27 Global Distribution of Polyfluoroalkyl and Perfluoroalkyl Substances and their Transformation Products in Environmental Solids Holly Lee and Scott A. Mabury 27.1 Introduction 27.2 Global Contamination of PFASs in Environmental Solid Matrices 27.2.1 Sediments 27.2.2 Temporal Trends in Sediment Cores 27.2.3 Wastewater Treatment Plant Sludge 27.2.4 Soils 27.2.5 Case Study: Contamination of Agricultural Farmlands in Decatur, Alabama 27.3 Fate of PFASs in Environmental Solids

753 755

755 756 756 758 762 765 767 770 775 775 775 779 780 780 782 783 786 786 787 792 793 793 797 797 801 801 803 804 806 807 809

xviii Contents

27.3.1 Sorption 27.3.2 Leaching to Surface Waters and Groundwater 27.3.3 Biodegradation in WWTP Media and Soils 27.4 Uptake into Vegetation 27.5 Summary and Future Outlook References PART VI

RISK ASSESSMENT, MANAGEMENT AND REGULATORY FRAMEWORK

28 Toxicity and Risk of Transformation Products of Emerging Contaminants for Aquatic Organisms: Pharmaceutical Case Studies Marina DellaGreca, Marina Isidori and Fabio Temussi 28.1 Introduction 28.2 Photolysis in the Environment: Pharmaceutical Case Studies 28.3 Effect-Driven Approach 28.3.1 Amiloride 28.3.2 Chlorpromazine 28.3.3 Diclofenac 28.3.4 Dipyrone 28.3.5 Propranolol 28.3.6 Ranitidine and Tramadol 28.3.7 Spiramycin 28.3.8 Discussion 28.4 Exposure-Driven Approach 28.4.1 Amlodipine 28.4.2 Estrone 28.4.3 Furosemide 28.4.4 Naproxen 28.4.5 Prednisone, Prednisolone, and Dexamethasone 28.4.6 Ranitidine 28.4.7 Tamoxifen 28.4.8 Discussion 28.5 Conclusion References 29 Quantitative Structure–Activity Relationship/Quantitative Structure–Toxicity Relationship (QSAR/QSTR) Modeling as Tools for Assessing Effects and Predicting Risks of Transformation Products of Emerging Contaminants Lars Carlsen and Bulat Kenessov 29.1 Introduction 29.2 The Toolbox 29.2.1 EPI Suite 29.2.2 PASSOnline 29.2.3 ADME/Tox Boxes 29.2.4 Partial Order Ranking

810 812 812 815 817 818 827

829 829 830 830 832 834 835 836 837 838 839 840 841 841 848 849 850 851 851 853 854 855 856

859 859 861 862 863 863 863

Contents xix

29.3 Environmental Behavior 29.3.1 EPI Suite Results 29.4 Ecotoxicological Effect 29.4.1 ECOSAR Results 29.5 Effects on Humans 29.5.1 Predictions of Selected Biological Activities 29.5.2 Organ-Specific Adverse Health Effects 29.6 Conclusions References 30 Steps Toward a Regulatory Framework for Transformation Products in Water Maria D. Hernando Guil, Maria J. Mart
ınez-Bueno, Laura Duran, Jos
e M. Navas and Amadeo R. Fern
andez-Alba 30.1 Introduction 30.2 Scientific Advances and Technical Knowledge of Transformation Products. Relevant Cases of Study 30.2.1 Polar Pesticides 30.2.2 Biocides 30.2.3 Pharmaceuticals 30.2.4 Industrial Chemicals 30.3 Toxicological Considerations in Assessing Mixtures of Chemicals and Significance of Transformation Products in EU Regulations. Interaction Between Regulatory Frameworks References 31 NORMAN Association: A Network Approach to Scientific Collaboration on Emerging Contaminants and their Transformation Products in Europe Jaroslav Slobodnik and Valeria Dulio 31.1 Introduction 31.1.1 Major Challenges 31.2 The NORMAN Network as a Science-to-Policy Interface 31.2.1 Prioritisation of Emerging Contaminants and their Transformation Products 31.2.2 Transformation Products and Mixture Toxicity 31.2.3 Transformation Products 31.2.4 Mixtures 31.2.5 Toxicity Profiling 31.3 Effect-Directed Analysis for Identification of Relevant Emerging Contaminants and their Transformation Products in Complex Environmental Samples 31.3.1 Publications 31.3.2 Databases 31.3.3 International Projects 31.4 Quality Control Aspects 31.4.1 Method Validation 31.4.2 Interlaboratory Studies

864 864 867 867 868 868 870 872 873 877

877 879 879 881 889 891

894 896 903 903 904 905 905 907 907 908 909

909 911 911 912 912 912 912

xx Contents

31.5 Conclusions Acknowledgments References PART VII

OUTLOOK

32 Outlook Dimitra A. Lambropoulou and Leo M. L. Nollet 32.1 General Remarks 32.2 Gaps, Recommendations and Future Needs 32.2.1 Elucidation, Detection, Quantification, and Environmental Occurrence 32.2.2 Environmental Fate and Transformation 32.2.3 Health Effects, Risk Assessment, and Prioritization 32.2.4 Remediation Approaches Index

913 914 914 917 919 919 921 921 922 922 923 925

Preface

Over the last 15 years, the focus of chemical pollution has definitely shifted from conventional “priority” pollutants, to so-called “emerging” or “new” unregulated contaminants. Concerns during this period about the potential health and ecological impacts of exposure to emerging contaminants (ECs) have led to the establishment of new, multi-stakeholder research and testing initiatives, committees, expert groups, newsletters, databases, etc., throughout the world. Up to date, despite these actions, the term “emerging contaminants” remains problematic and sometimes it is difficult to determine which chemicals should or should not be classified as ECs, because they represent a changing reality, dependent on perspective as well as timing. In general, ECs are a structurally diverse and heterogeneous group of chemical compounds, which have widely varying fate properties and adverse effects on environmental ecosystems and can be classified into the following categories: 

“new” ECs, which are chemicals that are recently manufactured and suddenly appear everywhere, and therefore, are not currently covered by existing regulations or legislation  “old” ECs, which are the ones that were actually around for several decades, but simply were not under regular investigation or for which analytical methods did not exist until recently.  “ECs within complex mixtures”, such as industrial effluents, oil residues, hospital effluent, etc. of which either the mixture itself or newly identified (subgroups) of components within may be considered ECs. In recent years, research in all branches of science and technology has been carried out on occurrence, fate and risks of ECs in the environment. Nowadays, their occurrence has been documented worldwide in various compartments of the water cycle including both natural and technical aquatic systems impacted by wastewater discharges and waste disposal sites and it has become a hot topic for environmental analytical chemists. Despite the increasing number of published studies covering EC input, occurrence, fate and effects, there is still a lack of understanding and knowledge about these substances in the aquatic environment. Even more, we know almost nothing about the impacts of the environmental exposure to trace concentrations of their transformation products (TPs) and/or metabolites, but the detection of TPs in the environment is worrying. TPs of ECs in aquatic environments are still rarely considered in water quality and chemical risk assessment, although they have been found in concentrations that are of concern. Since many different TPs can potentially be formed in the environment and analytical standards are typically lacking for these compounds, knowledge on the prevalence of TPs in aquatic environments is fragmentary. In this view, this book intends to gather, specify, synthesize and advance existing knowledge of the most important TPs of the major groups of ECs with potential concern to human health and the environment. The topics covered range from the sources of TPs of ECs and their environmental behaviour, to their occurrence and impacts on engineering systems and natural environment, to risk assessment and management, to the technologies and strategies

xxii Preface

available for control. The objective was to give as much information as possible on TPs of the most potent ECs categories, which nowadays are the most commonly studied and monitored, like for example, TPs of pharmaceuticals and personal care products (PPCPs), hormones, UV filters, perfluorinated compounds, poylybrominated flame retardants (BFRs), pesticides, plasticizers, alkyl phenols, benzotriazoles, artificial sweeteners, illicit drugs, algal toxins, endocrine disruptor compounds (EDCs) etc. The book is divided into 7 sections and 32 chapters, each with a distinct focus organized into two volumes.. In volume 1, the first section with 2 chapters covers general aspects regarding the TPs. In the first chapter the reader finds a classification of the TPs of ECs into their impact on the environment and human health. Chapter 2 deals with the role of TPs of ECs as future groundwater and drinking water contaminants. The second section (chapters 3 to 8) focuses on the fate of TPs in the environment and treatment strategies. In volume 2 the third section with 4 chapters (chapters 9 to 12) is about analytical strategies for identification and structure elucidation. Analytical challenges such as sampling and sample preparation as well as MS and NMR techniques for structure elucidation of TPs, are comprehensively discussed. The fourth section (chapters 13 to 23) is dedicated to the occurrence, fate and effects of TPs in various compartments of the environment by focusing on specific classes of ECs. The fifth and sixth sections (each 4 chapters) cover subjects related to global spaciality and environmental status of TPs in the environment, and risk assessment, management and regulatory framework of TPs. Finally, in chapter 32 (section 7), we summarize and synthesize the major findings and conclusions, and try to predict future trends of discovery, occurrence, fate and risks of TPs of ECs in the environment. This book would be helpful in multifarious ways to analysts, environmental chemists, toxicologists, hydrologists, environment scientists and technologists, engineers, risk assessors, managers of industries, water treatment consultants, firms engaged in water treatment and policy makers. In addition to professionals, anyone with a keen interest in the covered fields, as well as teachers and students at the undergraduate and postgraduate level, would be able to use some of the materials presented here to gain new insights and reach new perspectives in their fields. We hope that most readers will approach this book with knowledge of one or more of the technical areas covered, and hope that after reading this book, they will fill comfortable to discuss and work with experts in all subject areas. We would like to take this opportunity to express our gratitude to Ms. Rebecca Stubbs, Ms. Emma Strickland and Ms. Sarah Tilley as well as their team at John Wiley and Sons, Ltd Publisher, who strongly support the idea and helped make this book a reality. Last but not least, we would like to warmly thank all authors for their excellent contributions to this book which have resulted in an outstanding and superior book. We hope this book will help to keep the environment more green. Science is organized knowledge. Wisdom is organized life. Immanuel Kant April 2013 Dimitra A. Lambropoulou and Leo M.L. Nollet

List of Contributors

Jaume Ace~ na, Department of Environmental Chemistry, IDAEA-CSIC, Spain Ana Ag€ uera L opez, Department of Chemistry and Physics, University of Almerıa, and CIESOL (Solar Energy Research Center), Joint Centre of the University of AlmerıaCIEMAT, Spain Triantafyllos Albanis, Laboratory of Analytical Chemistry, Department of Chemistry, University of Ioannina, Greece Joel Andersen, School of Energy, Environmental Biological and Medical Engineering, University of Cincinnati, USA Maria G. Antoniou, Department of Environmental Science and Technology, Cyprus University of Technology, Cyprus Maria Antonopoulou, Department of Environmental and Natural Resources Management, University of Patras, Greece Adriana Arca-Ramos, Department of Chemical Engineering, University of Santiago de Compostela, Spain Ewelina Baginska, Institute of Environmental Chemistry, Faculty for Sustainability, Leuphana University, Germany Carlos Barata, Department of Environmental Chemistry, IDAEA-CSIC, Spain Damia` Barcel o, Water and Soil Quality Research Group, IDAEA-CSIC, and Catalan Institute for Water Research (ICRA), Cientific and Technologic Park of Girona University, Spain Sara Bogialli, Department of Chemistry, University of Padua, Italy Gabino Bolıvar-Subirats, Department of Environmental Chemistry, IDAEA-CSIC, Spain Vasiliki Boti, Laboratory of Analytical Chemistry, Department of Chemistry, University of Ioannina, Greece Heinz-J€ urgen Brauch, DVGW Water Technology Center, Germany Alejandro Cabrera-Reina, Department of Chemical Engineering, University of Almerıa, and CIESOL, Joint Centre of the University of Almerıa-CIEMAT, Spain Paola Calza, Dipartimento di Chimica, Universita di Torino, Italy Lars Carlsen, Awareness Center, Denmark Serge Chiron, UMR HydroSciences 5569, France

xxiv List of Contributors

Monia Cieszy nska, Department of Environmental Toxicology, Faculty of Health Sciences, Medical University of Gdansk, Poland Montserrat Cortina-Puig, Escola Universitaria Salesiana de Sarria, Spain Adrian Covaci, Toxicological Center, University of Antwerp, Belgium Radosław Czernych, Department of Environmental Toxicology, Faculty of Health Sciences, Medical University of Gdansk, Poland Maria de F atima Alpendurada, IAREN – Water Institute of the Northern Region, and Faculty of Pharmacy, University of Porto, Portugal Armah A. de la Cruz, Office of Research and Development, U.S. Environmental Protection Agency, USA Maria A.D. de Sousa, IAREN – Water Institute of the Northern Region, and Department of Bromatology and Hydrology, Faculty of Pharmacy, University of Porto, Portugal Marıa del Mar G omez Ramos, European Union Reference Laboratory (EURL), Department of Chemistry and Physics, University of Almerıa, Spain Marina DellaGreca, Department of Chemical Sciences, University Federico II, Italy Dionysios D. Dionysiou, School of Energy, Environmental Biological and Medical Engineering, University of Cincinnati, USA Alin C. Dirtu, Toxicological Center, University of Antwerp, Belgium Valeria Dulio, INERIS, National Institute for the Environment and Industrial Risks, France Laura Duran, Parque Cientıfico Tecnol ogico, University of Alcala, Spain Gemma Eibes, Department of Chemical Engineering, University of Santiago de Compostela, Spain Eleni Evgenidou, Department of Chemistry, Aristotle University of Thessaloniki, Greece Debora Fabbri, Dipartimento di Chimica, Universita di Torino, Italy Despo Fatta-Kassinos, Department of Civil and Environmental Engineering, and NIREAS, International Water Research Centre, University of Cyprus, Cyprus Gumersindo Feijoo, Department of Chemical Engineering, University of Santiago de Compostela, Spain Amadeo R. Fern andez-Alba, Department of Chemistry and Physics, University of Almerıa, and CIESOL (Solar Energy Research Center), Joint Centre of the University of AlmerıaCIEMAT, and European Union Reference Laboratory (EURL), Department of Chemistry and Physics, University of Almerıa, Spain ~ez, Plataforma Solar de Almerıa (CIEMAT), and CIESOL, Joint Pilar Fern andez-Ib an Centre of the University of Almerıa-CIEMAT, Spain Imma Ferrer, Center for Environmental Mass Spectrometry, University of Colorado, USA

List of Contributors xxv

Theodora Fotiou, Institute of Advanced Materials, Physicochemical Processes, Nanotechnology and Microsystems, National Center for Scientific Research “Demokritos”, Greece Graz_ yna Gałe˛ziowska, Department of Environmental Toxicology, Faculty of Health Sciences, Medical University of Gdansk, Poland Markus Godejohann, Bruker BioSpin, Germany Carlos GonSc alves, IAREN – Water Institute of the Northern Region, Portugal Tarek Haddad, Institute of Environmental Chemistry, Faculty for Sustainability, Leuphana University, Germany Evroula Hapeshi, Department of Civil and Environmental Engineering, and NIREAS, International Water Research Centre, University of Cyprus, Cyprus Xuexiang He, School of Energy, Environmental Biological and Medical Engineering, University of Cincinnati, USA Ester Heath, “Jozˇef Stefan” Institute, Department of Environmental Sciences, and “Jozˇef Stefan” International Postgraduate School, Slovenia Maria D. Hernando Guil, Spanish National Institute for Agricultural and Food Research and Technology, INIA, Spain Anastasia Hiskia, Institute of Advanced Materials, Physicochemical Processes, Nanotechnology and Microsystems, National Center for Scientific Research “Demokritos”, Greece Alin C. Ionas, Toxicological Center, University of Antwerp, Belgium Marina Isidori, Department of Environmental, Biological and Pharmaceutical Sciences and Technologies, Seconda Universita di Napoli, Italy Aleksandra Jelic, Department of Environmental Chemistry, IDAEA-CSIC, Spain Triantafyllos Kaloudis, Athens Water Supply and Sewerage Company (EYDAP SA), Organic Micropollutants Laboratory, Greece Bulat Kenessov, Center of Physical Chemical Methods of Research and Analysis, al-Farabi Kazakh National University, Kazakhstan Stuart J. Khan, UNSW Water Research Centre, School of Civil & Environmental Engineering, University of New South Wales, Australia Ioannis K. Konstantinou, Department of Environmental and Natural Resources Management, University of Patras, Greece Tina Kosjek, “Jozˇef Stefan” Institute, Department of Environmental Sciences, Slovenia Klaus K€ ummerer, Institute of Environmental Chemistry, Faculty for Sustainability, Leuphana University, Germany Silvia Lacorte, Department of Environmental Chemistry, IDAEA-CSIC, Spain

xxvi List of Contributors

Dimitra A. Lambropoulou, Environmental Pollution Control Laboratory, Department of Chemistry, Aristotle University of Thessaloniki, Greece Dan J. Lapworth, British Geological Survey, UK Holly Lee, Department of Chemistry, University of Toronto, Canada Juan M. Lema Rodicio, Department of Chemical Engineering, University of Santiago de Compostela, Spain Lucı´a Lloret, Department of Chemical Engineering, University of Santiago de Compostela, Spain Thelmo A. Lu-Chau, Department of Chemical Engineering, University of Santiago de Compostela, Spain Scott A. Mabury, Department of Chemistry, University of Toronto, Canada Govindan Malarvannan, Toxicological Center, University of Antwerp, Belgium Sixto Malato, Plataforma Solar de Almerıa (CIEMAT), and CIESOL, Joint Centre of the University of Almerıa-CIEMAT, Spain Maria J. Martınez-Bueno, Department of Hydrogeology and Analytical Chemistry, University of Almerıa, Spain Shigeki Masunaga, Graduate School of Environment and Information Sciences, Yokohama National University, Japan Irene Michael, Department of Civil and Environmental Engineering, and NIREAS, International Water Research Centre, University of Cyprus, Cyprus Sara Miralles-Cuevas, Plataforma Solar de Almerıa (CIEMAT), and CIESOL, Joint Centre of the University of Almerıa-CIEMAT, Spain Mª Teresa Moreira, Department of Chemical Engineering, University of Santiago de Compostela, Spain Jacek Namiesnik, Department of Analytical Chemistry, Faculty of Chemistry, Gdansk University of Technology, Poland Jose M. Navas, Department of Hydrogeology and Analytical Chemistry, University of Almerıa, Spain Leo M.L. Nollet, University College Ghent, Belgium Isabel Oller, Plataforma Solar de Almerıa (CIEMAT), and CIESOL, Joint Centre of the University of Almerıa-CIEMAT, Spain Kevin O’Shea, Department of Chemistry and Biochemistry, Florida International University, USA Sandra Perez Solsona, Department of Environmental Chemistry, IDAEA-CSIC, Spain Kristina Pestotnik, “Jozˇef Stefan” Institute, Department of Environmental Sciences, Ecological Engineering Institute Ltd, and “Jozˇef Stefan” International Postgraduate School, Slovenia

List of Contributors xxvii

Mira Petrovic, Catalan Institute for Water Research (ICRA), Cientific and Technologic Park of Girona University, and Instituci o Catalana de Recerca i EstudisAvanSc ats (ICREA), Spain Alfred Preiss, Fraunhofer Institute for Toxicology and Experimental Medicine, Germany Lucia Prieto-Rodriguez, Plataforma Solar de Almerıa (CIEMAT), and CIESOL, Joint Centre of the University of Almerıa-CIEMAT, Spain Stefano Polesello, Water Research Institute, IRSA-CNR, Italy Jose Benito Quintana, Department of Analytical Chemistry, Nutrition and Food Sciences, IIAA – Institute for Food Analysis and Research, University of Santiago de Compostela, Spain Rosario Rodil, Department of Analytical Chemistry, Nutrition and Food Sciences, IIAA – Institute for Food Analysis and Research, University of Santiago de Compostela, Spain Isaac Rodrıguez, Department of Analytical Chemistry, Nutrition and Food Sciences, IIAA – Institute for Food Analysis and Research, University of Santiago de Compostela, Spain Vasilios Sakkas, Laboratory of Analytical Chemistry, Department of Chemistry, University of Ioannina, Greece Marco Scheurer, DVGW Water Technology Center, Germany Jaroslav Slobodnik, Environmental Institute, Slovak Republic Weihua Song, Department of Environmental Science & Engineering, Fudan University, PR China Marianne E. Stuart, British Geological Survey, UK Roberto Taboada, Department of Chemical Engineering, University of Santiago de Compostela, Spain Fabio Temussi, Department of Chemical Sciences, University Federico II, Italy Frank Thomas Lange, DVGW Water Technology Center, Germany E. Michael Thurman, Center for Environmental Mass Spectrometry, University of Colorado, USA Theodoros M. Triantis, Institute of Advanced Materials, Physicochemical Processes, Nanotechnology and Microsystems, National Center for Scientific Research “Demokritos”, Greece Sara Valsecchi, Water Research Institute, IRSA-CNR, Italy Marlen Ines Vasquez, Department of Civil and Environmental Engineering, and NIREAS, International Water Research Centre, University of Cyprus, Cyprus Davide Vione, Dipartimento di Chimica Analitica, Universita degli Studi di Torino, Italy Dimitra Voutsa, Environmental Pollution Control Laboratory, Department of Chemistry, Aristotle University of Thessaloniki, Greece

xxviii List of Contributors

Lidia Wolska, Department of Analytical Chemistry, Faculty of Chemistry, Gdansk University of Technology, and Department of Environmental Toxicology, Faculty of Health Sciences, Medical University of Gdansk, Poland Bozo Zonja, Department of Environmental Chemistry, IDAEA-CSIC, Spain Yasuyuki Zushi, Center for Environmental Measurement, Organochemical Measurement Laboratory, National Institute for Environmental Studies, Japan

Part I General Considerations

1 Classifying the Transformation Products (TPs) of Emerging Contaminants (ECs) for Prioritizing Research into their Impact on the Environment and Human Health Jacek Namie
snik1, Lidia Wolska1,2, Radosław Czernych2, nska2 Gra_zyna Gałe˛zowska2 and Monika Cieszy
1

Department of Analytical Chemistry, Faculty of Chemistry, Gdansk University of Technology, Poland 2 Department of Environmental Toxicology, Faculty of Health Sciences, Medical University of Gdansk, Poland

1.1 Introduction Continuous degradation of an abiotic part of an environment as well as disruption of the homeostatic state of biota can be attributed to the increasing rate of human impact on the environment. An idea about the diversity of pollutants that can potentially appear in the environment can be gained by monitoring the amount of already existing chemicals and the rate of appearance of new ones. If a compound is known and in use it seems obvious that in the end it will be introduced into the environment. Figure 1.1 shows a classification of environmental pollutants. Analysts have concentrated mainly on:


Non-identified pollutants – pollutants that exist in the environment for a long time but due either to lack of proper analytical procedures, insufficient development of analytical devices, or lack of new methodological parameters have not yet been identified.

Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

4 Transformation Products of Emerging Contaminants in the Environment

Figure 1.1 Classification of environmental pollutants.




New emerging contaminants (ECs) that start to appear in the environment due to development of new technological processes and other human activities.

Although there are numerous publications on the monitoring of ECs, the vast majority of them do not address the monitoring of metabolites and transformation products (TPs) of parent compounds. The problem must be addressed if we consider that the properties and occurrence of TPs in an environment may vary considerably from those of their parent compounds. ECs, being a subject of intense examination, can be subdivided into four groups: veterinary and human antibiotics, human drugs, industrial and household wastewater products, and sex and steroidal hormones. Each of these groups can be subdivided into groups of compounds with varying chemical characteristics and biological activity. This division has been suggested by the United States Geological Survey (USGS) together with the United States Environmental Protection Agency (US EPA) and, among any other suggested classifications, this one not only indicates the origin of ECs, but also shows that contaminants defined as “emerging” are usually biologically active (drugs, antibiotics, hormones, endocrine-disrupting chemicals). It is commonly thought that the parent compounds may exist in smaller amounts and be less toxic than their TPs. This is not always true. In fact the literature study carried out in this chapter shows that there is no common pattern. ECs, depending on their stability in an environment and metabolic pathways, may exist in either substantially smaller or larger amounts than their transformation products or metabolites. Thus they may have smaller or larger environmental impact. Therefore, it is essential to develop reliable methods of qualitative and quantitative analysis, not only of ECs but also of their metabolites and TPs. Lack of such methods and knowledge about their environmental and health impact makes TPs and metabolites of ECs much more dangerous than ECs themselves. The aim of this chapter is to carefully investigate the modern scientific literature containing crucial information about the physiochemical properties and toxicological data of chosen ECs, with particular attention paid to TPs of ECs or their metabolites. Gathering such data will show the present state of the art for analytical and toxicological assessment of ECs and their TPs, and finally indicate which compounds should have the highest priority in setting legislative regulations and restrictions as well as developing new analytical methods.

Classifying the Transformation Products 5

1.2 Emerging Contaminants – Emerging Problem ECs constitute a very broad group of compounds that can be divided into distinct classes which can be further divided into into sub-classes. ECs and their TPs possess varying physiochemical characteristics and toxicological profiles. The EPA defines ECs as: “a chemical or material characterized by a perceived, potential, or real threat to human health or the environment or by a lack of published health standards. A contaminant also may be ’emerging’ because of the discovery of a new source or a new pathway to humans.” This definition well summarizes the complex nature of ECs. However, it has been made even more precise by The Massachusetts Department of Environmental Protection (MassDEP): “emerging contaminants have been defined as hazardous materials or mixtures (naturally occurring or manmade chemical, microbial or radiological substances) that are characterized by having: a perceived or real threat to human health, public safety or the environment, no published health standards or guidelines, insufficient or limited available toxicological information or toxicity information that is evolving or being re-evaluated, or significant new source, pathway, or detection limit information.” This comprehensive definition touches a very important aspect of ECs, namely: the lack of knowledge about their (eco-)toxicity. It should also be emphasized that ECs cannot only be regarded as single compounds but also as mixtures of ECs, occurring either naturally or introduced into the environment as xenobiotics (anthropogenic activity), as a consequence of metabolism, oxidation, degradation or other transformations of the parent compound. The research on ECs developed after an improvement in modern analytical techniques, such as liquid chromatography and gas chromatography coupled to tandem mass spectrometry, which made it possible to detect trace levels of ECs and characterize their TPs in various environmental samples. Sources of ECs are varied and include agriculture and aquaculture, animal husbandries, life-style and domestic use compounds, household and industrial wastewater discharges, industrial additives and by-products, and compounds found in sewage sludge or manure applied on lands (see Figure 1.2). It should be stressed, however, that for

Figure 1.2 Origin and routes of ECs.

6 Transformation Products of Emerging Contaminants in the Environment

many ECs the pathway from the pollution source to the receptor is very ambiguous, due to lack of information about the nature and environmental fate and effect of many newly introduced compounds. Another problem with ECs is that conventional sewage treatment plants (STPs) fail to remove a large portion of ECs from sewage. These survive the STP biodegradation process and thus may enter the aquatic systems with the STP effluents. The problem with ECs in the environment is that until now many detected ECs have escaped regulations and routine monitoring, and only some have been proposed to be included in the regulatory list [1,2]. A brief description of some representative ECs is given in the next sections. 1.2.1 Veterinary and Human Antibiotics Antibiotics are complex molecules of natural or semi-synthetic origin which were developed with the intention to stimulate physiological response in target organisms. They are therapeutic agents used to prevent and treat various infectious diseases by either inhibition of the growth rate of pathogenic microorganisms, reduction of their metabolic activity, or simply killing. Antibiotics are commonly applied in human and veterinary medicine, but they are also used in agriculture to enhance animal growth, treat vegetable diseases, and in aquaculture to overcome fish infections [3]. Antibiotics have a tendency to bioaccumulate and provoke many effects in the environment, which is a reason why they are listed among important ECs in the environment. They also combine a high potential for entering the environment, with many unresolved issues and problems on the extent of occurrence, transport routes, environmental fates (e.g., degradation, adsorption, leaching, metabolism), ecological risks and public health effects [4,5]. Extensive antibiotics application also induces antimicrobial resistance in bacteria in the environment and develops horizontal transfer of resistant genes to pathogenic bacteria which, unaffected by antibiotics, are then capable of infecting plants, animals and humans with impunity. Antibiotics also affect soil and water natural microbial activities. It has been observed that they can considerably influence organic matter decomposition and modify key fate processes of chemical substances in the environment. Thus, concerning the biological effect, even low levels of antibiotics cannot be excluded a priori from the study. The antibiotics are thought to target bacteria, but released to the environment they can affect non-target organisms such as soil-dwelling earthworms, plants, fish, and invertebrates [6,7]. 1.2.2 Human Drugs Among all emerging substances in the aquatic environment, the occurrence of pharmaceutical compounds and their metabolites or/and TPs has become an increasing concern over recent years. Exact assessment of the pharmaceutical impact on the environment is difficult due to the multiplicity of input sources to the aquatic system, with no evident quantitative data available concerning the relative distribution of pharmaceuticals from all emission sources (Figure 1.2). Although humans and animals treated with pharmaceutical products are the main contamination source of potable water resources, drugs are qualitatively, quantitatively, spatially and temporally shared out into different routes, depending on where the patients are located. Different studies reveal the release of pharmaceuticals in wastewater effluent, surface waters,

Classifying the Transformation Products 7

drinking water sources and even in some treated drinking waters. Pharmaceuticals are typically released into the aqueous systems following their ingestion and subsequent excretion through the wastewater treatment plants or, for countryside households, directly released into septic tanks where overflows may release raw wastewater with drugs in the case of heavy rain. Additionally, the inappropriate disposal of expired or unused drugs in households and hospital facilities by flushing down toilets or sending to landfill, and pharmaceutical residues from accidental spillages at manufacturing sites may also contribute to the overall burden. Depending on the nature of the drugs and on the process design, the elimination rates may be anywhere in the range 0–100% and are then found with other not removed pharmaceuticals in treated waste water treatment plant (WWTP) discharges in rivers, lakes, streams, and estuaries. Groundwater and sewage waters are physically closely linked and, therefore, can contaminate one another. These facts have raised substantial concern in public and regulatory agencies, even though pharmaceuticals are often detected only at trace levels they are very important in terms of the high environmental potential risk and the possible chronic adverse effects on humans [8–11]. Waste waters contain by-products, such as metabolites excreted via urine or feces and TPs formed under biological and physicochemical conditions (potential route presented in Figure 1.3) present in WWTP [12]. However, very often many of the excreted pharmaceuticals are only slightly transformed or even unchanged, mostly conjugated to polar molecules (e.g., as glucoronides). These conjugates can easily be cleaved during sewage treatment and the original drugs will then be released into the aquatic environment, mostly in effluents from WWTP [8–11,13,14].

Figure 1.3 Transformation pathways of pharmaceutical products.

8 Transformation Products of Emerging Contaminants in the Environment

1.2.3 Industrial and Household Wastewater Products Plastics additives encompass four main functional categories: property modifiers, property extenders, stabilizers, and processing aids. Property modifiers include flame retardants, antioxidants, antimicrobials, plasticizers, blowing agents and impact modifiers. Modifiers are rarely used alone, being typically added to plastics in combination with processing aids and property extenders. Plasticizers are the fastest-growing segment among property modifiers, followed by flame retardants. The development of end-use markets for plastics means that the consumption of plastics additives continues to increase. The global revenue for the industry reached US$39.7 billion in 2011, and with a forecast compound annual growth rate of 3.8% will reach $47.8 billion by 2016 [15]. Organophosphate esters (OPs) are a group of important chemicals widely used as flame retardants and plasticizers in a variety of products, such as polyurethane foams, electronics equipment, upholstery, and textiles. Chlorinated OPs, such as tris (2-chloroethyl) phosphate (TCEP), and tris (1,3-dichloro-2-propyl) phosphate (TDCPP), are preferentially used as flame retardants, while nonchlorinated OPs, such as tris (2-butoxyethyl) phosphate (TBEP) and triphenyl phosphate (TPhP), are applied as plasticizer additives and also as extreme pressure additives and antiwear agents in hydraulic fluids, lubricants, transmission fluids, motor oils, and similar products [16,17]. OPs can undergo diffusion processes and be emitted into the environment. Because of the large variation in physico-chemical properties, OPs can be transported in different environmental media and have been detected in various environmental areas. High concentrations, up to a few mg/m3, of ethanol-2-butoxy-phosphate (TBEP) were detected in schools [18,19] and up to 0.73 g/m3 in offices [20], whereas in outdoor air concentrations of a few ng/m3 were measured [21]. In water concentrations of TBEP were calculated to be 300–35 000 ng/l [22,23]. For a different type of plasticizer (triphenylphosphate TPhP) the contents in air, water and sediment were 23.2 ng/m3, 7900 ng/l and up to 4000 ng/g, respectively [24]. Bacoloni et al. in 2007 found that the concentration of TPhP in surface water was 165 ng/l. Generally, water concentrations of plasticizers in the environment are low and toxic effects on aquatic organisms are unlikely (TBEP: Daphnia magna/LC50/24 h and 48 h/N/A, NOEC/32 mg/l [25]; fathead minnow (Pimephales promelas)/LC50/96 h/16 mg/l [26]; killifish (Oryzias latipes)(10, 20, 30  C)/LC50/48 h/44 mg, 27 mg, 6.8 mg/l [27]; goldfish (Carassius auratus)/168 h/N/A at 5 mg/l [28]; rainbow trout (Oncorhynchus mykiss)/LC50/ 96 h/24 mg/l, NOEC/10 mg/l [29]; TPhP: rainbow trout/LC50/96 h/0.36 mg/l [24], daphnia/ NOEC/0.1 mg/l [30]). These substances are removed rapidly from the tissues of fish when exposure ends and bioconcentration factors are moderate (TPhP: BCF 113 [31], bioaccumulation is not considered to be a hazard. The same situation holds for flame reagents where BCF for tris(2-chloroethyl)phosphate (TCEP) and tri(dichlorisopropyl)phosphate (TDCPP) were calculated from 1.4 to 107 [32]. The concentrations of flame retardants in the environment are similar to the concentrations of plasticizers (Table 1.1). All environmental plasticizers and flame retardants derive from human activities but the input rate to the environment cannot be estimated from the available data. The input is expected to be mainly to soil, sediments and surface waters from leachates from plastics on landfills, from spillages, and from effluents. The low vapor pressure, the high soil sorption coefficient (Koc) and the water solubility suggest that substances in the environment will be found mainly in water and sediment.

40 Transformation Products of Emerging Contaminants in the Environment

Many chemicals, even banned ones, persist in the environment and continue to accumulate in ecosystems. Pesticides with high water solubility, low tendency to adsorb to soil particles, and long persistence or half-life have the highest potential to move into water. These three factors, soil adsorption, water solubility and persistence, are commonly used to rate pesticides for their potential to leach or move with surface runoff after application. Carbaryl is considered moderately to highly toxic to fish – Rainbow Trout/96 h/LC50/ 4.38 ppm, Bluegill Sunfish/96 h/LC50/6.76 ppm, Goldfish/96 h/LC50/13.2 ppm [210]. It is highly toxic to the aquatic invertebrate Daphnia magna – Daphnia Magna/48 h/LC50/ 18.6 ppb, Daphnia Magna/48 h/EC50/0.26 ppb [209]. The concentrations of carbaryl in air were 0.0035 to 0.107 mg/m3 [205]. The half-life coefficient depends on the pH of the water (t1/2 hydrolysis (pH ¼ 5/7/9) ¼ >1500 d [212]/12.1 d/3.2 h [213]. This relationship is also correct for other pesticides, for example diazinone t1/2(water) ¼ 12 h (highly acidic) 6 months (neutral) [195]. Contamination of water by pesticides is mainly due to runoff, usually within a few weeks after application, diazinon in water (river): 0 (winter) 26.7 (summer) mg/l [188], and carbaryl in water: sunlight 10.1 ppm and pH 5 [206]. 1.2.4 Sex and Steroidal Hormones Hormones occur in an environment both as natural and synthetic compounds. They possess a very high biological activity and thus are able to exert a biological effect even at very low doses. Once introduced into the environment hormones may interfere with endogenous hormones (endocrine-disrupting compounds – EDCs). The most commonly occurring hormones are estrogens, namely: ethinylestradiol (EE2) and 17b-estradiol (E2). The disruptive effect of environmental estrogens has been investigated by monitoring physiological changes in fish exposed to elevated amounts of estrogens, both naturally and in controlled laboratory conditions. Fish exposed to the synthetic drug EE2 exhibited various physiological and behavioral changes, such as decrease in fertilization rate, distorted sex proportions, and partial and complete feminization of exposed specie [109]. Fish feminization is usually manifested by the development of testis-ova at an initial phase of exposure and development of an ovary in a terminal phase or absence of secondary male characteristics. Similar endocrine-disrupting effects (sex transition, distortions in secondary sex characteristics development) could have been observed in amphibians and reptiles exposed to environmental estrogens. Environmental monitoring (Table 1.1) shows that either environmental or synthetic hormones are found in almost any type of environment, but especially in an aquatic environment. Each type of hormone has been found in surface waters (rivers and streams), oceans, and sediments in concentrations high enough to exert physiological effects on fish and other aquatic vertebrates. Trace amounts of estradiol, progesterone and ethinylestradiol have been found in drinking water reservoirs. Scientists insist that such amounts of hormones present in drinking water cannot inflict any biological effect on humans. Nevertheless, it is crucial to point out here that hormones, as non-polar compounds, are highly lipophilic, therefore undergoing the process of bioaccumulation (by drinking water polluted with environmental or synthetic hormones) and biomagnifications (increase in concentration of a pollutant along food chain trophic levels). Therefore, they can eventually exist in substantial amounts within fatty tissue of not only aquatic vertebrates but also humans. This prompts the idea of enhancing water treatment technologies so as to minimize or eradicate the presence of any type of hormone from effluents from WWTPs.

Classifying the Transformation Products 41

1.3 Transformation Products of ECs 1.3.1 Veterinary and Human Antibiotics After being taken, antibiotics may be excreted through urine or feces (depending on lipophilicity) as: an unchanged parent compound, a conjugated metabolite or other TP with an inactivating substituent (depending on the pharmacology) attached to the molecule. When administered, antibiotics are metabolized in the target organisms’ bodies; however, metabolism of an antibiotic is frequently incomplete. Studies have shown that 60–80% of an antibiotic is excreted unchanged as an active parent compound. For some antibiotics it was also noted that excreted inactive metabolites can be transformed back to the original bioactive parent compound under mild conditions. Intensive metabolism was mentioned for sulfamethoxazole (15% of parent compound released) [33,280–282]. After release to the environment antibiotics or their metabolites may be biochemically transformed in sewage treatment plants or in the environment. Hydrophobic antibiotics can be bound to particles, for example, by sorption to sediments (fluoroquinolones), others may form complexes with various ions (e.g., tetracyclines with magnesium and calcium) or, if sensitive to light, they can undergo phototransformation (quinolones, tetracyclines, sulfonamides). Fluoroquinolones are susceptible to photodegradation by UV-radiation. Water soluble antibiotics, such as tetracyclines and sulfonamides, may be submitted to hydrolysis [283]. Most of the antibiotics used are not easily biodegraded in field studies or laboratory tests. Environmental degradation rates and soil absorption rates may vary significantly for different antibiotics. Trimethoprim is not mineralized at all, tylosin A has been shown to be biodegraded after 24 h in the laboratory tests [284]. Enrofloxacin and ciprofloxacin may be degraded by various types of fungi (Mucor ramannianus, Gloeophyllum striatum) [72,75,285] while chlortetracycline and sulfamethazine may be absorbed by some plants. It can be concluded from Table 1.1 that for many of the antibiotics (especially those used in agriculture) TPs are not known, not certain, or described only by molecular mass of “unknown degradate.” It is also evident that the toxic properties of most TPs of antibiotics are insufficiently examined and described in the literature. The toxicity of TPs may be different to or even greater than that of the parent molecule at the same concentration level. Another issue is that even if individual antibiotics are assessed for their quality, efficacy, toxicity, and safety, the information is not sufficient since they are rarely used alone, but rather in combination with others to improve their action (e.g., trimethoprim with sulfonamides). This may significantly modify the properties and ecotoxicity values noted for a component in the natural environment mixture of pure antibiotics, their metabolites and TPs (additivity, synergism) [85,86]. 1.3.2 Human Drugs Since drugs, in a prevailing number of instances, are introduced into the municipal sewage system with urine and feces, they will exist in completely or partially metabolized form, whereas environmental abiotic degradation will not constitute a very important process of drugs transformation. Drugs can also be introduced into the environment directly. However, this takes place to a much smaller extent. In the first phase of metabolism a C H bond is converted to C OH via an oxidation reaction. In the second phase a drug is conjugated with glucuronic acid, sulfonates, glutation or amino acids, thus reducing the compound’s biological activity. Due to the two-phase metabolization process metabolites become more polar

42 Transformation Products of Emerging Contaminants in the Environment

and thus are more easily excreted with urine. An increase in polarity is manifested by a decrease in the partition coefficient (Ko/w) of metabolites with respect to parent compounds (Table 1.1). Digoxine, having a partition coefficient Ko/w ¼ 1.26, as a drug dissolves slightly better in fats, whereas its metabolite – digoxigenin – dissolves almost equally in water and in fats. A similar situation can be observed in the case of the analgesic drug acetaminophen (Ko/w ¼ 0.46) and its metabolite 1,4-benzoquinone (Ko/w ¼ 0.2) and N-acetyl-p-benzoquinone (Ko/w ¼ 0.45), as well as codeine (Ko/w ¼ 1. 19) and norcodeine (Ko/w ¼ 0.69) and morphine (Ko/w ¼ 0.89). Comparing concentrations of drugs and their metabolites in an aquatic environment it can be observed that the metabolites may at times exist in concentrations a few times higher than their parent compound. Weigel et al. [136,137] conducted research on the presence of ibuprofen and its metabolites, hydroxy- and carboxy-ibuprofen, in surface waters (rivers, lakes) in which he showed that concentrations of ibuprofen metabolites are three times higher than that of ibuprofen itself. A year later (2005) Weigel’s findings were acknowledged by Bendz et al. [135]. Having analyzed changes in concentration of chosen drugs in wastewater influents and effluents, certain conclusions can be drawn: (i) drugs present in wastewater are not completely removed from it, (ii) drugs’ metabolites are hardly removed at all, (iii) drugs subjected to biological treatment in WWTPs may undergo further transformation/metabolization that results in an increase in metabolites concentration in wastewater effluents when compared with influents. It should also be emphasized that metabolites/TPs are not only present in higher concentration than their parent compounds, but they may also often be characterized by higher toxicity (Table 1.1: metformin, ranitidine, diltiazem, acetaminophen – LD50, EC50). What is more, many of these drugs are non-prescription drugs, which means that controlling their traffic is practically impossible. This results in an extensive collection of drugs at home and further disposal of expired ones, which naturally contributes to an increased drugs concentration in the environment. 1.3.3 Industrial and Household Wastewater Products Pesticides have been detected in the environment. Their fate depends on volatilization, hydrolysis, and photolysis. In an oxidation process dianizon is transformed to diazoxon [186], in hydrolysis to 3,5,6-trichloro-2-pyridinol (TCP) [187] and 2-isopropyl-6-methyl-4-pyrimidinol (IMP) [186]. Microbes play a significant role in the degradation of pesticides in soil, including carbaryl. IMP was reported as a major TP of diazinon in compost, soil and water. IMP and TCP are regarded as less toxic than their parent compound – TCP – Pimephales promelas (fathead minnow)/LC50/48 h/125 mg/l; fish/LC50/96 h/1.8–2.7 mg/l [201] and IMP- Lepomis macrochirus/LC50/96 h/1.2 mg/l. The other product, diazoxon, is more toxic than its parent compound – killifish/LC50/48 h/0.22 mg/l [200]; Rana boylii/LC50/96 h/0.760 mg/l [194]. pH, temperature and buffer capacity of an aqueous solution did not affect the efficiency of ozonation of diazinon. TCP is also a TP of chloropyrifos. 1.3.4 Sex and Steroidal Hormones An environmental and health hazard imposed by hormones and their TPs derives not from their toxicity but rather their endocrine-disrupting effect. In fact, data on the toxicological

Classifying the Transformation Products 43

effect on living organisms are rather scarce, and limited to the LD50 parameter. However, it seems useful to take into consideration the very high values of the BCF (bioconcentration factor). The lipophilic character of hormones is due to their characteristic steroidal structure. Therefore, even existing in the environment at very low concentrations but being continuously accumulated in fatty tissue may lead to hormones existing in substantial amounts within living organisms. Hormones, regardless of their class, have a very high soil organic carbon/water partition coefficient that provides the potential to adsorb to soil particles and sediments. For this reason, hormones can be found not only in surface waters but occasionally also in sediments (Table 1.1: cholesterol). The literature study showed that, presumably due to their low solubility in water, concentrations of hormones in water remain at trace levels. Comparison of the occurrence of cholesterol and its metabolites: coprostanol and cholestanol, show that both metabolites predominate in sediments [276]. Similar disproportions can be observed in the case of other hormones and their metabolites (estradiol and estriol, [262]).

1.4 Minimizing Environmental Risk of ECs and their TPs 1.4.1 Designing a Risk Minimization Strategy The prioritization procedure has been based upon data of occurrence in an environment and toxicological assessment of generic groups of ECs (stage 1) as well as their TPs (stage 2). Taking into consideration the points obtained from stages 1 and 2, a final priority list for monitoring and environmental impact assessment of ECs has been made. The first stage (Figure 1.4) included the analysis of environmental occurrence and potential hazard imposed by ECs. By utilizing data on presence, and pathways of entry to the environment (Table 1.1), the ECs that were considered to reach the environment in potentially significant amounts were identified. Groups of substances were

Figure 1.4 Environmental occurrence and potential hazard evaluation matrix.

44 Transformation Products of Emerging Contaminants in the Environment Table 1.2 Classification of compounds occurrence levels. Amounts Trace Microtrace Nanotrace

Upper limit

Lower limit

– 0.01 mg/l 0.01 ng/l

0.01 mg/l 0.01 ng/l –

Table 1.3 Toxicity ranges for different groups of organisms. Organism (time of exposure)

Parameter (units)

Mammals Birds Fish (96 h) Aquatic invertebrates (48 h) Aquatic crusteceans (96 h) Algae (72 h)

LD50 (mg/kg) LD50 (mg/kg) LC50 (mg/l) EC50 (mg/l) LC50 (mg/l) EC50 (mg/l)

Low

Moderate

High

>2000 >2000 >100 >100 >100 >10

100–2000 100–2000 0.1–100 0.1–100 0.1–100 0.01–10

diclofenac. Maximum concentrations of antibiotics measured in groundwater from published studies ranged from 5.7 to 2  103 ng/L (Figure 2.2) with the major sources being waste water sources, landfills, septic tanks and animal waste lagoons [45–47]. The three most commonly reported antibiotics were found to occur in the following order of maximum concentration, triclosan > sulfamethoxazole > lincomycin > erythromycin [3]. Veterinary medicines are an important sub-group of pharmaceutical compounds. Detection of these compounds is often associated with leaching from farm waste lagoons, associated with concentrated animal feeding operations in the USA, although they have also been reported in groundwater reconnaissance studies in Germany, the USA and Switzerland [48– 50]. The antibiotic sulfamethazine has been reported in groundwater in at least five separate studies, with maximum concentrations ranging from 120–616 ng/L [3]. Two other veterinary antibiotics reported in groundwater include monensin and tylosin [51,52]. A number of laboratory column experiments have been carried out to investigate the transformation of pharmaceuticals in soil and aquifer material, but very few studies have reported the occurrence of these compounds in groundwaters. For example Barbieri et al. [53] showed the quantitative transformation of diclofenac and sulfamethoxazole to nitro-diclofenac and 4-nitro-sulfamethoxazole, respectively, in biotic nutrient rich conditions. Clofibric acid, the metabolite of clofibrate (cholesterol regulator), has been detected in soil pore waters at concentrations between 0.6 and 143 mg/L at a site in Germany where sewage effluent had been applied to fields [54]. A recent microcosms degradation study of the analgesic compounds phenazone and propyphenazone and their degradates by Burke et al. [55] found that the degradation of six of the seven investigated compounds was strongly influenced by the prevailing redox conditions. One study in Berlin, Germany, detected metabolites of phenazone drugs (analgesics) in groundwater impacted by spills at a pharmaceutical production plant [44]. Three phenazonetype metabolites, 1-acetyl-1-methyl-2-dimethyl-oxamoyl-2-phenylhydrazide (AMDOPH), 1-acetyl-1-methyl-2-phenylhydrazide (AMPH), and dimethyloxalamide acid-(N-methyl-Nphenyl)-hydrazide (DMOAS), were detected in aerated groundwater in the following

30 26 22 18 1

4

7 10

14

Major EC groups

34

38

42

46

72 Transformation Products of Emerging Contaminants in the Environment

0.1

1

10

100

1000 10000

1e+06

Max. EC concentration in groundwater [ng/L]

Figure 2.2 Box plot of maximum concentrations of parent compounds in groundwater for major groups of ECs, including PPCPs, industrial compounds, steroids and hormones, artificial sweeteners and preservatives. Note the log scale on the y-axis. Suspected outliers (þ) are 25th and 75th percentile þ/1.5(IQR). For comparison, the solid horizontal line is the EU drinking water limit for pesticides. The data are from a BGS database of published studies worldwide (14 different countries) investigating ECs in groundwater from. Explanation of groups of compounds on y axis: 1 ¼ alcoholism treatment, 2 ¼ analgesic, 3 ¼ anti-inflamatory, 4 ¼ antianginal, 5 ¼ antiarrhythmic, 6 ¼ antibiotic, 7 ¼ anticoagulant, 8 ¼ anticonvulsants, 9 ¼ antidepressant, 10 ¼ antidiabetic, 11 ¼ antioxidant, 12 ¼ antipruritic, 13 ¼ artificial sweetener, 14 ¼ barbiturate, 15 ¼ beta-blocker, 16 ¼ blood pressure/hypertension, 17 ¼ soil fumigant,18 ¼ coccidostat, 19 ¼ corrosion inhibitor, 20 ¼ detergents, 21 ¼ diuretic, 22 ¼ dye, 23 ¼ fire retardant, 24 ¼ fluorescent whitening agent, 25 ¼ food additive, 26 ¼ fragrances, 27 ¼ fungicide, 28 ¼ glaucoma treatment, 29 ¼ illicit substance, 30 ¼ insect repellent, 31 ¼ insomnia drug, 32 ¼ lipid regulators, 33 ¼ metabolic (diet pills), 34 ¼ muscle relaxant, 35 ¼ skin cosmetic, 36 ¼ plasticiser, 37 ¼ psychiatric drug, 38 ¼ scabicide/miticide, 39 ¼ sedative, 40 ¼ solvent plasticiser and anti-foamiing, 41 ¼ solvent stabiliser, 42 ¼ steroids and steroidal hormones, 43 ¼ stimulant, 44 ¼ sunscreen, 45 ¼ surfactant, 46 ¼ veterinary Medicine, 47 ¼ X-ray contrast media.

concentrations, 1200, 20–100 and about 10 mg/L, respectively. At the same site, Zuehlke et al. [56] found the occurrence of metabolites of phenazone-type pharmaceuticals (analgesics) significantly higher than the parent compounds in treated groundwater as a result of degradation during filtration. AMPH, and the TPs of dimethylaminophenazone (4-acetylaminoantipyrine and 4-formylaminoantipyrine) were reported in Dutch river bank filtrates with the following maximum concentrations, 109, 20 and 45 ng/L, respectively [57].

Transformation Products of Emerging Organic Compounds 73

2.4.3 Personal Care Products and Synthetic Musks A range of compounds associated with skin care products, including UV blockers (oxybenzone and drometrizole), isopropyl myristate, phenoxy-ethanol and lilial have been detected in groundwater in a limited number of case studies [5,135]. Paraben compounds (fungicide/ microbiocide) which are used in food, creams and other personal care products, are also found in groundwater, see Figure 2.2. For example methylparaben has been detected relatively frequently (about 2%) in UK groundwaters [135]. The occurrence of these types of compounds in groundwater is often associated with waste water sources, such as septic tanks and artificial recharge of treated waste water. In the past the antimicrobial triclosan, now thought to be an endocrine disrupting compound (EDC) [58], was used in personal care productss such as hand soaps and toothpaste. Polycyclic musks, including galaxolide (HHCB), tonalide (AHTN), celestolide (ADBI) and phantolide (AHDI), and the nitro musks (musk xylene and musk ketone) are used as fragrances for personal care and household products. Their route into the environment is therefore predominantly in wastewater and high concentrations are found in the influents and effluents of treatment plants [59,60]. Chase et al. [61] detected traces of HHCB, AHTN, ADBI, and AHDI in groundwater samples below a wastewater land application site, mainly at 3:1 107 ½NO NPOC > 3  þ 5:7 10
> ½ OH ¼ > 4 > > jðICÞ þ 5:0 10 NPOC þ 5:2 107 ½NO 3 <   IC 8:5 106 10pH þ 0:025 jðICÞ ¼ > > 10:2 pH > 12000 ð10 þ 10 Þ > > > : pH ¼ 1:95 1  100:075 IC  þ 6:32

(3.30)

The cited series of approximations finally allows a manageable equation to be obtained, by which [
NO2 ] can be plotted as a function of nitrite (nitrate), at the fixed ratio shown above, of NPOC and of IC. Figure 3.18 reports [
NO2 ] vs. NPOC and nitrate, Figure 3.19 reports [
NO2 ] vs. NPOC and IC. One can see that [
NO2 ] obviously increases with increasing nitrate and nitrite, while it decreases with NPOC, in particular at high nitrate/nitrite. The most likely explanation is
OH scavenging by DOM, which inhibits reaction (3.25). Scavenging of
OH is also the most likely explanation of the decrease of [
NO2 ] with IC. Interestingly, [
NO2 ] decreases with NPOC at low IC (NPOC measures DOM that is a major
OH scavenger), but slightly increases with NPOC at high IC. In the latter case, most
OH is scavenged by IC and the oxidation of nitrite by 3 CDOM , which would be favoured at high NPOC, could become more important as an
NO2 source.

3.4 Towards the Modelling of Phototransformation Kinetics in Surface Water It is possible to model the transformation kinetics of a substrate, a generic pollutant P, in surface water as a function of water chemistry and substrate reactivity, via the main photochemical reaction pathways (direct photolysis and reaction with
OH, CO3
, 1 O2 and 3 CDOM  ). The reaction kinetics is modelled within a cylindrical volume of 1 cm2 surface area and depth d. The model may use actual data of the water absorption spectrum, or it can approximate the spectrum from dissolved organic carbon (DOC) values. The model will now be described in greater detail.

Phototransformation Processes of Emerging Contaminants in Surface Water 107

Figure 3.18 Trend of [
NO 2 ] as a function of nitrate and NPOC, in the presence of constant IC ¼ 10 mg C L1. Sunlight UV irradiance: 22 W m2.

Figure 3.19 Trend of [
NO 2 ] as a function of IC and NPOC, in the presence of constant 10 mM nitrate. Sunlight UV irradiance: 22 W m2.

108 Transformation Products of Emerging Contaminants in the Environment

3.4.1 Surface-Water Absorption Spectrum It is possible to find reasonable correlation between the absorption spectrum of surface waters and their content of dissolved organic matter, expressed as NPOC, which is a measure of DOM. The following equation holds for the water spectrum, referred to an optical path length of 1 cm [34]: A1 ðlÞ ¼ ð0:45  0:04Þ NPOC eð0:0150:002Þ l

(3.31)

As an obvious alternative, A1(l) can be spectrophotometrically determined on real water samples. 3.4.2 Reaction with
OH [34] In natural surface waters under sunlight illumination, the main
OH sources are (in order of average importance) CDOM nitrite and nitrate. All these species produce
OH upon absorption of sunlight. The calculation of the photon fluxes absorbed by CDOM, nitrate and nitrite requires taking into account the mutual competition for sunlight irradiance, also considering that CDOM is the main absorber in the UV region where nitrite and nitrate absorb radiation as well. At given wavelength l, the ratio of the photon flux densities absorbed by two different species is equal to the ratio of the respective absorbances. The same is also true of the ratio of the photon flux density absorbed by species to the total photon flux density absorbed
by the solution, ptot a ðlÞ [76]. Accordingly, the following equations hold for the different OH sources (note that A1(l) is the specific absorbance of the surface water layer over a 1 cm optical path length, in units of cm1; d is the water column depth in m; Atot(l) the total absorbance of the water column, and p (l) the spectrum of sunlight): Atot ðlÞ ¼ 100 A1 ðlÞ d

(3.32)

ANO3 ðlÞ ¼ 100 eNO3 ðlÞ d ½NO 3

(3.33)

ANO2 ðlÞ ¼ 100 eNO2 ðlÞ d ½NO 2

(3.34)

ACDOM ðlÞ ¼ Atot ðlÞ  ANO3 ðlÞ  ANO2 ðlÞ Atot ðlÞ

(3.35)

 Atot ðlÞ ptot Þ a ðlÞ ¼ p ðlÞ ð1  10

(3.36)

1 pCDOM ðlÞ ¼ ptot

ptot a a ðlÞ ACDOM ðlÞ ½Atot ðlÞ a ðlÞ

(3.37)

NO 2

1  ðlÞ ¼ ptot a ðlÞ ANO2 ðlÞ ½Atot ðlÞ

(3.38)

NO 3

1 ðlÞ ¼ ptot a ðlÞ ANO13 ðlÞ ½Atot ðlÞ

(3.39)

pa

pa

An important issue is that p (l) is usually reported in units of E cm2 s1 nm1 (see for instance Figure 3.20), thus the absorbed photon flux densities are expressed in the same units. To express the formation rates of
OH in M s1, the absorbed photon fluxes Pia should be expressed in E L1 s1. Integration of pia ðlÞ over wavelength would give units of E cm2 s1 that represent the moles of photons absorbed per unit surface area and unit time. Therefore, assuming a cylindrical volume of unit surface area (1 cm2) and depth d (expressed in m), the absorbed photon fluxes in E L1 s1 units would be expressed as

Phototransformation Processes of Emerging Contaminants in Surface Water 109

Figure 3.20 Sunlight spectral photon flux density at the water surface per unit area. The corresponding UV irradiance is 22 W m2 [77].

follows (note that 1 L ¼ 103 cm3 and 1 m ¼ 102 cm): ¼ 10d 1 PCDOM a

Z pCDOM ðlÞdl a

(3.40)

l

NO 2

Pa

¼ 10d 1

Z

NO 2

pa

ðlÞdl

(3.41)

ðlÞdl

(3.42)

l

NO 3

Pa

¼ 10d 1

Z

NO 3

pa l

Various studies have yielded useful correlation between the formation rate of
OH by the photoactive species and the respective absorbed photon fluxes of sunlight. In particular, it has been found that [34,78]: ¼ ð3:0  0:4Þ 105 PCDOM RCDOM
OH a Z    NO NO NO R OH2 ¼ F OH2 ðlÞpa 2 ðlÞdl





(3.43) (3.44)

l NO

R OH3 ¼ ð4:3  0:2Þ 102


½IC þ 0:0075 NO Pa 3 2:25½IC þ 0:0075

(3.45)

where [IC] ¼ [H2CO3] þ [HCO3] þ [CO32] is the total amount of inorganic carbon. The NO wavelength-dependent data of F OH2 ðlÞ are reported in Table 3.1.


110 Transformation Products of Emerging Contaminants in the Environment Table 3.1 Values of the quantum yield of
OH photoproduction by nitrite, for different wavelengths of environmental significance. NO

NO

NO

l, nm

F OH2 ðlÞ

l, nm

F OH2 ðlÞ

l, nm

F OH2 ðlÞ

292.5 295.0 297.5 300.0 302.5 305.0 307.5 310.0 312.5

0.0680 0.0680 0.0680 0.0678 0.0674 0.0668 0.066 0.065 0.063

315.0 317.5 320.0 322.5 325.0 327.5 330.0 333.3 340.0

0.061 0.058 0.054 0.051 0.047 0.043 0.038 0.031 0.026

350 360 370 380 390 400 410 420 430

0.025 0.025 0.025 0.025 0.025 0.025 0.025 0.025 0.025










At the present state of knowledge it is reasonable to hypothesise that CDOM, nitrite and nitrate generate
OH independently, with no mutual interaction. Therefore, the total formation rate of
OH (Rtot OH ) is the sum of the contributions of the three species:


NO

NO







CDOM þ R OH2 þ R OH3 Rtot OH ¼ R OH





(3.46)

Accordingly, having as input data d, A1(l), NPOC, [NO3], [NO2] and p (l) (the latter referred to a 22 W m2 sunlight UV irradiance, see Figure 3.20), it is possible to model the
expected Rtot OH of the sample. The photogenerated OH radicals could react either with the pollutant P or with the natural scavengers present in surface water (mainly organic matter, bicarbonate, carbonate and nitrite). The natural scavengers have an
OH scavenging rate constant:


X

k ½S  i Si i i

¼ 5  104 NPOC þ 8:5  106 ½HCO3   þ 3:9  108 ½CO3 2  þ 1:0  1010 ½NO2   (3.47)

(SikSi [Si] has units of s1; NPOC ¼ non-purgeable organic carbon is a measure of DOC, expressed in mg C L1, and the other concentration values are in molarity). Accordingly, the reaction rate between the pollutant P and
OH can be expressed as follows:


RPOH ¼ Rtot OH


kP; OH ½P P kP; OH ½P þ i kSi ½Si 


(3.48)




where kP, OH is the second-order reaction rate constant between P and
OH, and [P] is a molar concentration. Note that, in the vast majority of environmental cases kP, OH [P] SikSi [Si] would hold, thus the kP, OH [P] term can be neglected in the denominator of Equation 3.48. The pseudo-first order degradation rate constant of P is kP ¼ RPOH ½P1 , and the half-life time is tP ¼ ln 2 k1 P . The time tP is expressed in seconds of continuous irradiation under sunlight, at constant 22 W m2 UV irradiance. It has been shown that the sunlight energy reaching the ground on a summer sunny day (SSD) such as 15 July at 45 N latitude corresponds to 10 h ¼ 3.6 104 s of continuous irradiation at 22 W m2 UV irradiance [74]. Accordingly the











Phototransformation Processes of Emerging Contaminants in Surface Water 111

half-life time of P, because of reaction with
OH, would be expressed as follows in SSD units: P P ln 2 i kSi ½Si  5 i kSi ½Si  ¼ ¼ 1:9 10 (3.49) tSSD P; OH tot R 3:6 104 Rtot k OH kP; OH OH P; OH














It is 1.9 105 ¼ ln 2 (3.6 104)1. The steady-state [
OH] under 22 W m2 UV irradiance would be: Rtot (3.50) ½
OH ¼ P OH i kSi Si 


3.4.3 Direct Photolysis [79,80] The calculation of the photon flux absorbed by P requires taking into account the mutual competition for sunlight irradiance between P and the other surface water components (mostly CDOM, which is the main sunlight absorber in the spectral region of interest, around 300–500 nm). Under the Lambert–Beer approximation, at a given wavelength l, the ratio of the photon flux densities absorbed by two different species is equal to the ratio of the respective absorbances [76]. Accordingly, the photon flux absorbed by P in a water column of depth d (expressed in m) can be obtained by the following equations (note that A1(l) is the specific absorbance of the surface water sample over a 1 cm optical path length, Atot(l) the total absorbance of the water column, p (l) the spectrum of sunlight, referred to a UV irradiance of 22 W m2 as per Figure 3.20, eP(l) the molar absorption coefficient of P, in units of M1 cm1, and pPa ðlÞ its absorbed spectral photon flux density; also pPa ðlÞ ptot a ðlÞ and AP(l) Atot(l) hold in the very vast majority of environmental cases): Atot ðlÞ ¼ 100 A1 ðlÞ d

(3.51)

AP ðlÞ ¼ 100 eP ðlÞ d ½P

(3.52)

 Atot ðlÞ ptot Þ a ðlÞ ¼ p ðlÞ ð1  10

(3.53)

1 pPa ðlÞ ¼ ptot a ðlÞ AP ðlÞ ½Atot ðlÞ

(3.54)

The absorbed photon flux PPa is the integral over wavelength of the absorbed photon flux density: Z P (3.55) Pa ¼ pPa ðlÞdl l

The sunlight spectrum p (l) is referred to a unit surface area (units of E s1 nm1 cm2, Figure 3.20), thus PPa (units of E s1 cm2) represents the photon flux absorbed by P inside a cylinder of unit area (1 cm2) and depth 100 d (d is expressed in metres, thus 100 d is in cm). The rate of photolysis of P, expressed in M s1, can be expressed as follows (note that 1 L ¼ 103 cm3 and 1 m ¼ 102 cm): Z (3.56) RateP ¼ 10d 1 FP ðlÞpPa ðlÞdl l

where FP(l) is the photolysis quantum yield of P in the relevant wavelength interval (also note that 1 L ¼ 103 cm3). The pseudo-first order degradation rate constant of P is

112 Transformation Products of Emerging Contaminants in the Environment

kP ¼ RateP ½P1 , which corresponds to a half-life time tP ¼ ln 2ðkP Þ1 . The time tP is expressed in seconds of continuous irradiation under sunlight, at 22 W m2 UV irradiance. The sunlight energy reaching the ground on a SSD, such as 15 July at 45 N latitude, corresponds to 10 h ¼ 3.6  104 s continuous irradiation at 22 W m2 UV irradiance [74]. Accordingly, the half-life time expressed in SSD units would be given by: tSSD ¼ ð3:6  104 Þ1 ln 2ðkP Þ1 ¼ 1:9  105 ½Pd103 ðFP PNCP Þ1 ¼ 1:9  105 ½NCPd103 a P Z Z 1 1 1 5 3 ðFNCP pNCP ðlÞdlÞ ¼ 1:9  10 ½NCPd10 ðF ptot NCP a a ðlÞ ANCP ðlÞ ½Atot ðlÞ dlÞ l

¼ FNCP

l

1:9  108 d

R l

p ðlÞð1  10A1 ðlÞd Þ

eNCP ðlÞ dl A1 ðlÞ

(3.57)

Note that 1.9 108 ¼ 103 (ln 2) (3.6 104)1. 3.4.4 Reaction with CO3
[81]

The radical CO3
can be produced upon oxidation of carbonate and bicarbonate by
OH, upon carbonate oxidation by 3 CDOM , and possibly also from irradiated Fe(III) oxide colloids and carbonate [82]. However, as far as the latter process is concerned, there is still insufficient knowledge about the Fe speciation in surface waters to enable a proper modelling. The main sink of the carbonate radical in surface waters is the reaction with DOM. OH þ CO3 2 ! OH þ CO3 


½k58 ¼ 3:9  108 M1 s1 

(3.58)

OH þ HCO3  ! H2 O þ CO3 


½k59 ¼ 8:5  106 M1 s1 

(3.59)

CDOM  þ CO3 2 ! CDOM
þ CO3 
½k60 1  105 M1 s1 

(3.60)

½k61 102 ðmg CÞ1 s1 

(3.61)







3

DOM þ CO3 
! DOMþ
þ CO3 2


The formation rate of CO3
in reactions (3.58) and (3.59) is given by the formation rate of OH times the fraction of
OH that reacts with carbonate and bicarbonate, as follows:


OH tot RCO  ¼ R OH 3





8 2 8:5 106 ½HCO 3  þ 3:9 10 ½CO3  6  8 2 5 104 NPOC þ 1:0 1010 ½NO 2  þ 8:5 10 ½HCO3  þ 3:9 10 ½CO3  (3.62)

The formation of CO3
in reaction (3.60) is given by: CDOM RCDOM ¼ 6:5 103 ½CO2 CO 3  Pa


3

(3.63)

OH CDOM ¼ RCO . The transformation rate of P The total formation rate of CO3
is Rtot  þ R CO CO 3 3 3 

by CO3 is given by the fraction of CO3 that reacts with P, in competition with











reaction (3.61) between CO3
and DOM: RP;CO3 ¼


 Rtot CO k P;CO3 ½P





3

k61 NPOC þ kP;CO3 ½P


(3.64)

Phototransformation Processes of Emerging Contaminants in Surface Water 113

where kP;CO3 is the second-order reaction rate constant between P and CO3
. In the very [P] k61 NPOC. vast majority of environmental cases kP;CO
3 In a pseudo-first order approximation, the rate constant of P transformation is kP ¼ RP;CO3 ½P1 and the half-life time is tP ¼ ln 2 k1 P . Considering the usual conversion ( 10 h) between a constant 22 W m2 sunlight UV irradiance and a SSD unit, the following is obtained: expression for tSSD NCP;CO 3 ! k61 NPOC SSD 5 (3.65) tP;CO ¼ 1:9 10 3  Rtot CO k P;CO3

















3

Note that 1.9 105 ¼ ln 2 (3.6 104)1. 3.4.5 Reaction with 1 O2 [83] The formation of singlet oxygen in surface waters takes place upon energy transfer between ground-state molecular oxygen and the excited triplet states of CDOM (3 CDOM  ). Accordingly, irradiated CDOM is practically the only source of 1 O2 in aquatic systems. In contrast, the main 1 O2 sink is the energy loss to ground-state O2 by collision with water molecules, with a pseudo-first order rate constant k1 O2 ¼ 2:5  105 s1 . Dissolved species, including dissolved organic matter that is certainly able to react with 1 O2 , would play a minor role as sinks of 1 O2 in aquatic systems. The main processes involving 1 O2 and P in surface waters would be the following: 3

CDOM  þ O2 ! CDOM þ 1 O2

1

O2 þ H2 O ! O2 þ H2 O þ heat 1

O2 þ P ! Products

(3.66) (3.67) (3.68)

In the Rh^ one delta waters it has been found that the formation rate of 1 O2 by CDOM is CDOM ¼ 1:25 103 PCDOM [84]. Considering the competition between the deactivation of R1 a 1 O2 O2 by collision with the solvent (reaction (3.67)) and reaction (3.68) with P, one gets the following expression for the degradation rate of P by 1 O2 (note that kP;1 O2 ½P k1 O2 ): 1

O2

RP 1

¼ RCDOM 1 O2

kP;1 O2 ½P k1 O2

(3.69)

In a pseudo-first order approximation, the rate constant of P transformation is kP ¼

O RP 2 ½P1

and the half-life time is tP ¼ ln 2 k1 P . Considering the usual conversion ( 10 h) between a constant 22 W m2 sunlight UV irradiance and a SSD unit, the following expresis obtained (remembering that RCDOM ¼ 1:25 103 PCDOM and that sion for tSSD 1 a O2 P;1 O2 Z ¼ 103 d 1 pCDOM ðlÞdl): PCDOM a a l

¼ tSSD P;1 O2

4:81 ¼ RCDOM kP;1 O2 1 O 2

kP;1 O2

3:85 d Z pCDOM ðlÞdl a l

Note that 3.85 ¼ (ln 2)k1 o2 (1.25 103 3.60 104 103)1.

(3.70)

114 Transformation Products of Emerging Contaminants in the Environment

3.4.6 Reaction with 3 CDOM  [83] The formation of CDOM excited triplet states (3 CDOM  ) in surface waters is a direct consequence of radiation absorption by CDOM itself. In aerated solution, 3 CDOM  could undergo thermal deactivation or reaction with O2, and a pseudo-first order quenching rate constant k3 CDOM  5 105 s1 has been observed. The quenching of 3 CDOM  would be in competition with reaction between 3 CDOM  and P:

3

CDOM þ hn ! 3 CDOM 

(3.71)

CDOM  ðO2 Þ ! Deactivation and 1 O2 production

(3.72)

3

CDOM  þ P ! Products

(3.73)

In the Rh^ one delta waters it has been found that the formation rate of 3 CDOM  is [84]. Considering the competition between reaction (3.73) R3 CDOM ¼ 1:28 103 PCDOM a with P and other processes (reaction (3.72)), the following expression for the degradation rate of P by 3 CDOM  is obtained (note that kP;3 CDOM ½P k3 CDOM, where kP;3 CDOM is the second-order reaction rate constant between P and 3 CDOM  ): 3 CDOM

RP

¼ R3 CDOM 

kP;3 CDOM  ½P

(3.74)

k3 CDOM

In a pseudo-first order approximation, the rate constant of P transformation is kP ¼  3 RPCDOM ½P1 and the half-life time is tP ¼ ln 2 k1 P . Considering the usual conversion ( 10 h) between a constant 22 W m2 sunlight UV irradiance and a SSD Z unit, one gets the

(remembering that PCDOM ¼ 103 d 1 following expression for tSSD a P;3 CDOM  tSSD P;3 CDOM 

7:52 d R ¼ kP;3 CDOM  l pCDOM ðlÞdl a

pCDOM ðlÞdl): a l

(3.75)

Note that 7:52 ¼ ðln 2Þk3 CDOM  ð1:28 103 3:60 104 103 Þ1 . 3.4.7 Photochemical Transformation of Organic Pollutants The model described so far can be used to predict the environmental persistence of dissolved molecules. In recent years, surface-water pollution by pharmaceuticals has become a considerable environmental problem, which accounts for the importance of predicting the persistence and fate of these compounds. Table 3.2 reports the quantum yields and rate constant values that have been determined for carbamazepine (CBZ, antiepileptic drug) and ibuprofen (IBP, analgesic) for the main photochemical processes that are active in surface waters. Note that carbamazepine would mainly be degraded upon direct photolysis and reaction with
OH, while ibuprofen would react upon direct photolysis as well as with
OH and 3 CDOM  . Reactions with 1 O2 and CO3
would be insignificant for both compounds [85,86]. Figures 3.21 and 3.22 report the modelled half-lives of CBZ and IBP as a function of water depth and chemical composition (NPOC, nitrite and carbonate).

Phototransformation Processes of Emerging Contaminants in Surface Water 115 Table 3.2 Parameters describing the photochemical reactivity of CBZ and IBP, toward processes that are relevant to surface waters. Carbamazepine FP (polychromatic, UVB) kP; OH ; M1 s1 kP;3 CDOM; M1 s1 kP;1 O2 ; M1 s1


4

(7.8  1.8) 10 (1.8  0.2) 1010 (7.0  0.2) 108 (1.9  0.1) 105

Ibuprofen 0.33  0.05 (1.0  0.3) 1010 (9.7  0.2) 109 (6.0  0.6) 104

Figure 3.21 Half-life of CBZ as a function of: (a) NPOC and depth, with constant 50 mM nitrate, 1 mM nitrite, 2 mM bicarbonate and 10 mM carbonate. (b) Carbonate and nitrite, with constant 5 m depth, 3 mg C/L NPOC, 50 mM nitrate and 2 mM bicarbonate.

116 Transformation Products of Emerging Contaminants in the Environment

Figure 3.22 Half-life of IBP as a function of: (a) NPOC and depth, with constant 50 mM nitrate, 1 mM nitrite, 2 mM bicarbonate and 10 mM carbonate. (b) Carbonate and nitrite, with constant 5 m depth, 3 mg C/L NPOC, 50 mM nitrate and 2 mM bicarbonate.

First, CBZ would be more persistent than IBP in surface waters. The half-life time of CBZ increases with increasing depth, NPOC and carbonate, and decreases with increasing nitrite. The depth effect is caused by sunlight irradiance that decreases with depth. Moreover, NPOC scavenges
OH and competes with CBZ for irradiance, thereby inhibiting both
OH reaction and direct photolysis. Finally, carbonate scavenges
OH, and nitrite produces it. As far as IBP is concerned, its half-life has a maximum as a function of NPOC, because reaction with
OH prevails at low NPOC and reaction with 3 CDOM  at high NPOC. The

Phototransformation Processes of Emerging Contaminants in Surface Water 117

increase in the half-life with increasing depth is due to the fact that the deeper layers of a water body are poorly illuminated and, therefore, do not constitute a good environment for photochemical reactions. The increase in the half-life with increasing carbonate and the decrease with nitrite is accounted for by the fact that CO32 is an
OH scavenger that inhibits the hydroxyl-related pathway of IBP transformation, with nitrite that is an
OH source then the opposite effect occurs. The model approach presented here has been able to successfully predict the photochemical degradation kinetics of IBP and CBZ observed in the epilimnion of Lake Greifensee (Switzerland) [85,86]. 3.4.8 Photo-Transformation of Intermediates It is also possible to model the formation rate constants and yields of intermediates. For instance, acridine (ACR) is a mutagenic compound that is formed from CBZ upon direct OH
photolysis (yield hPhot ACR ¼ 0:036, i.e. 3.6%) and OH reaction (yield hACR ¼ 0:031) [86]. In p the generic process p, CBZ could produce ACR with yield hACR , experimentally determined as the ratio between the initial formation rate of ACR and the initial transformation rate of CBZ [86]. The pseudo-first order rate constant of ACR formation in the generic process p is ðkpACR Þ0 ¼ hpACR kpCBZ . Therefore, the overall rate constant of ACR formation upon direct photolysis and
OH reaction of CBZ is:


Phot OH OH ðkACR Þ0 ¼ hPhot ACR kCBZ þ hACR kCBZ





(3.76)

One can also obtain the overall yield of ACR formation from CBZ (hACR ), as , where kCBZ is the overall rate constant of CBZ photochemical hACR ¼ ðkACR Þ0 ðkCBZ Þ1P transformation (kCBZ ¼ p kpCBZ ). Figures 3.23 and 3.24 report ðkACR Þ0 and hACR , respectively, as a function of depth and NPOC. It can be observed that ðkACR Þ0 decreases with both d and NPOC, because ACR is

Figure 3.23 First-order rate constant of ACR formation, as a function of NPOC and depth d, with constant 50 mM nitrate, 1 mM nitrite, 2 mM bicarbonate and 10 mM carbonate.

118 Transformation Products of Emerging Contaminants in the Environment

Figure 3.24 Yield of ACR formation from CBZ, as a function of NPOC and depth d, with constant 50 mM nitrate, 1 mM nitrite, 2 mM bicarbonate and 10 mM carbonate.

formed from CBZ upon photolysis and
OH reactions that are both inhibited at greater depth (due to reduced sunlight irradiance) and at high NPOC (because of competition for irradiance between CDOM and CBZ and of
OH scavenging by DOM, respectively). The yield hACR also decreases with d and NPOC, but more slowly than ðkACR Þ0 . This happens because direct photolysis and the
OH reaction are the main CBZ transformation processes over a wide range of d and NPOC conditions, although the respective rates decrease with increasing d and NPOC. Reaction between CBZ and 3 CDOM  , which does not yield ACR [86], plays a significant role only at elevated d and NPOC, where its effect on the decrease in hACR can be noticed.

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Phototransformation Processes of Emerging Contaminants in Surface Water 121 48. Canonica, S. and Freiburghaus, M. (2001) Electron-rich phenols for probing the photochemical reactivity of freshwaters. Environmental Science & Technology, 35, 690–695. 49. Maurino, V., Bedini, A., Borghesi, D. et al. (2011) Phenol transformation photosensitised by quinoid compounds. Physical Chemistry Chemical Physics, 13, 11213–11221. 50. Maddigapu, P.R., Bedini, A., Minero, C. et al. (2010) The pH-dependent photochemistry of anthraquinone-2-sulfonate. Photochemical & Photobiological Sciences, 9, 323–330. 51. Canonica, S. and Laubscher, H.U. (2008) Inhibitory effect of dissolved organic matter on triplet-induced oxidation of aquatic contaminants. Photochemical & Photobiological Sciences, 7, 547–551. 52. Wenk, J. and Canonica, S. (2012) Phenolic antioxidants inhibit the triplet-induced transformation of anilines and sulfonamide antibiotics in aqueous solution. Environmental Science & Technology, 46, 5455–5462. 53. Czaplicka, M. (2006) Photo-degradation of chlorophenols in the aqueous solution. Journal of Hazardous Materials, B134, 45–59. 54. Vione, D., Bagnus, D., Maurino, V. and Minero, C. (2010) Quantification of singlet oxygen and hydroxyl radicals upon UV irradiation of surface water. Environmental Chemistry Letters, 8, 193–198. 55. Boreen, A.L., Edhlund, B.L., Cotner, J.B. and McNeill, K. (2008) Indirect photodegradation of dissolved free amino acids: The contribution of singlet oxygen and their potential as probes for assessing DOM photoreactivity. Environmental Science & Technology, 42, 5492–5498. 56. Latch, D.E. and McNeill, K. (2006) Microheterogeneity of singlet oxygen distributions in irradiated humic acid solutions. Science, 311, 1743–1747. 57. Grandbois, M., Latch, D.E. and McNeill, K. (2008) Microheterogeneous concentrations of singlet oxygen in natural organic matter isolate solutions. Environmental Science & Technology, 42, 9184–9190. 58. Minella, M., Romeo, F., Vione, D. et al. (2011) Low to negligible photoactivity of lake-water matter in the size range from 0.1 to 5 mm. Chemosphere, 83, 1480–1485. 59. Minella, M., Merlo, M.P., Maurino, V. et al. (2013) Transformation of 2,4,6-trimethylphenol and furfuryl alcohol, photosensitised by Aldrich humic acids subject to different filtration procedures. Chemosphere, 90, 306–311. 60. Cavani, L., Halladja, S., Ter Halle, A. et al. (2009) Relationship between photosensitizing and emission properties of peat humic acid fractions obtained by tangential ultrafiltration. Environmental Science & Technology, 43, 4348–4354. 61. Trubetskoj, O.A., Trubetskaya, O.E. and Richard, C. (2009) Photochemical activity and fluorescence of electrophoretic fractions of aquatic humic matter. Water Resources, 36, 518–524. 62. Halladja, S., Ter Halle, A., Aguer, J.-P. et al. (2007) Inhibition of humic substances mediated photooxygenation of furfuryl alcohol by 2, 4, 6-trimethylphenol. Evidence for reactivity of the phenol with humic triplet excited states. Environmental Science & Technology, 41, 6066–6073. 63. Chiron, S., Barbati, S., De Meo, M. and Botta, A. (2007) In vitro synthesis of 1,N6-etheno-20 -deoxyadenosine and 1,N2-etheno-20 -deoxyguanosine by 2,4-dinitrophenol and 1,3-dinitropyrene in presence of a bacterial nitroreductase. Environmental Toxicology, 22, 222–227. 64. Heng, Z.C., Ong, T. and Nath, J. (1996) In vitro studies on the genotoxicity of 2,4-dichloro-6-nitrophenol ammonium (DCNPA) and its major metabolite. Mutation Research: Genetic Toxicology and Environmental, 368, 149–155. 65. Tognazzi, A., Dattilo, A.M., Bracchini, L. et al. (2012) Chemical characterization of a new estuarine pollutant (2,4-dichloro-6-nitrophenol) and assessment of the acute toxicity of its quinoid form for Artemia salina. International Journal of Environmental Analytical Chemistry, 92, 1679–1688. 66. Bedini, A., Maurino, V., Minero, C. and Vione, D. (2012) Theoretical and experimental evidence of the photonitration pathway of phenol and 4-chlorophenol: a mechanistic study of environmental significance. Photochemical & Photobiological Sciences, 11, 418–424. 67. Chiron, S., Minero, C. and Vione, D. (2007) Occurrence of 2,4-dichlorophenol and of 2,4dichloro-6-nitrophenol in the Rh^one river delta (Southern France). Environmental Science & Technology, 41, 3127–3133.

122 Transformation Products of Emerging Contaminants in the Environment 68. Mack, J. and Bolton, J.R. (1999) Photochemistry of nitrite and nitrate in aqueous solution: a review. Journal of Photochemistry and Photobiology A: Chemistry, 128, 1–13. 69. Vione, D., Maurino, V., Minero, C. and Pelizzetti, E. (2002) New processes in the environmental chemistry of nitrite: Nitration of phenol upon nitrite photoinduced oxidation. Environmental Science & Technology, 36, 669–676. 70. Cullen, J.T., Bergquist, B.A. and Moffett, J.W. (2006) Thermodynamic characterization of the partitioning of iron between soluble and colloidal species in the Atlantic Ocean. Marine Chemistry, 98, 295–303. 71. Maddigapu, P.R., Minero, C., Maurino, V. et al. (2010) Enhancement by anthraquinone-2-sulphonate of the photonitration of phenol by nitrite: implication for the photoproduction of nitrogen dioxide by coloured dissolved organic matter in surface waters. Chemosphere, 81, 1401–1406. 72. Chiron, S., Comoretto, L., Rinaldi, E. et al. (2009) Pesticide by-products in the Rh^ one delta (Southern France). The case of 4-chloro-2-methylphenol and of its nitroderivative. Chemosphere, 74, 599–604. 73. Maddigapu, P.R., Vione, D., Ravizzoli, B. et al. (2010) Laboratory and field evidence of the photonitration of 4-chlorophenol to 2-nitro-4-chlorophenol and of the associated bicarbonate effect. Environmental Science and Pollution Research, 17, 1063–1069. 74. Minero, C., Chiron, S., Falletti, G. et al. (2007) Photochemical processes involving nitrite in surface water samples. Aquatic Sciences, 69, 71–85. 75. Minero, C., Lauri, V., Maurino, V. et al. (2007) A model to predict the steady-state concentration of hydroxyl radicals in the surface layer of natural waters. Annali di Chimica (Rome), 97, 685–698. 76. Braslavsky, S.E. (2007) Glossary of terms used in photochemistry, 3rd edition. Pure and Applied Chemistry, 79, 293–465. 77. Frank, R. and Kl€opffer, W. (1988) Spectral solar photon irradiance in Central Europe and the adjacent North Sea. Chemosphere, 17, 985–994. 78. Vione, D., Khanra, S., Cucu Man, S. et al. (2009) Inhibition vs. enhancement of the nitrate-induced phototransformation of organic substrates by the OH scavengers bicarbonate and carbonate. Water Research, 43, 4718–4728. 79. Vione, D., Feitosa-Felizzola, J., Minero, C. and Chiron, S. (2009) Phototransformation of selected human-used macrolides in surface water: kinetics, model predictions and degradation pathways. Water Research, 43, 1959–1967. 80. Vione, D., Minella, M., Minero, C. et al. (2009) Photodegradation of nitrite in lake waters: role of dissolved organic matter. Environmental Chemistry, 6, 407–415. 81. Vione, D., Maurino, V., Minero, C. et al. (2009) Modelling the occurrence and reactivity of the carbonate radical in surface freshwater. Comptes Rendus Chimie, 12, 865–871. 82. Chiron, S., Barbati, S., Khanra, S. et al. (2009) Bicarbonate-enhanced transformation of phenol upon irradiation of hematite, nitrate, and nitrite. Photochemical & Photobiological Sciences, 8, 91–100. 83. Vione, D., Das, R., Rubertelli, F. et al. (2010) Modelling of indirect phototransformation reactions in surface waters, in Advances in Synthetic Chemistry (ed. B. Pignataro), Wiley-VCH, Weinheim, Germany, pp. 203–234. 84. Al-Housari, F., Vione, D., Chiron, S. and Barbati, S. (2010) Reactive photoinduced species in estuarine waters. Characterization of hydroxyl radical, singlet oxygen and dissolved organic matter triplet state in natural oxidation processes. Photochemical & Photobiological Sciences, 9, 78–86. 85. Vione, D., Maddigapu, P.R., De Laurentiis, E. et al. (2011) Modelling the photochemical fate of ibuprofen in surface waters. Water Research, 45, 6725–6736. 86. De Laurentiis, E., Chiron, S., Kouras-Hadef, S. et al. (2012) Photochemical fate of carbamazepine in surface freshwaters: laboratory measures and modeling. Environmental Science & Technology, 46, 8164–8173.

4 Transformation Products of Emerging Contaminants upon Reaction with Conventional Water Disinfection Oxidants Jos
e Benito Quintana, Rosario Rodil and Isaac Rodr
ıguez Department of Analytical Chemistry, Nutrition and Food Sciences, IIAA – Institute for Food Analysis and Research, University of Santiago de Compostela, Spain

4.1 Introduction Nowadays, it is recognized that many emerging contaminants (ECs) (pharmaceuticals, personal care products, hormones, polymer additives, etc.) are not completely degraded during conventional biological treatments applied in urban sewage treatment plants (STPs). Consequently, they are released and spread in surface and groundwater [1]. A possibility to improve the efficiency of STPs consists in implementing tertiary treatments. Most of them are based either on the use of UV radiation and/or chemical oxidants, such as chlorinecontaining species, with free chlorine (HClO/ClO
) as one of the cheapest and more extensively used reagents. Depending on a relatively complex combination of factors (chemical structure of precursor species, water pH, dissolved organic matter, temperature, etc.) ECs surviving biological treatments might react with free chlorine without undergoing complete mineralization, generating the so-called disinfection by-products (DBPs), or, more generally, transformation products (TPs) [2,3]. Some of these TPs display different physico-chemical properties to their precursors, leading to (1) different behavior in the aquatic environment, including lower or higher resistance to further degradation, and (2) the need for eco-toxicological risk assessment studies [4,5].

Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

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Given that tertiary treatments are still not applied in most STPs, and that drinking water catchments of downstream populations are affected by discharges of treated sewage from upstream communities [1], the use of free chlorine in drinking water treatment plants (DWTPs) might also lead to the generation of TPs from ECs able to reach tap water resources. In addition to the two above source points of TPs, many personal care products come into contact with chlorinated waters at private homes, during daily-life activities, such as bathing and showering, and at swimming-pools. Compounds released from water distribution pipes, for example, bisphenol A (BPA), are also in permanent contact with free chlorine added to finished tap water. Obviously, humans are directly exposed to TPs generated in the above scenarios; moreover, these species are re-introduced into the aquatic environment, together with the remaining percentage of unchanged precursors, through urban sewage. On the basis of the above considerations, it is not surprising that the reactivity of ECs with chlorine-based disinfectants is receiving growing attention from analytical chemists and environmentalists. In most cases, pilot studies, using ultrapure water are first carried out in order to assess the stability of precursor pollutants, evaluating those variables affecting their half-lives (t1/2) and identifying the generated TPs, as well as their relative stabilities towards further transformation reactions. Thereafter, the occurrence of reported reactions in more complex water samples is normally evaluated and, in some cases, further tests are also applied to estimate potential eco-toxicological risks associated with TPs. This chapter covers some relevant aspects related to the formation of TPs from the interaction of ECs with free chlorine, including a discussion of analytical approaches employed to follow their time-course, parameters affecting the kinetics of transformation reactions and the most typical reaction mechanisms for different classes of ECs. A brief overview of reactions occurring between ECs and other disinfection agents (chlorine dioxide, chloramines, permanganate and ferrate) is also given.

4.2 Analytical Methodology for Transformation Products Identification Analytical approaches of interest to investigate the reactivity of ECs in chlorinated water samples must be able (1) to follow the time-course of precursor compounds in matrices with different complexities, (2) to detect the formation of TPs and (3) to provide enough information to assign a chemical structure to these TPs. Assuming that a maximum chlorine level of 10 mg/L is normally set in these studies and, that experiments are usually performed under pseudo-first order conditions (20–30 fold molar excess of chlorine), a good sensitivity is also required to detect the formation of minor TPs. In practice, the combinations of gas chromatography (GC) and liquid chromatography (LC) with mass spectrometry (MS) are the most used techniques. However, distinction must be made between targeted and non-targeted studies. Targeted studies are usually focused on the formation of some classes of well-known (and sometimes regulated) DBPs, such as trihalomethanes or nitrosamines [6–10]. In such cases, structural information is not that relevant, as pure standards are commercially available, extraction (if needed) and determination are tuned towards those target DBPs, and a quantification and mass (molar) balance can be established. Hence, GC can frequently be employed, due to the volatile nature of trihalomethanes and most nitrosamines, and MS is not always necessary.

Transformation Products of Emerging Contaminants 125

On the other hand, non-targeted studies aim to detect and to identify the products formed. Hence, coupling with MS is mandatory. In both types of study, chromatograms for different reaction times and conditions are compared with those corresponding to zero time and to control experiments (buffered sample aliquots without chlorine addition); thus, chromatographic peaks displaying changing intensities are related to the chlorination processes. Major by-products can be easily identified by visual inspection of the chromatograms; however, automated search strategies, using dedicated software, are required to detect the formation of minor byproducts, particularly when total ionic current (TIC) chromatograms show a high complexity, due either to the composition of the water sample (e.g., sewage) and/or to the characteristics of the ionization source and the mass analyzer. TPs identification relies on the information obtained from scan MS spectra with further confirmation against pure standards when available. If this latter verification is not possible, the use of at least two different ionization techniques, for example, electron ionization (EI) and electrospray ionization (ESI), brings a dramatic improvement in the reliability of identifications derived from MS spectra interpretation [11]. This comment is also valid when MS and MS/MS are used for structure assignments and/or when high resolution (HR) mass analyzers are employed. Additionally to GC-MS and LC-MS, some authors have also reported the use of nuclear magnetic resonance (NMR) to further characterize the structure of TPs. This technique is particularly useful to discriminate between positional isomers [11,12] that display identical MS spectra and whose identity cannot be established using retention times, since they are not separated by the chromatographic column or, simply, because pure standards are not available. On the other hand, NMR requires a relatively large amount of purified substance the isolation of which is often tedious. Whatever the methodology considered for evaluating the time-course of chlorination reactions, an accurate control of reaction times requires quenching the excess of chlorine with a reductor, such as ascorbic acid, sodium thiosulfate or sodium sulfite, before chromatographic analysis of the reaction mixture. Although a systematic comparison of the performance of different reductors has not been presented in the literature, the risk of secondary reactions between the reductor and labile TPs has been reported or, at least, suggested. In this sense, it is known that N-chloramine-type TPs, generated from emerging pollutants with primary and secondary amine groups, are back converted to the parent species in the presence of sodium thiosulfate [13]. Benzoquinone derivatives are also reduced to phenols in the presence of sodium thiosulfate or sulfite [14]; whereas they appear to be stable using ascorbic acid as chlorine quencher [15]. 4.2.1 GC-MS-Based Approaches When the precursor compound is amenable to GC separation, GC-MS constitutes the most popular tool to investigate its reactivity with free chlorine. Advantages of this technique are excellent separation capability, availability of EI-MS libraries, moderate cost and widespread availability in analytical chemistry and environmental laboratories. On the other hand, a relevant drawback of GC versus LC is that analytes (precursor compounds and their TPs) must be first extracted from the aqueous media and then transferred to an organic solvent suitable for injection in the GC-MS system. Thus, following the time-course of chlorination reactions becomes a time-consuming and labor-intensive task. Furthermore, only those TPs effectively

126 Transformation Products of Emerging Contaminants in the Environment

concentrated during the extraction step and compatible with GC determination can be identified. Table 4.1 summarizes sample preparation conditions considered during chlorination studies of ECs by GC-MS. Solid-phase extraction (SPE), normally based on the use of non-selective hydrophilic-lipophilic balanced sorbents (such as the OASIS HLB copolymer), is one the most common extraction approaches because of its capability to concentrate compounds within a broad range of polarities. Assuming that quantitative recoveries are attained for the precursor compound and its TPs, extraction of moderate sample volumes (about 100 mL) allows the use of chlorine and analyte concentrations close to levels existing in real-life water samples (few mg/L and mg/L, respectively), maintaining pseudo-first order conditions and ensuring the detection of even those TPs formed in minor amounts. Independently of the determination technique, an effective concentration step is required to further investigate the existence of TPs, previously detected under model conditions, in real-life samples, such as swimming-pool, tap or sewage water. In this vein, Canosa et al. [16] have employed the combination of OASIS HLB cartridges and GC-MS to investigate the reactivity of parabens with free chlorine, first considering ultrapure water fortified with these bactericides and chlorine, then mixing tap water with paraben-containing personal care products and finally processing real wastewater samples. Following this scheme, they obtained the first evidence of the existence of chlorinated forms of methyl and propyl parabens in urban sewage water, a fact which was further confirmed with pure standards and HR-MS [17]. A similar sample preparation strategy has been considered in GC-MS chlorination studies dealing with medium and low polarity UV filters, such as benzophenone-3 (BP-3) and 2-ethylhexyl-pamino benzoate (EHPABA) [18,19], lipid regulators [11] and bactericides [7,20]. Most of the analytes involved in the above studies contain polar functionalities, such as phenolic or carboxylic groups; thus, to improve their detectability, different derivatization reactions, based on the use of silylation or pentafluorobenzylation agents, have been introduced in the sample preparation scheme, Table 4.1. Alternatively to SPE sorbents, liquid-liquid extraction (LLE) has been used as a sample preparation technique for studying the chlorination of emerging pollutants by GC-MS. Yamamoto and Yasuhara [21] reported the LLE of chlorinated water samples (400 mL) with dichloromethane (2  50 mL), followed by evaporation to dryness and GC-MS analysis, to investigate the reactivity of BPA with chlorine. Although this option provides similar enrichment factors to SPE, it involves an even more expensive and tedious strategy. Both drawbacks are overcome when sample and organic solvent volumes are downscaled below 1 mL, as proposed by Rodil et al. [15], Table 4.1. Detection of non-targeted TPs arising during chlorination experiments requires operating the MS spectrometer in the scan mode. Thus, ion-trap systems are preferred to quadrupole instruments because of the lower detection limits of the former in this acquisition mode. The recent commercialization of accurate and HR- time-of-flight (TOF) MS analyzers, combined with GC, is expected to facilitate the identification of TPs displaying poor ionization efficiencies in LC-MS sources, particularly if HR-EI-MS libraries become available. On the other hand, analytical strategies considered when chlorination experiments are designed to quantify the formation of targeted TPs (e.g., trihalomethanes, N-nitrosamines and others hazardous DBPs) require the use of well-tuned sample preparation conditions to improve the detectability of these species. As an example, Rule and coworkers [7,10] have evaluated the formation of chloroform and bromodichloromethane from triclosan-containing personal care products by Purge and Trap (P&T) followed by GC with electron capture detection (ECD).

128 Transformation Products of Emerging Contaminants in the Environment

As regards N-nitrosamines, the chemical ionization mode is recommended to enhance their detectability by GC-MS [9]. Indeed, GC is well suited for targeted studies based on the volatility of most DBPs, lack of matrix effects unlike LC-MS, and poor or negligible ionization efficiency of some of those DBPs in LC-MS using ESI sources (e.g., trihalomethanes). 4.2.2 LC-MS-Based Approaches LC-MS is the technique of choice for most non-targeted studies because of its good sensitivity and the fact that the sample can be directly injected (after a simple filtration) into the system, minimizing the possibilities of the TPs being lost or degraded during the sample preparation process required by GC-MS [22]. The ionization interface most often used in these studies is ESI [23], based on its ability to produce ionized species for several analyte classes, particularly if (weak) acidic or basic groups are present in the original molecule or introduced during the oxidative reaction. Nevertheless, it must be kept in mind that some chemicals are hardly ionized by ESI and, sometimes, even by atmospheric-pressure chemical ionization (APCI). As, so far, the possibilities of alternative ionization sources, such as atmospheric-pressure photochemical ionization (APPI), remain unexplored [23], GC-MS must still be considered in those cases, as for example when studying butylated hydroxytoluene (2,6-di-tert-butyl-4methyl-phenol, BHT) reactions [15]. Moreover, some authors even decided to combine both GC-MS and LC-MS analysis in order to cover the maximum number of possible TPs [24,25]. The main weakness of LC-MS lies in the fact that a soft ionization is employed (normally ESI) which results mostly in a (de)protonated molecular ion being obtained, without forgetting the possibility of adduct formation. This produces little structural information and prevents spectral libraries being available, unlike with GC-EI-MS. On the other hand, as many TPs are not known and, hence, not present in GC-MS libraries, this feature becomes useful to ensure a correct identification of their empirical formulas, which is more difficult by GC-EI-MS. Then, once these formulas are known, MS/MS experiments are carried out in order to acquire further structural information. Hence, the most frequently employed analyzers are triple-quadrupoles (QQQ) and HR systems: TOF, QTOF and Orbitrap. Although, QQQ systems were used in the first studies in the literature, HR systems are nowadays widely available in research laboratories. Indeed, QQQ can be useful when the transformation pathway is simple and TPs structures are predictable. However, HR systems are required for better TPs identification. LC-HR-MS(/MS) instruments have improved both their resolution, being capable of producing full-width-at-half-maximum (FWHM) resolutions higher than 20 000, mass accuracy values below 5 ppm (equivalent in most cases to less than 2 mDa), good sensitivity, linear range and also, very importantly, a reliable isotopic pattern. Moreover, modern LC-HR-MS software includes not only mass accuracy coincidence with candidate formulas but an algorithm that considers the isotopic fit to the proposed formula. A few years ago the mass accuracy was the only parameter considered and the analyst needed to visually inspect the spectra for isotopic distribution. It was relatively simple to identify halogenated products, however, a good isotopic fit (combined with mass accuracy) can nowadays be obtained with compounds that do not contain halogens, based, for instance, on the 1.11% abundance of 13 C, 0.36% of 15 N, or 0.76% of 33 S and 4.22% of 34 S. An example is presented in Figure 4.1, showing the good identification of a chlorinated and a hydroxylated derivative of propranolol [26]. In that case, the boxes represent the theoretical isotopic pattern and the overall score the combined mass error, isotopic and distance between isotopes fit [17,27,28] and, as shown, a score above 90% is obtained in both cases.

130 Transformation Products of Emerging Contaminants in the Environment

Once the molecular formula of a TP is known, then MS/MS experiments are performed at different collision energies in order to collect as much fragmentation information as possible [29]. However, interpretation of MS/MS spectra is not always an easy task, particularly with low resolution MS. Thus, again HR systems are preferred. TOF analyzers cannot provide MS/MS spectra, but in-source fragmentation, with the risk of background fragments being identified as part of the molecule. However, based on the popularization of QTOF and Orbitrap, those problems are now avoided. Hence, empirical formulas can be obtained for the fragments, based on the exact masses of MS/MS products and the precursor formula. Moreover, it is a good idea to obtain and interpret the MS/MS fragmentation pattern of the original chemical before proceeding to the TPs. In this way, the appearance or disappearance of typical fragments can be very helpful in structure elucidation. Figure 4.2 shows as an example of the MS/MS spectra of propranolol and chloro-propranolol [26] where, although the exact Propranolol C16H21NO2

x103

O

4 3 2

H2C

C3H6N+ (–0.3 mDa)

–H2O

50

+

OH

N H

C6H14NO+ (+0.32 mDa) 116.1067

56.0501 C3H8NO+ (–0.3 mDa) 74.0603

1 0

N H

O H

C11H9O+ (+0.99 mDa) 157.0638 157 0638

C6H12N+ (+0 07 mDa) (+0.07 98.0964 –H2O –C 3H6

100

C13H11O+ (+0.69 mDa) 183.0797

–C2H2

[M+H]+ 260.1645

–H2 O, –C3H9N

150 200 Counts vs. Mass-to-Charge (m/z)

250

Chloro-propranolol C16H20NO2Cl x103 2

H2C

+

O OH

1.5

1

C3H8NO+ + (–0.5 mDa) C6H12N C3H6N+ 74.0605 (–0.04 mDa) (–0.51 mDa) 98.0966 56.0500 –C3H6 –H2O

0.5

–H2O

N H

O H

C6H14NO+ (+0.21 mDa) 116.1070

[M H]+ [M+H] Cl

C13H10O+ (+1.07 mDa) C11H8OCl+ 182.0719 (–0.06 mDa) 191.0260

–C2H2

50

100

294.1249

C13H10OCl+ (+0.46 mDa) 217.0408

–Cl –

0

N H

150 200 Counts vs. Mass-to-Charge (m/z)

C13H15NO2Cl+ ((+1.06 mDa)) 252.0775 –H2O, –NH3

–C3H6

250

Figure 4.2 HR-MS/MS spectra of propranolol and chloro-propranolol [26], with some relevant fragments being highlighted.

Transformation Products of Emerging Contaminants 131

position of the halogen cannot be clearly determined, it was shown not to be attached to the 1-isopropylamino-2-propanol group, as its ion (m/z 116.1067) and further fragments were observed in both the original pharmaceutical and the chlorinated TP. Further help can be obtained from the double-bond equivalents (DBE) of the TP and its fragments, which can indicate, for example, the loss of double bonds and ring openings and, if possible, by acquisition of positive and negative ionization modes spectra. Structure elucidation may be a time-consuming and skills-requiring operation. Yet, it is also important to detect as many TPs as possible. As mentioned, if the initial pollutant concentration tested is quite high, it may be easy to detect the presence of new chromatographic peaks for treated samples. However, minor peaks may remain unnoticed, particularly with LC-MS, where the baseline is higher than in GC-MS, masking many sample components. Again, the solution to this has been simplified by the availability of software dedicated to finding those peaks by comparing chromatograms of samples from different conditions (normally untreated and treated at different times), originally developed for metabolomics [29]. Nowadays, most manufacturers have their own commercial packages, for example, Mass Profiler from Agilent, MetaboLynx from Water, ProfileAnalysis from Bruker, and so on. Also, free software have been developed by different research groups, as for instance XCMS, developed by the Scripps Center for Metabolomics (http://metlin.scripps.edu/xcms/ index.php) or MetAlign, from the RIKILT - Institute of Food Safety (http://www.wageningenur.nl/en/show/MetAlign-1.htm). All this software works in a very similar way following the steps shown below or something very similar: 

Extraction of features: extraction of peaks with a characteristic mass or masses. Normally hundreds of features are obtained for each chromatogram, due to artifacts, background and contaminant peaks, TPs of the quenching agent and the actual pursued TPs.  Peak alignment: small retention time shifts are corrected, so that chromatograms can be more easily compared.  Detection of relevant features: once aligned, peak intensities are compared within replicates of the same sample and samples at different reaction times. Then a statistical comparison can be performed, so that only those peaks that show a significant change over time are retained.  Generation of formulas: formulas are produced. Then, the analyst has a reduced number of putative TPs and the process continues by manual verification and further acquisition of their MS/MS spectra. A good example of the advantages of the use of metabolomics software is presented by Gervais et al. [30] who studied the reaction of 17-a-ethinylestradiol (EE2) with chlorine and compared the metabolomics approach to the classical one. As a result, six new TPs were discovered, which could not be easily detected by the classical approach.

4.3 Factors Influencing the Kinetics of Chlorination The stability of ECs is normally assessed using an excess of chlorine. Although in some studies levels of this oxidant close to, or even above, 100 mg/L have been employed [31,32], a value of 10 mg/L is normally established as the highest assayed concentration [21,33]. It is very unlikely that species remaining stable for several hours in ultrapure water solutions

132 Transformation Products of Emerging Contaminants in the Environment

containing such a chlorine level, would undergo significant transformations in any of the water matrices reported in the introduction to this chapter. As far as it has been reported, in excess of free chlorine, reactive ECs degraded following pseudo-first-order kinetics, providing that the sample temperature, pH and concentrations of other salts (e.g., halides) remain constant. In practice, most of the primary TPs compete with their precursors for the available chlorine. Therefore, using a moderate molar excess of free chlorine (about 20 times), determination coefficients of the natural logarithm plots of concentration (or related response) versus time worsen with increase in reaction time. Thus, degradation of precursor species is normally followed up to 2–3 times their t1/2, with longer periods considered just to investigate the stability of the primary TPs. For a given concentration of chlorine, analyte stabilities usually change with the pH of the water samples. As a general rule, hypochlorous acid (HClO), pKa 7.5, is more reactive than the hypochlorite anion (ClO
); thus, transformation reactions are favored when the acid– base equilibrium is shifted to the neutral form of the acid [34]. This statement explains why non-ionizable ECs, for example, the UV filters 2-ethylhexyl-p-methoxycinnamate (EHMC) and EHPABA [35], and phenazone-type analgesics [36] display minimum t1/2 values at pHs below 7.5. The above rule is also valid to predict the effect of chlorine on the reactivity of ECs existing as ionized species at the natural pH (5–9) of water samples, under the condition that the ionized functionality is not involved in the reaction with free chlorine. Some examples are the carboxylic acids gemfibrozil (pKa 4.75) and naproxen (pKa 4.89) whose stabilities increase steadily with the pH of the water sample [13]. In the case of phenol-type pollutants, phenolate species are much more reactive than neutral forms. Since their pKa values are normally higher than that of HClO, their minimum t1/2 values correspond to pHs above 7.5, ranging from 7.5–8.3 for BPA (pKa 9.6), triclosan (pKa 8.1), propyl-paraben (pKa 8.2) and salbutamol (pKa 10.3) [7,16,26,37] to 8–9 for acetaminophen (pKa 9.5) [13] and 9.6 for nonylphenol (NP) and steroidal hormones (pKa values 10.4– 10.7) [38]. The dependences of the apparent second-order rate constants on sample pH for triclosan, acetaminophen and naproxen are depicted in Figure 4.3. Plotted data were calculated considering the individual rate constants, and the constant rate for the acid catalyzed reaction (case of naproxen), reported by Deborde et al. [38] for pairs of species involved in the chlorination reaction (phenol/phenolate and HClO/ClO
). Amines also react with free chlorine at different rates depending on their degree of ionization; although, in this case, the most reactive species is the neutral form. For those ECs whose amino moieties are only protonated at very acidic pHs (e.g., diclofenac), the effect of pH on their reactivity with free chlorine (pH range 6–10) depends only on the degree of dissociation of HClO, thus minimum t1/2 values have been reported at pHs below 7.5 [32,33]. On the other hand, the amino moiety of amphetamine-like illicit drugs stays in the protonated form at pHs below 9–10; thus, acid–base equilibrium constants of precursor species and HClO, and second-order constant rates for their neutral and ionized forms need to be considered to describe the dependence of their t1/2 on water pH. However, in the only chlorination study dealing with these emerging pollutants, the effect of pH on their t1/2 is not discussed [39]. Finally, the effect of pH on the reactivity of more complex molecules, containing several ionizable moieties able to participate in the reaction with chlorine, such as trimethoprim [34], sulfonamides [40] and tetracyclines [41], is more difficult to predict. In this case, acid–base distribution diagrams and the relative reactivity of neutral and ionized forms of each functionality with HClO and ClO
have to be considered.

Transformation Products of Emerging Contaminants 133 CI

CH3

OH O

OH H3C

CI

CI Triclosan

O

O Naproxen

H N

CH3 O

HO

Acetaminophen Triclosan

Acetaminophen

Naproxen

log k (M

–1

–1

s )

2.8 1.8 0.8 –0.2 –1.2 –2.2 5.0

5.5

6.0

6.5

7.0

7.5

8.0

8.5

9.0

9.5

10.0

pH

Figure 4.3 Chemical structures and evolution of apparent second-order rate constants (k), as a function of water pH, for the phenolic species triclosan and acetaminophen and the anti-inflammatory drug naproxen [3]. Reprinted with permission from [3] Copyright (2008) Elsevier Ltd.

Additionally to chlorine concentration and sample pH, kinetics of chlorine-induced transformation reactions can be enhanced by the presence of bromide salts. Several authors have demonstrated that low bromide levels, close to concentrations reported in surface water from areas affected by marine infiltrations [42], speed up the reaction rates of several ECs with free chlorine, contributing to the generation of new TPs [15,16,19,26,33,43]. In excess of free chlorine, bromide traces are converted into HBrO which, depending on the predominant reaction mechanism, might compete with HClO for the ECs. Thus, when halogenation processes are the basic route controlling the degradation of ECs, bromide traces normally lead to a significant reduction in their half-lives. This trend is illustrated in Table 4.2, which presents the t1/2 values for several phenolic ECs. On the other hand, the effect of bromide on the rates of reactions with free chlorine is less significant when precursor compounds are mainly degraded through hydroxylation and oxidation reactions, as is the case for BHT [15] and indomethacine [33]. Free chlorine-mediated reactions are competitive processes. This means that in surface, bath and/or recreational waters and sewage, other anthropogenic and/or natural organic chemicals might also react with the available chlorine. Therefore, considering free chlorine concentrations existing in the above matrices (up to 2 mg/L), or those typically used in tertiary treatments (up to 10 mg/L), the only expected TPs are those arising from primary reactions of ECs which display t1/2 in the range of a few minutes in ultrapure water assays. As regards chlorinated tap, bath and recreational waters, it has been proved that highly

134 Transformation Products of Emerging Contaminants in the Environment Table 4.2 Effect of bromide on the half-lives (t1/2) of selected ECs in free chlorine containing ultrapure water. Compound

pH

Free chlorine (mg/L)

Bromide (mg/L)

t1/2 (min)

Ref.

BHA

7

5.5

7.2

0.3

BP-4

7.3

1.7

EHPABA

7.2

0.6

Diclofenac

7

1

Naproxen

7

1

Phenazone

7

1

Salbutamol

7

1

7.4 0.6 2.7 0.8 0.78 0.55 0.19 26.7 21.3 4.6 325 15.7 446 94.4 54.4 24.3 20.9 5.5

[15]

BP-3

0 0.1 0 0.01 0 0.01 0.1 0 0.001 0.01 0 0.1 0 0.1 0 0.1 0 0.1

[19] [43]

[19]

[33] [33] [36] [26]

BHA: 2-tert-butyl-4-methoxy-phenol; BP-3: benzophenone-3; BP-4: benzophenone-4; EHPABA: 2-ethyhexylp-dimethylamino benzoate.

reactive species, such as parabens [16], benzophenone-type UV filters [19,43], EHPABA [18] and triclosan [7] are partially converted into halogenated TPs when these samples come into contact with personal care products containing these compounds. Furthermore, the chlorinated forms of methyl- and propyl-paraben are stable to further transformation reactions and have been detected in raw samples of urban wastewater [16,17]. Vikesland and coworkers [10] have even demonstrated a significant formation of trihalomethanes when triclosan-containing soaps are in contact with chlorinated tap water. Other TPs arising from chlorination of ECs, and detected at DWTPs or in tap water, are amphetamine derivatives [39], chlorinated forms of salicylic acid found in tap and wastewater [33], antioxidants TPs [15] and a derivative of the antiepileptic drug oxcarbazepine [44]. As regards treated wastewaters, Nakamura et al. [45] confirmed the presence of significant concentrations (up to 14 ng/L) of monochlorinated estrone in the outlet stream of STPs using sodium hypochlorite for final disinfection of treated sewage. Temperature is another key variable affecting the kinetics of any reaction. Thus, the relevance of transformation reactions in STPs and/or DWTPs, using free chlorine as disinfection agent, might change from winter to summer. The same comment is also valid for TPs generated during bathing and showering, depending on the temperature of the water. In this vein, Canosa et al. [16] have demonstrated that degradation of butyl paraben, contained in a bath gel in contact with tap water (free chlorine content 0.82 mg/L), occurred 1.5 times faster at 38  C than at 15  C; furthermore, an increase in the formation of brominated derivatives at 38  C was also reported. The yield of chloroform from triclosan-containing personal care products in contact with tap water also rose significantly when the water temperature increased from 30 to 40  C [10].

Transformation Products of Emerging Contaminants 135

4.4 Overview of Typical Reaction Mechanisms During Free Chlorine Treatments Reaction pathways of organic compounds with free chlorine vary depending on the functionalities existing in their chemical structures. Electrophilic substitution of hydrogen by chlorine is one of the most common reactions for aromatic compounds. Additionally to those parameters already commented on in Section 4.3 (chlorine concentration, sample pH, bromide and temperature), the rate of this reaction is strongly affected by the moieties attached to the benzene ring. Donor groups, such as hydroxy and amino, increase the electronic densities on carbon atoms located in ortho- and para-positions [12], speeding up the rate of the substitution reaction at these positions. On the other hand, aromatic compounds without activating (donor) groups and those bonded to electron-withdrawing substituents (e.g., carboxylic groups), display negligible reaction rates in chlorinated water samples. An intermediate reactivity corresponds to species with alkyloxy-functionalities (aromatic ethers). The above rule is valid to predict the relative t1/2 values of the analgesic and anti-inflammatory drugs: acetaminophen (phenol) < naproxen (aromatic ether) < ibuprofen (carboxylic acid) with free chlorine at pH 7 [3,13]. Reaction rates of phenolic compounds with free chlorine are further modulated by additional substituents attached to the aromatic ring, again depending on their electron-donating or -withdrawing character. On the basis of this contribution, Deborde and von Gunten [3] have estimated the apparent reaction rate constants for several phenolic endocrine disrupters, obtaining a good agreement with empirical values reported in previous studies. An illustrative example of the effect of these substituents on the t1/2 of precursor compounds has been reported for the UV-filters BP-3 and benzophenone-1 (BP-1). The chemical structures of both compounds differ just in the replacement of a moderate activating methoxy group (BP-3) by a strong donor hydroxy one (BP-1) in the phenolic aromatic ring. At neutral pH (7.2), the t1/2 of BP-3 was 2.7 min using 0.3 mg/L of chlorine, whereas, a t1/2 value of 0.4 min was measured for BP-1 in the presence of 0.1 mg/L of free chlorine [19]. These values result in calculated apparent second-order constant rates differing by one order of magnitude (20.5  103 and 1.01  103 M
1 s
1 for BP-1 and BP-3, respectively). ECs with amino functionalities are also prone to electrophilic substitution reactions. For aromatic species (aniline-type compounds) this reaction leads to the replacement of hydrogen by chlorine in the aromatic ring, as reported for diclofenac [32,33] and EHPABA [19]. In addition, anilines and primary and secondary aliphatic amines might also undergo N-chlorination to render chloramines. However, these TPs are considered as unstable intermediates, which normally suffer further rearrangements, or are simply back converted to their precursors by reaction with the quenching reagent or any moderate reductor present in the medium. It is likely that this behavior is responsible for differences between the number and type of TPs found for certain amine compounds, such as EHPABA, without considering a reductor to remove the excess of chlorine [35] and using Na2S2O3 as quencher [19]. In the latter study only aromatic halogenated TPs were noticed; whereas, N-demethylation by-products were also reported in the former case. Likely, N-demethylation was preceded by formation of a chloramine intermediate. Another often reported reaction during aqueous chlorination of aromatic emerging pollutants is the formation of hydroxy derivatives. Probably, these TPs are the result of a nucleophilic attack of the hydroxy group, from HClO, on electron deficient carbon atoms in the

136 Transformation Products of Emerging Contaminants in the Environment

structure of the parent compounds. Thus, the likelihood of hydroxylation reactions, either on the parent compounds or their chlorinated primary derivatives, increases with the number of electron-withdrawing moieties attached to the aromatic ring or to unsaturated bonds. Among the list of compounds undergoing aromatic hydroxylation as the primary transformation step in chlorinated water are the b-blocker propranolol [26], the antioxidant BHT [15] and the anti-inflammatory drug indomethacine [33]. A third mechanism reported during reaction of emerging pollutants and free chlorine consists of oxidation of hydroxy, carbonyl and alkyl moieties. Bedner and MacCrehan [14] described that oxidation of the phenolic moiety to a benzoquinone competes with aromatic chlorination during the reaction of acetaminophen with free chlorine, rendering most concerning by-products in terms of toxicity, Figure 4.4. Benzoquinones are also produced after oxidation of hydroxylated compounds, generated as primary TPs during chlorination of butylated hydroxyanisole (2-tert-butyl-4-methoxy-phenol, BHA) and other BHT analogs [15]. Oxidation of the carbonyl moiety has also been reported when studying the reaction of the UV filter benzophenone-4 (BP-4) with chlorine. After an initial chlorination in the phenolic ring of the precursor UV filter, conversion of the carbonyl group into an ester, followed by a rearrangement of the molecule through a Baeyer–Villiger oxidation reaction was observed [43]. Sulfur-containing species are also susceptible to oxidation reactions in the presence of chlorine. One of the examples reported in the literature is the conversion of the sulfide group of the antacid drug cimetidine into the corresponding sulfoxide, which has been identified as the first transformation step during the rapid degradation of the precursor drug in chlorinated water samples [31].

H N HO Electrophilic aromatic halogenation

O

Oxidation

H N

Cl

N O

HO

O

Electrophilic aromatic halogenation

Oxidation H N

Cl

O O

HO

O

O

Cl

Figure 4.4 Illustration of competitive reaction routes of acetaminophen with free chlorine [14]. Reprinted with permission from [14] Copyright (2006) US Government.

Transformation Products of Emerging Contaminants 137

Thus, electrophilic substitutions, nucleophilic hydroxylations and oxidations are the main reactions responsible for primary TPs generated from ECs in chlorinated water samples. Secondary TPs can be formed through cleavage of bonds and/or coupling reactions between two TPs, or between a TP and the parent species. In these cases, the compounds formed share fewer similarities than primary TPs with their precursors; therefore, they might display completely different persistence, mobility and toxicity. A well-known example of a relatively stable and toxic by-product is 2,4-dichlorophenol which arises from cleavage of the ether bond existing in triclosan and/or its chlorinated TPs (Figure 4.5) [7,20]. Similarly, cleavage of C

C bonds to produce the toxic 2,4,6-trichlorophenol has been reported as a secondary process during chlorination of BPA [12,21]. One of the final steps of cleavage reactions is the formation of trihalomethanes (basically chloroform), as has been demonstrated for the bactericide triclosan (Figure 4.5) [10]. Additionally to degradation through hydrolysis processes, TPs generated from ECs might interact through coupling reactions, resulting in higher molecular weight oligomers, as reported for BPA [12] and suggested for acetaminophen [14].

OH

Cl O

Cl

Cl Electrophilic aromatic halogenation

Cleavage OH

Cl

OH

Cl O

O

+

Cl

Cl

Cl

Cl

Cl

Cl OH

Cl

Electrophilic aromatic halogenation

Cleavage

OH

+

HO

Cl OH

Cl

Cl n

Cleavage

n=0,1 or 2

Cl

O

Electrophilic aromatic halogenation

Cl

Cl Cl

Cl OH

Cleavage & Halogenation CHCl3

Cl

Cl

Cl

Figure 4.5 Reaction routes of triclosan with free chlorine [7,10,21].

138 Transformation Products of Emerging Contaminants in the Environment

4.5 Review of Current Knowledge of Emerging Pollutant Reactions with Free Chlorine This section reviews the literature on the TP studies for eight selected groups of ECs. Further studies exist with other contaminants which are sometimes also considered as belonging to the EC class (e.g., herbicides). However, they are not considered here in order to keep this chapter a reasonable length. Tables 4.3 and 4.4 compile a revision of second order rate constants (k) and the most relevant transformation mechanisms observed for pharmaceuticals and other emerging pollutants, respectively. Only those ECs for which not only kinetics but also TPs were investigated are included in the tables.

Table 4.3 Revision of main mechanisms and kinetics of free chlorine reaction with pharmaceuticals. Family b-blockers

Compound Metoprolol Atenolol

Propranolol

b-agonists

Salbutamol

Analgesics Acetaminophen and antipyretics Salicylic acid Naproxen Diclofenac

Indomethacine

Phenazone

Transformation Reactions N-chlorination N-chlorination followed by N-dealkylation Oxidation of the amide group Electrophilic aromatic halogenation Electrophilic aromatic halogenation Aromatic hydroxylation Electrophilic aromatic halogenation Electrophilic aromatic halogenation Oxidation to quinoidal products Electrophilic aromatic halogenation Electrophilic aromatic halogenation Electrophilic aromatic halogenation Decarboxylation Aromatic hydroxylation followed by decarboxylation Dehydrogenation Cleavage of C

N bond Halogenation of the pyrazolone ring Hydroxylation N-demethylation

k (M
1 s
1)a
2

Refs.

1.4  10 1.3  10
2
9.4  10
2

[13,47] [13,26,48]

1.9  10
1
6.3

[13,26]

3.9  101

[26]

1.5
2.8  101

[13,14]

2.4  10
2

[33]

1.8
2.5

[13,33,46]

2–2.5

[32,33]

5.8  101
2.0  102

[13,33]

9.0  102

[36]

Transformation Products of Emerging Contaminants 139 Table 4.3 (Continued) Family

Lipid lowering drugs Antibiotics

Others

a

Compound

Transformation Reactions

k (M
1 s
1)a

Refs.

Propyphenazone

Halogenation of the pyrazolone ring Hydroxylation N-demethylation Electrophilic aromatic halogenation

1.5  103

[36]

8.8  10
1

[11,13,50]

7.6  105

[53]

5.1  102

[53]

Gemfibrozil

Ciprofloxacin

N-chlorination Electrophilic aromatic halogenation Cleavage of piperazine C

N bonds Enrofloxacin Electrophilic aromatic halogenation Decarboxylation Sulfamethoxazole N-chlorination Electrophilic aromatic halogenation Cleavage of C

S bond Oligomerization Carbadox Hydroxylation Intramolecular nucleophilic attack Cleavage of C

N bond Trimethoprim Electrophilic aromatic halogenation Aromatic hydroxylation Tetracyclines Electrophilic aromatic halogenation Hydroxylation Cimetidine Oxidation to sulfoxide S-chlorination followed by cleavage of C

S bond Intramolecular nucleophilic attack Fluoxetine N-chlorination Phenothiazine Oxidation to dioxide Electrophilic aromatic halogenation N-chlorination Oxcarbazepine Halogenation on a-carbon to the carbonyl group Cleavage of C

C bond and oxidation Intramolecular nucleophilic attack

Apparent second-order rate constant at pH 7.

9.2  102
1.5  103 [40,52]

1.0  103
1.5  104 [51,54]

5.6  101
2.0  102 [34,51] 1.1  104
1.8  106 [41] >6.9  101

[31]

– –

[47] [56]

–

[44]

140 Transformation Products of Emerging Contaminants in the Environment Table 4.4 Review of main mechanisms and kinetics of chlorine reaction with ECs. Family

Compound

Transformation Reactions

k (M
1 s
1)a

Refs.

Androgenic steroidal compounds

Trenbolone

Electrophilic halogenation Hydroxylation Electrophilic halogenation Hydroxylation Electrophilic aromatic halogenation Oxidation Hydroxylation Electrophilic aromatic halogenation Oxidation Cleavage of C

C bonds Electrophilic aromatic halogenation Hydroxylation Dehydrogenation Oxidation Chlorination N-dealkylation Cleavage of C

N bond N-dealkylation Hydrolysis Oxidation from N-methyl to N-formyl Electrophilic aromatic halogenation Cleavage of C

C bond Electrophilic aromatic halogenation Cleavage of C

C bond Cleavage of dioxole moiety Electrophilic aromatic halogenation Cleavage of C

C bond Electrophilic aromatic halogenation Cleavage of C

C bond Oligomerization Electrophilic aromatic halogenation Cleavage of C

C bond Hydroxylation

1.9  101 b

[58]

2.1 – 1.6  103c

[57]

1.3  102

[38,45]

1.2  102

[38,59]

1.1– 1.4  102

[30,38, 60,61]

–

[63]

1.4  10
1

[62]

–

[39]

–

[39]

–

[39]

6.2  101

[12,21,37]

1.3  101

[38,64]

Estrogenic steroidal compounds

2ENE, 4ENE, 9ENE,16ENE, DHEA E1

E2

EE2

Substances of abuse

Nicotine

Cocaine

MDA

MDEA

MDMA

Bisphenol A and Nonylphenol

Bisphenol A

Nonylphenol

Transformation Products of Emerging Contaminants 141 Table 4.4 (Continued) Family

Compound

Transformation Reactions

k (M
1 s
1)a

Refs.

Bactericides

Triclosan

Electrophilic aromatic halogenation Cleavage of C

O bond Electrophilic aromatic halogenation Electrophilic aromatic halogenation N-demethylation Electrophilic aromatic halogenation Oxidation (esterification) Cleavage of C

C bond Electrophilic aromatic halogenation Oxidation (esterification) O-demethylation Oxidation to quinone Aromatic hydroxylation Aromatic hydroxylation Electrophilic aromatic halogenation Aromatic hydroxylation Oxidation to quinone Electrophilic halogenation Hydroxylation Electrophilic halogenation Cleavage of C

C bonds and Oxidation

4.7  102
9.3  102

[7,20]

6.0  101
1.2  102 4.4– 5.1  101

[16]

1.0  103
1.1  103 d

[19,43]

6.2  102 d

[43]

2.1  101

[15]

2.4 1.6  101 1.3  102

[15] [15]

9.2  101

[68–70,79]

1.2  103

[72–74]

Parabens UV filters

EHPABA

BP-3

BP-4

Antioxidants

BHA

BHT BHT-OH BHT-COOH Cyanotoxins

Microcystin-LR

Cylindrospermopsin

[19,35]

a

Apparent second-order rate constant at pH 7, except: b pH 6.5, c pH 8, and d pH 7.3. 2ENE: 2(5a)-androsten-17-one; 4ENE: 4-androsten-3b-ol-17-one; 9ENE: 9(11),(5a)-androsten-3b-ol-17-one; 16ENE: 16(5a)-androsten-3-one; DHEA: dehydroandrosterone; E1: estrone; E2: 17-b-estradiol; EE2: 17a-ethinylestradiol; MDA: 4-methylenedioxyamphetamine; MDEA: 3,4-methylenedioxyethamphetamine; MDMA: 3,4-methylenedioxymethamphetamine; EHPABA: 2-ethyhexyl-p-dimethylamino benzoate; BP-3: benzophenone-3; BP-4: benzophenone-4; BHA: 2-tert-butyl-4-methoxy-phenol; BHT: 2,6-di-tert-butyl-4methyl-phenol; BHT-OH: 2,6-di-tert-butyl-4-hydroxymethyl-phenol; BHT-COOH: 3,5-di-tert-butyl-4hydroxybenzoic acid.

4.5.1 Pharmaceuticals 4.5.1.1 b-Blockers Pinkston and Sedlak [13] have shown that b-blockers (atenolol, metoprolol, nadolol and propranolol) react with chlorine quite rapidly, with faster rates noticed for those compounds with meta- substituents (i.e., propranolol and nadolol) than for those with substituents in the

142 Transformation Products of Emerging Contaminants in the Environment

para- position (i.e., metoprolol and atenolol) (Table 4.3). The reaction of all four compounds exhibited a strong dependence on pH, with rates increasing below pH 7, and three of the compounds (atenolol, metoprolol and propranolol) showing also an increased reaction rate above pH 8 [13,46], associated with the presence of an a-hydrogen to the secondary amine (at the propyl group) in these three compounds [13]. On the other hand, in a deeper study of propranolol and atenolol performed by Quintana et al. [26] on three parameters influencing chlorination, that is, pH, chlorine dose and bromide concentration, using an experimental design methodology, the effect of pH was not statistically significant and only the chlorine concentration was a statistically significant factor, showing that the chlorination dosage dominates the kinetics of the reaction. Moreover, a statistically significant interaction between chlorine and bromide for atenolol indicated a faster reaction with HBrO, rapidly formed from bromide upon chlorination, but only if the chlorine dose is not much higher than the bromide levels [26]. Pinkston and Sedlak [13] proposed that the first step in the reaction between b-blockers and chlorine is the rapid formation of the N-chloro derivatives. These intermediates have been identified for metoprolol [47] and atenolol [48], but when a reducer is used to quench chlorine the N-chloro compounds are back converted to the parent compound [48]. The second step proposed is the decomposition of the chloramine, yielding the N-dealkylated product, which has been identified for atenolol [26,48]. For this drug, DellaGreca et al. [48] identified two other TPs from the secondary N-chloro intermediate, a formamide derivative and 4-hydroxyphenylacetamide. On the other hand, Quintana et al. [26] found an additional major route by oxidation of the amide group. This mechanism starts by hydrolysis of the amide group to yield the carboxylic acid followed by decarboxylation to produce the benzaldehyde TP or halogenation in the aromatic ring (chlorination or bromination, when bromide is present in the solution) [26]. In the case of propranolol, neither the chloramine nor the N-dealkylated products were detected. The reaction pathway proposed by Quintana et al. [26] for this pharmaceutical suggests the introduction of hydroxy and halogen groups (chlorine or bromide) in the naphthalene ring, yielding stable TPs under strong and long reaction conditions (up to 72 h). 4.5.1.2 b-Agonists Salbutamol, the most important b-agonist, reacts rapidly with chlorine, exhibiting a strong dependence on pH [26]. The role of the pH is the same as for other phenolic compounds, that is, faster kinetics at basic pH values. Chlorination of salbutamol produces mono- and dihalogenated TPs, the introduction of the halogen atoms was confirmed to take place in the aromatic ring and not in the amine side chain [26], but none of these products were stable at long reaction times (72 h). 4.5.1.3 Analgesics and Antipyretics The four most common classes of analgesic/antipyretic drugs are: salicylates (including aspirin), acetaminophen, non-steroidal anti-inflammatory drugs (NSAIDs) and phenazone (also known as antipyrine). The stability to chlorination of these different classes has been studied in the literature, including an aspirin metabolite (salicylic acid), acetaminophen, six NSAIDs (ibuprofen, ketoprofen, indomethacine, naproxen, diclofenac and fenoprofen) and two phenazone-type drugs (phenazone and propyphenazone) [13,14,32,33,36,46]. Among the

Transformation Products of Emerging Contaminants 143

compounds investigated, three NSAIDs, ibuprofen, ketoprofen and fenoprofen did not show any significant transformation [13,33]. Due to their different structures there are big differences in the k values for the remaining drugs, ranging from 0.024 M
1 s
1 for salicylic acid up to 1500 M
1 s
1 for propyphenazone (Table 4.3) [14,33,36]. The reaction rates of these compounds, except acetaminophen, with chlorine decrease with the pH [13,32,33,36,46], due to the higher reactivity of HClO with respect to ClO
[3]. In the case of acetaminophen, the effect of the pH is also conditioned by its phenolic character as described in Section 4.3 [13]. Moreover, bromide concentration was another statistically significant factor for naproxen and diclofenac [33,36]. The reaction of chlorine with substituted phenols, such as acetaminophen and salicylic acid, and aromatic ethers, such as indomethacine and naproxen, was modeled by Pinkston and Sedlak [13] as an attack of HOCl on the aromatic ring undergoing substitution of hydrogen by halogen. The formation of halogenated derivatives in the ring activated positions has been confirmed for acetaminophen (Figure 4.4) [14], salicylic acid, naproxen and even diclofenac [33], but the reaction of indomethacine with HClO does not lead to further halogenation but to oxidation TPs [33]. The major products of indomethacine were an indole ring hydroxylated and a decarboxylated and hydroxylated derivative [33]. Also, two other minor products likely correspond to deshydroindomethacine and 4-chlorobenzoic acid, probably produced by the final hydrolysis of the amide group [33]. For acetaminophen, another alternative pathway identified is the oxidation to quinoidal products, leading to the toxic compounds N-acetyl-p-benzoquinone imine and 1,4-benzoquinone (Figure 4.4) [14]. In the case of diclofenac, a decarboxylation reaction and hydroxylation leading to decarboxy-hydroxydiclofenac and its chlorinated variant was proposed [32,33]. For the last TP (C13H10NOCl3) two structures have been proposed [32,33], but hydroxylation in position 4 of the dichlorinated aromatic ring was confirmed by LC-ESI-QTOF-MS/MS [33]. The transformation path of phenazone-type drugs consisted mainly in halogenation of the pyrazolone ring, hydroxylation of the methyl group or the pyrazolone ring and demethylation of the N-methyl group [36]. 4.5.1.4 Lipid-Lowering Drugs So far, lipid-lowering drugs considered in chlorination studies belong to the fibrates family, which is a class of amphipathic carboxylic acids: bezafibrate, gemfibrozil and clofibric acid (a metabolite of clofibrate). Bezafibrate and clofibric acid were stable under strong chlorination conditions [33]. On the other hand, Glassmeyer and Shoemaker [49] indicated that gemfibrozil was chlorinated showing the formation of one single product. Alternatively, the incorporation of one, two or even three chlorine atoms under strong chlorination conditions into the aromatic region of gemfibrozil was demonstrated by Krko9sek et al. [11]. However, under conditions that more accurately represent those of a STP or DTWP, chlorination reactions yielded only p-monochlorinated gemfibrozil and p-monobrominated gemfibrozil TPs, in the presence of bromide salts. These halogenated TPs enhanced the antiandrogenicity of gemfibrozil in a model fish species (Japanese medaka) [50]. 4.5.1.5 Antibiotics Antibiotics comprise a wide array of chemicals that can be divided into different sub-groups, such as b-lactams, quinolones, tetracyclines, macrolides, sulfonamides and others. The

144 Transformation Products of Emerging Contaminants in the Environment

effect of chlorine has been considered for most of these different groups of antibiotics, including two b-lactams (amoxicillin and cephalexin), three quinolones (ciprofloxacin, enrofloxacin and flumequine), three tetracyclines (tetracycline, oxytetracycline, chlortetracycline), three macrolides (erythromycin, roxithromycin and tylosin), six sulfonamides (sulfamethoxazole, sulfadimethoxine, sulfamerazine, sulfamethazine, sulfamethizole and sulfathiazole), carbadox and trimethoprim [34,40,41,46,49,51–54]. All those compounds react rapidly with chlorine (Table 4.3), except flumequine, which exhibits no apparent reactivity [53]. Antibiotics are often complex molecules which may possess different functionalities within their structures. Because of the different functionalities within a single molecule, their reactivity is strongly pH-dependent. Generally, reactivity increases above pH 4, reaching a maximum between 6 and 8 and then decreases above pH 8 [34,40,41,46,52–54]. The reaction of chlorine with sulfamethoxazole, ciprofloxacin, enrofloxacin and carbadox leads to the attack of HClO on the most basic amine: the aniline group in the case of sulfamethoxazole, the piperazine of fluoroquinilones and the hydrazine of carbadox [52–54]. In the case of sulfamethoxazole (Figure 4.6) and ciprofloxacin the formation of the N-chloro intermediate was observed [52,53]. This unstable N-chloro intermediate rapidly decays by halogenation to form a ring-chlorinated TP, identified for sulfamethoxazole, ciprofloxacin and enrofloxacin [52,53], or by molecule breakdown at the sulfonamide group for sulfamethoxazole to form 3-amino-5-methylisoxazole and N-chloro-p-benzoquinone imine (Figure 4.6) [52] and of the piperazine group for ciprofloxacin [53]. Moreover, a dimeric O H N S

NH2

O

O N

N-chlorination O H N S

N H

O

O N

Cl Electrophilic aromatic halogenation

Cleavage Oligomerization NH2 O N

+

O

N

O H N S

Cl O N

NH2

O Cl

N O O

O

H N S O N

O

N

N

S N H

O

Figure 4.6 Sulfamethoxazole transformation pathway in the presence of free chlorine [52]. Reprinted with permission from [52] Copyight (2004) American Chemical Society.

Transformation Products of Emerging Contaminants 145

product of sulfamethoxazole, azosulfamethoxazole, was observed (Figure 4.6). However, in contrast to ciprofloxacin, all enrofloxacin identified products retained the piperazine moiety [53]. The proposed reactions of enrofloxacin suggest the displacement of the quinolone carboxyl group by a Cl atom, followed by a halogenation ortho to the amine group [53]. In the case of carbadox, the unstable N-chloro intermediate produces hydroxylation TPs, which gradually decay by cleavage of the C

N bond, and can also react by intramolecular attack at the imine C by the carbonyl group, leading to a side-chain ring closure (Figure 4.7) [54]. The reaction pathway of trimethoprim is strongly pH-dependent. Thus, at acidic pH, the primary reaction is via its 3,4,5-trimethoxybenzyl moiety and at neutral and alkaline pH the 2,4-diaminopyrimidinyl moiety [34]. This reactivity leads to halogenated and hydroxylated TPs [34]. In the case of tetracyclines, several chlorinated and hydroxylated products without any ring breakage were proposed; however, the mass spectral information obtained with a single Q LC-MS system was insufficient to identify the structures of these products [41]. 4.5.1.6 Other Classes of Pharmaceuticals Other classes of pharmaceuticals considered in the literature are cardiac agents, angiotensin agents, psychiatrics and antacid drugs [49,55]. Among them, the cardiac agents warfarin, furosemide and amlodipine, the angiotensin agents losartan and irbesartan, the psychiatric drugs oxazepan, pirimidone, sertraline, zolpidem, fluoxetine, phenotrizine and oxcarbazepine O +

H N N

N

O O

+

N

Intramolecular attack

O

O +

N

N

N

O

Hydroxylation

O N

+

O O

O

OH

+

N

OR

O

+

N

+

N

Cleavage

+

O

O

O

Cleavage

O O

O

+

N

N NH2

O

N N

N

H N N O

+

N

O Hydroxylation O

OH

+

N

O

N N

O

+

N

O

Figure 4.7 Reaction pathways of carbadox with free chlorine [54]. Reprinted with permission from [54] Copyright (2006) American Chemical society.

146 Transformation Products of Emerging Contaminants in the Environment

and the antacid drug cimetidine were shown to be degraded during chlorination treatments [44,47,49,55,56]. On the other hand, the angiotensin agent valsartan and the psychiatric drugs bromazepan, carbamazepine, chlorpromazine, phenytoin and venlafaxine survived chlorination [55]. The reaction of some of these compounds with chlorine was studied in detail. The proposed pathway for the antacid cimetidine begins with its fast oxidation to cimetidine sulfoxide, followed by S-chlorination and C

S bond cleavage generating 4-hydroxymethyl-5methyl-1H-imidazole, finally forming 4-chloro-5-methyl-1H-imidazole and the sulfinyl chloride, which may undergo an intramolecular nucleophilic attack of either secondary amine at the sulfur center to give either the b-sultone or the d-sultone [31]. For fluoxetine, only the N-chloramine product has been identified in the literature [47]. In the case of phenothiazine, the first step is oxidation to the sulfone, followed by chloro-substitution in either the amine or the aromatic rings [56]. The chlorination pathway of oxcarbazepine proceeded by mono- and di-chlorination at the a-carbon to the carbonyl group, followed by hydrolysis and formation of a new carbonyl group and, finally, ring enclosure by an intramolecular reaction to produce a quinazoline moiety [44]. 4.5.2 Androgenic and Estrogenic Steroidal Compounds The reactivity of chlorine with both androgenic and estrogenic steroidal compounds has been widely studied in the literature. The chemical stability of androgenic steroids was found to be highly dependent on the degree of saturation and the relative position of double bonds within their structures. Thus, conjugation between the ketonic moiety in position 3 and 1- or 4-carbon double bonds, for example, progesterone, testosterone, androstan-17b-ol-3-one (ANE), 1,4-androstadien-17b-ol-3-one (DIENE), 1,(5a)-androsten-17b-ol-3-one (1ENE) and 4-androsten-3,17-dione (4ENE-3,17N), inhibits reaction with free chlorine [57]. On the other hand, 2(5a)-androsten-17-one (2ENE), 4-androsten-3b-ol-17-one (4ENE), 9(11), (5a)-androsten-3b-ol-17-one (9ENE), 16(5a)-androsten-3-one (16ENE) and dehydroandrosterone (DHEA), which do not have the conjugation character are degraded during chlorination [57]. Finally, the extended conjugated structure of trenbolone imposes a somewhat unique response so that it readily reacts with chlorine [58]. A variety of oxidative TPs of androgenic compounds were observed to be produced through electrophilic substitutions by chorine and hydroxylations [57,58]. On the other hand, various authors have shown the rapid reaction of estrogenic steroids, viz. estrone (E1), 17-b-estradiol (E2), estriol (E3) and EE2, with free chlorine (Table 4.4) [38,45,59–61]. For these compounds, the reactivity increases above pH 5, reaching a maximum between 8 and 10 and then decreasing above pH 10 [38], similar pH-dependence profiles have been noted for other phenolic compounds previously. Moreover, the reaction rates of EE2 were strongly enhanced in the presence of bromide, which can be attributed to the higher reactivity of bromine than that of chlorine [60]. In the case of estrogenic steroids, the main pathway is an electrophilic attack of chlorine or bromide on ortho carbons in the phenolic ring [30,45,59–61]. For E1, in a second step, the dichlorinated TP undergoes an oxidative chlorination at position 10, leading to a trichlorinated product 2,4,10-trichloro-1,4-estradiene-3,17-dione, and a further substitution at the same position by a hydroxy group [45]. For EE2, eight hydroxylated and halogenated forms were detected, but some uncertainties remain concerning the exact positions of the hydroxy and halogen groups [30]. Finally, in the case of E2, besides three chlorinated

Transformation Products of Emerging Contaminants 147

products (2,4-dichloro-E2, monochloro-E1 and 2,4-dichloro-E1), four TPs formed by ring cleavage through 9–10 and 16–17 bonds were identified [59]. 4.5.3 Substances of Abuse The stability of nicotine, amphetamine-type-stimulants and cocaine has been investigated [39,62,63]. Under laboratory conditions, all these compounds react with chlorine. However, the study on nicotine was accomplished using huge amounts of chlorine and nicotine and, therefore, conclusions cannot be directly extrapolated to a real world situation [63]. Among the amphetamine-type-stimulants, amphetamine and methamphetamine showed the faster kinetics (t1/2 below 1 min), while 4-methylenedioxyamphetamine (MDA), 3,4-methylenedioxyethamphetamine (MDEA) and 3,4-methylenedioxymethamphetamine (MDMA) displayed slower reaction rates (t1/2 above 1 h) [39]. For cocaine half-lives of several hours were obtained, between 4 and 39 h (with a free chlorine concentration of 10 mg L
1) depending on the pH of the sample [62]. For amphetamine and methamphetamine, no TPs were identified. In the case of MDA and MDEA a common chlorinated compound (3-chlorobenzo)-1,3-dioxole was formed, probably as a result of an electrophilic aromatic substitution over the dioxole a-carbon followed by cleavage of a C

C bond through chloramine intermediates [39]. However, for MDMA the cleavage of the dioxole moiety occurs to give rise to 3-chlorocatechol [39]. For nicotine, all the products are derived through transformations of the pyrrolidine moiety by addition, substitution and oxidation reactions, leading to several TPs, two of them chlorinated [63]. In the case of cocaine, the main chlorination pathway is attack on the amine group yielding norcocaine by N-dealkylation, moreover hydrolysis produced demethylation of the ester group yielding benzoylecgonine (from cocaine) and norbenzoylecgonine (from norcocaine) [62]. Another minor pathway observed is an oxidative reaction of the N-methyl group to produce N-formylnorcocaine [62]. 4.5.4 Bisphenol A and Nonylphenol As many other phenolic compounds, BPA and NP rapidly react with free chlorine [12,21, 37,38,64]. In both cases, the mechanism of reaction proceeds through electrophilic attack on the phenoxide ion, at its ortho-position, and formation of chlorinated derivatives [12,21, 64,65]. This is followed by cleavage of the bond between the a-C on the isopropyl, or nonyl, moiety and the b0 -C on the benzene ring, to form trichlorophenol (for both BPA and NP) and 4-isopropyl-20 -hydroxyphenol (for BPA) [12,21,64,65]. Besides the aforesaid reactions, breakdown of the aliphatic chain of NP led to the formation of isoamil/propyl/ isobutyl/and isopentyl-20 -hydroxyphenol compounds [64]. In the case of BPA, the formation of polychlorinated phenoxyphenols, through coupling reactions, has also been shown [12]. 4.5.5 Bactericides: Triclosan and Parabens Triclosan and parabens are easily chlorinated with sodium hypochlorite following the typical pH dependence of phenols [7,16,20,66,67]. In the case of parabens, mono-and di-chlorinated compounds were identified for methyl, ethyl, propyl and butyl derivatives through electrophilic aromatic substitution at ortho-positions [16]. Conversely to the limited stability of monochlorinated paraben by-products, the di-chlorinated ones were shown to be rather

148 Transformation Products of Emerging Contaminants in the Environment

persistent [16]. Triclosan is also chlorinated to produce tetra- and penta-hydroxylated diphenyl ethers (normally referred to as tetra- and pentaclosan) [7,20,66,67]. The cleavage of the ether bridge contained in the structure of triclosan, tetraclosan and pentaclosan leads to the generation of 2,4-dichlorophenol [7,20,67] and resorcinols containing one, two and three chlorine atoms (Figure 4.5) [20]. Moreover, a trichlorophenol is reported as a triclosan TP [7,20,66,67]. Different structures were suggested for this trichlorophenol in the literature: 2,4,6-trichlorophenol (obtained by chlorination of 2,4-dichlorophenol) or 2,3,4-trichlorophenol (produced by cleavage of the ether link in the pentaclosan molecule) [7,20,67], the first option being the most accepted [7,20]. 4.5.6 UV Filters The reactivity of five UV filters, namely ethylhexyl salicylate (EHS), EHMC, EHPABA, BP-3 and BP-4, was studied in chlorinated water samples [19,35,43]. The stability of these UV filters increased in the following order: BP-3 < BP-4 < EHPABA < EHMC < EHS (Table 4.4) [19,35,43]. Actually, the extension of EHS degradation was estimated as negligible under a real world situation [19]. The reactivity of UV filters is also pH-dependent. EHPABA and EHMC are more stable at higher pH-values (pH 8–9) [19,35], while BP-3 and BP-4 are less stable at pH 8 than at pH 6–7 [19,43]. Moreover, addition of bromide to the sample reduces the stability of UV filters [19,43]. Regarding the reaction mechanism, for the UV filters EHPABA, BP-3, BP-4 and EHMC, the first step consisted of an aromatic substitution of one or two atoms of hydrogen by chlorine and/or bromide ortho to the amino moiety (EHPABA) and those ortho and para to the hydroxy group (BP-3 and BP-4) [19,35,43], Figure 4.8. In the case of EHPABA, N-demethylation products were also identified [35]. For the benzophenone-like UV filters, BP-3 and BP-4, the second step consisted of the conversion of the carbonyl group into an ester, which links the two aromatic rings, and then a further electrophilic substitution reaction might occur [43]. Additionally to the above mentioned species, in the case of BP-3, halogenated forms of 3-methoxyphenol were generated from cleavage of the carbonyl bond in the structure of BP-3 and its halogenated TPs, Figure 4.8 [19]. 4.5.7 Antioxidants Only one work has been published on the stability of antioxidants during chlorination [15]. The degradation of three antioxidants: BHT, BHA and 2-tert-butylbenzene1,4-diol (TBHQ, also a BHA metabolite); and four metabolites of BHT: 2,6-di-tertbutyl-4-hydroxymethyl-phenol (BHT-OH), 3,5-di-tert-butyl-4-hydroxybenzaldehyde (BHT-CHO), 3,5-di-tert-butyl-4-hydroxybenzoic acid (BHT-COOH) and 2,6-di-tertbutylcyclohexa-2,5-diene-1,4-dione (BHT-Q), was investigated [15]. Among the target compounds, only BHT-Q and BHT-CHO were stable during chlorination. Reaction kinetics were mainly pH-dependent for BHA and BHT, showing that an increase in the pH, within the considered range (5.7–8.3), led to faster kinetics. Moreover, the presence of bromide reduced the stability of BHA [15]. The transformation of these compounds has revealed two main routes: first, hydroxylation, possibly via an intermediate halogenated product observed in the case of BHT-OH and BHTCOOH and second, oxidation, leading to the quinone derivative (except for BHT). Finally, in the case of BHA a further hydroxylation reaction of the quinone derivative was observed [15].

Transformation Products of Emerging Contaminants 149 O

OH

O Electrophilic aromatic halogenation OH

OH

O

Cl

O

AND O

O Electrophilic aromatic halogenation

Cl

O

OH Cl Cleavage

O Cl

OH Oxidation

Cl

O

OH

O

O

Cl

Cl Electrophilic aromatic halogenation

O Cl OH Cl

Cl

Electrophilic aromatic halogenation

O

O

OH Cl

O

Cl O

Cl Cl

Figure 4.8 Benzophenone-3 transformation pathway in the presence of free chlorine [19,43]. Reprinted with permission from [19] and [43] Copyright (2007) and (2012) Elsevier Ltd.

4.5.8 Cyanotoxins Cyanotoxins can be classified into two groups, namely hepatotoxic cyclic peptides, including microcystins (MCs) and nodularins, and neuro- or cytotoxic alkaloids, including cylindrospermopsin, anatoxins and saxitoxins. Chlorine has been demonstrated to be an effective oxidant for MCs [68–71], cylindrospermopsin [72–75] and saxitoxins [76]. On the other hand,

150 Transformation Products of Emerging Contaminants in the Environment

anatoxin-a was shown to be very stable during chlorination at pH 7 [74]. Kinetic analysis of the oxidation of toxins in natural water revealed significant differences in their susceptibility to chlorine, with saxitoxins being the most reactive species, followed by cylindrospermopsin and MCs (Table 4.4) [77]. According to literature data, MCs and saxitoxins reaction with chlorine depends on the substituents [69,76], for example, the reactivity of microcystins follows the order: MC-YR > MC-RR > MC-LR > MC-LA [69]. Moreover, chlorination efficiency depends largely on experimental conditions. Hence, microcystins transformation depends strongly on the pH-value, decreasing while pH increases in the range 4–9 [68,71], simply explained by the speciation of chlorine. Senogles et al. [75] concluded that degradation of cylindrospermopsin is reduced at pH lower than 6. In contrast, other studies suggest that pH should be maintained equal to 7 because higher values would decrease the toxin transformation rate [74,78]. Finally, saxitoxins reaction rates simply rise when pH is increased from 4 to 9 [76]. Several reactive sites of chlorine attack can be expected depending on the considered MC: phenolic ring, ethylguanidyl group and adda group. Chlorination by-products of MC-LR have been studied and six chlorinated and hydroxylated compounds have been identified, as well as their respective isomers. The reaction pathway proposed includes the addition of chlorine atoms on the conjugated diene of adda followed by hydroxylation, yielding monochloro-MC, monochloro-hydroxy-MC and dihydroxy-MC [79]. These products further reacted with chlorine to form trichloro-hydroxy-MC, monochloro-dihydroxy-MC and dichloro-dihydroxy-MC [70]. Concerning cylindrospermopsin, the main chlorine reactivity was observed on the uracil moiety, yielding 5-chloro-cylindrospermopsin and a carboxylic acid derivative, cylindrospermic acid [72,74]. Moreover, a TP with the empirical formula C13H18N4O7S has been reported in the literature, but a structural formula has not been proposed [73].

4.6 Other Disinfection Agents Although free chlorine is the most commonly employed conventional disinfection agent, other chemicals are also used as a complement to chlorine, or have been suggested in order to reduce the amount of DBPs. Among them, ozone is, perhaps, the most relevant. Yet, as ozonation is discussed in detail in Chapter 10, it will not be considered here. Other common agents are also chlorinated species, such as chloramine and chlorine dioxide. Moreover, potassium permanganate is relatively frequently used as a peroxidation treatment. Recent research also points to the use of ferrate. Hence, these four chemicals will be considered in this section. A comparison of the redox potentials of the here-mentioned chemicals with those of chlorine [80,81] is presented in Table 4.5. 4.6.1 Chlorine Dioxide Chlorine dioxide is a more selective oxidant, as compared to free chlorine/hypochlorite, as a result of its lower redox potential (see Table 4.5). Its advantages over chlorine are a good inactivation of protozoa and reduction in the formation of regulated halogenated DBPs, such as trihalomethanes [82,83]. However, chlorine dioxide is not stable and needs to be generated in situ, by reaction of chlorite with chlorine, hence, if not properly controlled residual

Transformation Products of Emerging Contaminants 151 Table 4.5 Redox potentials of common water disinfection agents [80,81]. Oxidant

Reaction

E0 (V)

Chlorine Hypochlorite

Cl2 þ 2 e
$ 2 Cl
HClO þ Hþ þ 2 e
$ Cl
þ H2O ClO
þ H2O þ 2 e
$ Cl
þ 2 OH
ClO2 (aq) þ e
$ ClO2
NH2Cl þ H2O þ 2e
$ Cl
þ OH
þ NH3 MnO4
þ 4 Hþ þ 3 e
$ MnO2 þ 2 H2O MnO4
þ 8 Hþ þ 5 e
$ Mn2þ þ 4 H2O MnO4
þ 2 H2O þ 3 e
$ MnO2 þ 4 OH
FeO42
þ 8 Hþ þ 3 e
$ Fe3þ þ 4 H2O FeO42
þ 4 H2O þ 3 e
$ Fe(OH)3 þ 5 OH


1.36 1.48 0.84 0.95 0.74 1.68 1.51 0.59 2.20 0.70

Chlorine dioxide Chloramine Permanganate

Ferrate

chlorine/hypochlorite is also present in the medium [83]. Because of its lower oxidative potential, it cannot oxidize bromide, so brominated products are not expected. However, this lower redox potential does not imply slower reaction rates. Actually, the reaction rate is strongly dependent on the pollutant structure and the pH. Hence, some pharmaceuticals, such as EE2 or tetracycline and sulfonamide antibiotics, react much faster with chlorine dioxide than with free chlorine [41,84], whereas others, for example, atenolol or fluoroquinolone antibiotics react faster with the latter oxidant [84,85]. Unlike chlorine, mostly transformed to hypochlorite in water, chlorine dioxide remains as a dissolved gas. Hence its reaction potential is not affected by the pH. However, the reaction rate with ECs is influenced by the pH, providing that they contain ionizable substituents (e.g., phenols). Thus, reaction proceeds much faster with the electron-rich form (e.g., phenolate), as it typically occurs during chlorine disinfection. This fact has been demonstrated with several pharmaceutical classes, showing a faster reaction at higher pH values [41,82,85]. Most studies on chlorine dioxide reaction with emerging pollutants are dedicated to pharmaceuticals. A few publications discuss the oxidizability of a broad range of drugs with chlorine dioxide in real water matrixes, from groundwater to effluent wastewater [82,86,87]. As a short summary, the conclusions of those works are that estrogen hormones, several antibiotic classes and phenazone analgesics are easily transformed with ClO2. On the other hand, acidic drugs (NSAIDs and lipid regulators, with a few exceptions), carbamazepine, b-blockers/agonists and several basic antidepressant drugs are hardly removed under real conditions. However, further studies have shown that under strong dosages, carbamazepine [25] and citalopram [24] can be at least partially transformed. In both cases the reaction was investigated by a combination of LC-QTOF-MS/MS and GC-MS, permitting a good identification of the transformation products. In the case of carbamazepine, the reaction starts with an attack at the 10,11-double bond to yield 10,11-epoxycarbamazepine. Then, the reaction continues by elimination of HNCO to produce acridine and its hydroxylated and a-keto derivatives [25]. Citalopram also reacts by introduction of hydroxy and oxo groups and by dealkylation of the secondary amine [24]. Wang et al. [85] investigated the reaction of fluoroquinolone antibiotics with ClO2, confirming that attack at amines is one of the preferred reaction routes. Hence, reactions

152 Transformation Products of Emerging Contaminants in the Environment

of this class of antibiotics proceeded via the piperazine group, followed by ring opening, and the only fluoroquinolone not reactive was flumequine, which lacks the piperazine group in its structure. A series of structures and intermediates was proposed, but as they are based on single-quadrupole LC-MS, those TPs would need further confirmation. The same authors also studied tetracycline antibiotics, comparing free chlorine and chlorine dioxide [41]. Tetracyclines reacted with ClO2 by hydroxylation, introduction of oxo groups and elimination of the amino group, but no chlorinated TPs were observed, although the exact TP structures were not elucidated. In contrast, in the reaction with free chlorine, both groups of antibiotics yielded several TPs incorporating chlorine atoms in their structure by electrophilic aromatic substitution reactions [41,53]. However, a recent work from the same authors on trimethoprim [88] indicates that, besides attachment to the amine ring, introduction of chlorine into the benzene cycle is also possible with ClO2. Yet again the work was based on single LC-MS and the proposed structures should be taken with caution. Other studies, with carbamate pesticides and azo dyes show that chlorine dioxide can also attack and break azo groups and oxidize sulfides [89,90]. 4.6.2 Chloramination Inorganic chloramines are obtained by mixing chlorine with ammonia. Depending on the ratio of these two reactants and the water pH, different species, from monochloramine to trichloramine, can be obtained. However, under normal water pHs (6–9) mostly monochloramine is produced. Its redox potential is relatively low (Table 4.5) and, therefore, it is a poor disinfectant. However, it remains stable for long periods of time, so it is frequently used as a secondary oxidant in order to protect drinking water along the distribution system and to decrease the amount of trihalomethanes formed [83]. There are very few studies considering the reaction of chloramine with emerging (and other organic) pollutants, which can be partially understood in view of its low reactivity. Actually, a study of Chamberlain et al. [91] which compared the reaction of the pesticide friponil with free chlorine, chlorine dioxide, monochloramine and permanganate showed that reaction was quite fast with all the disinfection agents (t1/2 at pH 8.6 from 3 to 42 min) excepting chloramine, which did not react at all. However, the discovery of the reaction of some pesticides upon ozonation to yield important amounts of nitrosamines [8] has fostered the investigation of chloramination of ECs targeting this specific class of DBPs. There are two published studies on the production of nitrosamines, covering a range of pharmaceuticals, pesticides and other emerging pollutants [6,9]. In both cases, the conclusions are that there are several pharmaceuticals capable of yielding between a 1 and 10% conversion efficiency to N-nitrosodimethylamine (NDMA), while in the case of the antacid drug ranitidine, the molar conversion rate to NDMA is between 60 and 80%. Also, Duirk et al. have investigated the formation of iodinated DBPs upon chlorination and chloramination of waters containing iodinated X-ray contrast compounds [92]. They concluded that oxidation with hypochlorite was responsible for the production of hypoiodite, which then further reacted with natural organic matter to yield the iodinated DBPs. Actually, those DBPs were neither produced in pure water nor during chloramination at basic pHs. At acidic pHs (about 6) monochloramine is back converted to hypochlorite and can therefore start the reaction.

Transformation Products of Emerging Contaminants 153

4.6.3 Permanganate and Ferrate Mn(VII) is widely used in water treatment, mostly as a primary oxidant, in order to control odor, taste and color, Fe(II), Mn(II) and reduce the amount of dissolved organic matter prior to further disinfection with chlorine or other chlorinated agents. Among its advantages are the lack of halogenated DBPs and the production of MnO2, which can be easily separated by filtration/coagulation and that can also enhance the elimination of micropollutants by physical adsorption [93]. The low cost of KMnO4 and stability make it also a good option for in situ treatment and remediation of polluted water resources. At natural water pHs (10.4

91 85 75 83 29 92 88 83 78 70 82 87 85 54 43 95–70d decreases 90–85 70–78 80–90

Estradiola Atrazinea Simazinea Sulfamethoxazoleb Primidoneb Tartaric acidc Ibuprofena Ibuprofenc Estronec a

Uncharged hydrophobic compounds. Uncharged hydrophilic compounds. c Charged organic compounds. d R decreases with time. e RO-reverse osmosis, NF-nanofiltration, PA-polyamide, PS-polysulfone, PPA-polypiperazine, CA-cellulose acetate, SPES-sulfonated polysulfone. b

Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 167

Other studies [43] have focused on the evaluation of the parameters affecting membrane degradation, such as chlorine, typically found in the form of hypochlorite solution, and monochloramine, residuals used in the treatment of secondary effluents to avoid post-treatment biological growth. These and other oxidizing substances can cause structural changes in the polyamide membrane layer [44]. The main problem for membrane cleaning systems is fouling, which can lead to significant changes in the physicochemical properties of the membrane surface, affecting the separation mechanisms, including size exclusion and electrostatic interactions and thus the way in which the membranes interact with water and solutes [45]. Fouling is often regarded as an impediment, since it decreases membrane permeability and thus requires elevated pressures to maintain operational flows. However, these are not destructive techniques. The concentrates usually contain organic matter, micropollutants and wastewater treatment residuals (i.e., soluble microbial products, partially biodegraded organics, and anti-scaling chemicals) [46], and cannot be discharged into the environment, so they must be treated to minimize their environmental impact. P
erez-Gonz
alez et al. [34] reviewed concentrate treatment by AOPs. Sometimes the concentrate treated by AOPs is recycled back to the biological systems (conventional active sludge, CAS, see Figure 5.3) [32]. Studies on EC degradation by AOPs in NF/RO concentrates are scarce. Recent studies [47–51] have demonstrated that these techniques are very efficient concentrate treatments, but still must be optimized, as must their operating costs. Several key points about this are discussed in some detail in the section on solar AOPs. It should be highlighted that NF/RO systems are currently combined with membrane bioreactors (MBRs) for advanced treatment of their effluents [52,53]. There are lab-scale and full-scale studies on removal of ECs. Abdulhakem et al. [53] reported that the NF/RO membranes supplement MBR treatment very well, resulting in above 99% removal of most of the 40 trace organic contaminants selected in their study. This option is of great interest, because the effluent comes from UF, so no pretreatment is needed before NF/RO, whereas with CAS pretreatment would be required (MF/UF, coagulation, adsorption granular active, etc.) (Figure 5.3).

5.4 Membrane Bioreactors (MBR) A membrane bioreactor (MBR) combines the CAS process with membrane filtration. Membranes (pore sizes from 0.05 to 0.4 mm) can complete retention of bacterial flocs and virtually all the suspended solids, producing a high-quality effluent [54]. The MBR potential

Effluent +ECs

NF/RO

CAS

AOP Concentrate + ECs

MBR

Concentrate + ECs

UF

WW + ECs For reuse Permeate

AOP

Effluent Pretreatment

For reuse Permeate

WW+ECs

NF/RO

Figure 5.3 General diagram of two combination membrane–AOP systems for EC treatment.

168 Transformation Products of Emerging Contaminants in the Environment

for efficiently removing hazardous substances from wastewater has also been highlighted as it can operate with higher biomass concentrations than the CAS processes. Moreover, as all the bacteria are retained, the microorganisms are better adapted to mineralizing micropollutants and even hydrophobic substances that tend to accumulate on the sludge, which never reaches the effluent [55]. Thus the combination of water scarcity and more restrictive legislation has driven up market growth of the MBR technology, which since the turn of the millennium has increased by an average of 11.6–12.7% per year [56,57] as it presents many advantages over CAS treatments. This has also led to an increasing interest of researchers in the MBR technology and, indeed, the number of publications has grown around 20% since 1995, especially those on micropollutant (including ECs) treatment, which have increased the most (45%) [58]. A search done for “MBR” and “micropollutants” with the Scopus search engine produced a total of 63 articles. A word cloud generated with all the titles (Figure 5.4) associated word size with the frequency with which they appear in the titles (www.wordle.net). Thus, apart from general terms like removal or wastewater, it can be seen how pharmaceutical removal, comparisons of CAS and MBR, or landfill leachate treatment are frequent topics of study in this field. Biodegradation and sorption have been reported as the main micropollutant removal processes, where both are directly related to bioavailability, that is, the accessibility of micropollutants to the activated sludge population [59]. Abiotic degradation and volatilization can also take part in micropollutant elimination, however, the first has been found to be of limited importance [60,61] and the second depends on several micropollutant characteristics (like Henry’s constant and octanol–water partition coefficient) as well as some other operating factors like gas flow rate. Hydrophobicity and chemical structure seem to be the most important physicochemical characteristics affecting micropollutant removal. In general, hydrophobic compounds are eliminated by adsorption on the sludge particles in the system. Compounds containing complex structures and toxic groups show higher resistance to biodegradation [62].

Figure 5.4 Word cloud generated with the titles of articles found from a search for MBR and micropollutants in Scopus.

Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 169

The MBR operating parameters also have an important role in micropollutant elimination. Due to the high sludge retention time, the possibility for sludge to adapt to persistent organic pollutants increases. Tambosi et al. (2010) reported higher removal efficiencies for all the compounds in a pharmaceutical mixture when the solids retention time is increased from 15 to 30 days [54]. Another factor that positively affects the removal of micropollutants is the smaller bacterial floc size than in CAS processes, which minimizes the distances to be overcome by the substrate during diffusion into the flocs. The relationship between temperature and micropollutant removal has also been studied. Many authors have reported an improvement in micropollutant degradation rates with increase in temperature. For example, Hai et al. [63] reported that the removal of the hydrophilic compounds in a mixture of 22 micropollutants was lower at 10  C than at 30  C, whereas hydrophobic compound removal was stable between 10 and 30  C. In contrast, elimination of both groups of compounds deteriorated when the process was operated at 45  C. Consequently, important differences in micropollutant degradation rates can be expected when the MBRs are operated in different seasons and, indeed, it has been found that these differences are more important in MBRs than in CAS processes, which can attenuate variations in temperature better due to their larger surface. Finally, it should be mentioned that, although few studies focus on pH, this parameter also influences the removal of micropollutants as it affects both sorption and biodegradation [62]. Although the MBR has many theoretical advantages over CAS processes, the elimination of micropollutants in MBRs improves only slightly, and many micropollutants are not or are only partly eliminated or transformed [64]. Thus, the MBRs are not the definitive technology for coping with micropollutants, although they could be a good pretreatment for more advanced processes, like reverse osmosis, nanofiltration or AOPs, due to their high-quality effluents, as shown in Figure 5.3. Research currently studying this option, as in recent articles (e.g., [32,65,66]), has reported promising results with removal rates over 99% for all the compounds studied in the effluent.

5.5 AOPs Including Solar AOPs Among the advanced technologies able to degrade micro-pollutants and specifically ECs, advanced oxidation processes (AOPs) are a particularly attractive option [29,46]. Although there are different reactor systems (see http://www.jaots.net/), all of them are characterized by the same chemical feature: production of hydroxyl radicals (OH) able to oxidize and mineralize almost any organic molecule, yielding CO2 and inorganic ions. Rate constants (kOH) for formation of OH radicals, given by the expression (r ¼ kOH [OH] C) for most reactions involving hydroxyl radicals in aqueous solution, are usually of the order of 106 to 109 M
1 s
1. They are also characterized by their non-selective attack, which is a useful attribute for solving wastewater treatment pollution problems. The versatility of the AOPs is also enhanced by the fact that there are different ways of producing hydroxyl radicals, facilitating compliance with the specific treatment requirements. Methods based on UV, H2O2/UV, O3/UV and H2O2/O3/UV combinations (as described in the ozonation section) use photolysis of H2O2 and ozone to produce the hydroxyl radicals. Heterogeneous photocatalysis and homogeneous photo-Fenton are based on the use of a wide-band-gap semiconductor and addition of H2O2 to dissolved iron salts, respectively, and irradiation with UVA–Visible light [67–69].

170 Transformation Products of Emerging Contaminants in the Environment

Some of the disadvantages associated with AOPs are their high operating costs, depending on the specific process: (i) high electricity demand (e.g., ozone and UV-based AOPs), (ii) the relatively large amounts of oxidants and/or catalysts consumed (e.g., ozone, hydrogen peroxide, and iron-based AOPs), slow kinetics (photocatalysis with TiO2) and (iii) pH restrictions (e.g., Fenton and photo-Fenton). Nevertheless, by using solar energy as a light source, optimizing the pH and the amounts of oxidant/catalyst, processes like photo-Fenton, may be used for commercial applications. 5.5.1 Solar Driven Advanced Oxidation Processes The illumination of a wide-band-gap semiconductor with light energy greater than its bandgap energy produces excited high-energy states of electron and hole pairs (e-/hþ). The heterogeneous solar photocatalytic detoxification process consists of making use of the near-ultraviolet (UV) band of the solar spectrum (wavelength shorter than 400 nm), to photo-excite a semiconductor catalyst in contact with water and in the presence of oxygen. Under these circumstances, oxidizing species, either bound OH or free holes, react with oxidizable contaminants. With a typical UV-flux near the surface of the Earth of 20 to 30 W m
2, the Sun makes 0.2 to 0.3 mol photons m
2 h
1 in the 300 to 400 nm range available to the process. TiO2 is the most commonly used photocatalyst (band-gap energy of 3.2 eV, corresponding to approximate photon energy of < 400 nm). Although there are many different sources of TiO2, Degussa (now Evonik) P25 TiO2, has effectively become a standard [70], because it has (i) a reasonably well-defined nature (i.e., typically a 70 : 30 anatase:rutile mixture, non-porous, BET surface area 55  15 m2 g
1, average particle size 30 nm), (ii) a substantially higher photocatalytic activity than most other readily available (commercial) TiO2, (iii) low cost and low toxicity, (iv) chemical stability. Other semiconductor particles, for example, CdS or GaP absorb larger fractions of the solar spectrum and can form chemically activated surface-bond intermediates, but unfortunately, these photocatalysts are degraded during the repeated catalytic cycles involved in heterogeneous photocatalysis, generating toxic dissolved heavy metals in water, as reported by many authors in the 1990s and now part of the state of the art [71]. Fenton and Fenton-like processes are probably among the longest and most widely applied AOPs, and first proposals for wastewater treatment applications were reported in the 1960s. Yet it was not until the early 1990s, when scientists working in the field of environmental sciences published results on the role of iron in atmospheric chemistry, which attracted the attention of scientists and engineers working in wastewater treatment [68]. Irradiation with light up to 580 nm leads to photoreduction of dissolved ferric iron to ferrous iron. The primary step in the photoreduction of dissolved ferric iron is a ligand-to-metal charge-transfer reaction, followed by dissociation of intermediate complexes by irradiation, forming Fe2þ. The ligand can be any Lewis base able to form a complex with ferric iron (OH-, H2O, R-COO-, R-OH, R-NH2, etc.). Depending on the reacting ligand, the product may be a hydroxyl radical, such as in reaction (5.4), or another radical derived from the ligand. The direct oxidation of an organic ligand is possible, as shown for carboxylic acids in reaction (5.5). ½FeðOHÞ2þ þ hy ! Fe2þ þ OH

(5.4)

½FeðOOC
RÞ2þ þ hy ! Fe2þ þ CO2 þ R

(5.5)

Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 171

The ferric iron complex has different light absorption properties depending on the ligand and, therefore, the pH plays a crucial role in the efficiency of the photo-Fenton reaction because it strongly influences which complexes are formed. Thus, pH 2.8 was frequently postulated as an optimum pH for photo-Fenton treatment because at this pH precipitation does not take place, and the dominant iron species in solution is [Fe(OH)]2þ, the most photoactive ferric iron–water complex. The original solar photoreactor designs for photochemical applications were based on line-focus parabolic-trough concentrators (PTCs). In part, this was a logical extension of the historical emphasis on trough units for solar thermal applications. Furthermore, PTC technology was relatively mature and existing hardware could be easily modified for photochemical processes. The first outdoor engineering-scale reactors developed were a converted solar thermal parabolic-trough collector in which the absorber/glazing-tube combination was replaced by a simple Pyrex glass tube through which contaminated water could flow [72]. The main disadvantages of these solar collectors were described in the late 1990s [73], and solar concentrating devices were disregarded for photocatalysis. In the mid-1990s, the first non-concentrating collectors for photocatalysis were described as cheaper than solar concentrating devices because they have no moving parts or solar tracking devices. Extensive efforts in the design and testing of small non-tracking collectors resulted in several different non-concentrating solar reactors. However, the design of a robust non-concentrating photoreactor is not simple, and the use of these photoreactors was abandoned [74]. A series of constraints must be fulfilled in the design of a solar collector for photocatalysis: (i) effective collection of UV–Vis radiation, (ii) working temperatures as close as possible to ambient to avoid loss of volatile organic compounds, (iii) take into account that quantum efficiency decreases with irradiance, and that (iv) concentrators collect 1/RC (concentration factor, RC, ratio between radiation collection area and photoreactor area) of the available diffuse radiation and thus the concentration for photocatalytic applications should be RC ¼ 1. Finally, (v) its construction should be economical and easy, and (vi) the reactor should maintain a low pressure drop. Therefore, the use of tubular photoreactors has a decisive advantage because of the inherent structural efficiency of tubing. Tubing is also available in a large variety of materials and sizes and is a natural choice for a pressurized fluid system. State-of-the-art solar collector technology in photocatalysis focuses on low-concentration collectors called compound parabolic concentrators (CPCs). They are an attractive option combining the characteristics of both parabolic concentrators and stationary flat systems: they concentrate solar radiation, but like flat plate collectors, they are stationary and collect diffuse radiation [75]. Until about 15 years ago, only two engineering-scale demonstrations, one for groundwater treatment in the United States and another for industrial wastewater treatment in Spain (at Plataforma Solar de Almer
ıa) had been reported [72]. Since then, several installations have been constructed and various engineering systems have been developed. There are several examples in the literature, and they may be found in any scientific database with keywords such as compound parabolic collector or photocatalysis (Figure 5.5). Therefore, the technology may be considered mature for application [76], but it should be mentioned that most of the applications have been developed for industrial and non-biodegradable contaminants, and the information available on degradation of ECs under realistic conditions (i.e., real effluents from MWTP) is scarce.

172 Transformation Products of Emerging Contaminants in the Environment

Figure 5.5 View of CPC wastewater treatment plant. The inset shows the typical shape of a CPC. Image provided courtesy of PSA.

5.5.2 Different Approaches for Treating ECs by Solar AOPs TiO2 has been widely tested for removal of ECs, mainly working with model compounds dissolved in demineralized water. But studies in real MWTP secondary biological treatment effluents are unusual. Indeed, results when applying this solar AOP are discouraging, perhaps due to the low photonic efficiency of the process, the scavenging of OH radicals by components of the water (carbonates, chloride, etc.) and photocatalyst deactivation by adsorption of other components of MWTP effluents. Figure 5.6 shows an example of the degradation of

5000 I

a in um

ill

0 45 85 1 00 5 17 28 0 290 47 5

–1

10000

c [ng L ]

15000

n tio

e tim

in [m

1 1 14 5 1112 3 1 9 0 7 8 6 5 3 4 1 2

]

Figure 5.6 Degradation profile of 15 contaminants with an initial concentration over 1000 ng L
1. (1, ibuprofen; 2, hydrochlorothiazide; 3, diuron; 4, atenolol; 5, 4, AA; 6, diclofenac; 7, ofloxacin; 8, trimethoprim; 9, gemfibrozil; 10, 4, MAA; 11, naproxen; 12, 4, FAA; 13, 4, AAA; 14, caffeine; 15, paraxanthine) using solar heterogeneous photocatalysis with TiO2.

Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 173

different contaminants. Samples of MWTP effluents were analyzed by HPLC-QTRAP-MS and 66 target micropollutants were identified and quantified, but only 16 of those contaminants, found at initial concentrations over 1000 ng L
1, made up over 88% of the total effluent pollutant load. The starting dissolved organic carbon (DOC) of 23.2 mg L
1 remained almost constant during the whole photocatalytic treatment. It is quite clear that heterogeneous photocatalysis with TiO2 is not efficient for complete removal of micropollutants as the treatment time must be longer than 500 min. Figure 5.7 shows a conventional photo-Fenton experiment with minimal Fe (5 mg L
1) and minimal initial H2O2 (50 mg L
1) concentrations for the degradation of ECs in MWTP effluents from a municipal plant located in Almer
ıa (southeast Spain), which employs an activated sludge biological treatment. Wastewater was used as received within the following 3 days. Initial COD (chemical oxygen demand) and DOC were 49 and 26 mg L
1, respectively. Analyses (HPLC-Qtrap-MS) revealed the presence of 60 different microcontaminants at concentrations ranging from 3 ng L
1 (fenofibrate) to 43 660 ng L
1 (caffeine). Figure 5.7 shows the 22 contaminants detected at >100 ng L
1. During this experiment, 35% of the DOC, which varied from 26 to 17 mg L
1, was removed. H2O2 consumption was 80 mg L
1 at the end of the experiment (50 min solar illumination). The quite significant degradation of 97% of the initial concentration of the sum of all contaminants was achieved in less than one hour of treatment. But, in general, rates are slow due to the originally low concentrations of the contaminants. AOP degradation of pollutants at low concentrations follows pseudo-first-order kinetics (r ¼ kapC). Therefore, one of the solutions for increasing process efficiency would be to increase C0. The combination of AOPs with membrane processes under these conditions (see Figure 5.3)

–1

c [ng L ]

40000 30000 20000 10000

5 –2 illu m ina tio

0

n

25

tim

e

[m

in]

50

8 1 12 0 1 1 4 18 6 2 22 0

6

4

2 0

Figure 5.7 Solar photo-Fenton degradation of ECs in MWTP effluents. The ECs are: 1, 4-AAA; 2, 4-FAA; 3, 4-MAA; 4, antipyrine; 5, atenolol; 6, caffeine; 7, ciprofloxacin; 8, cotinine; 9, diclofenac; 10, diuron; 11, furosemide; 12, gemfibrozil; 13, hydrochlorothiazide; 14, ketoprofen; 15, naproxen; 16, nicotine; 17, ofloxacin; 18, paraxanthine; 19, ranitidine; 20, sulfamethoxazole; 21, sulfapyridine; 22, trimethoprim.

174 Transformation Products of Emerging Contaminants in the Environment

has attracted attention during the last few years as the concentrations in retentates are much higher than in raw effluents [34]. As mentioned above, membrane technologies produce water for reuse and a concentrate. These concentrates usually contain organic matter, micropollutants and wastewater treatment residuals (i.e., soluble microbial products, partially biodegraded organics, and anti-scaling chemicals) [46], which cannot be discharged into surface water so they must be treated to minimize their environmental impact. A combination of conventional NF and photo-Fenton is a more efficient tertiary treatment than photo-Fenton applied directly to the MWTP effluent. ECs dissolved in water were concentrated by NF, discharging a permeate free of contaminants. The retentate was then recirculated until the desired pre-concentration was achieved (up to 10 times). The concentrate was treated with solar photo-Fenton and compared to direct photo-Fenton treatment of the NF feed-in. Photo-Fenton kinetics and H2O2 consumption could be evaluated as the key parameters for comparing results. Results of experiments performed with NF/solar photo-Fenton (at different pre-concentrations) and solar photo-Fenton at 75 mg L
1 may be seen in Figure 5.8. The treatment time necessary to degrade 95% of the micropollutants by solar photo-Fenton only at starting concentrations of 75 mg L
1 was 89 min, with hydrogen peroxide consumption of 17.0 mg L
1. For NF/solar photo-Fenton with pre-concentration 4 and 10 times the treatment time needed to degrade 95% was quite similar, and hydrogen peroxide consumption was almost the same. So neither H2O2 consumption nor treatment time changed significantly with NF. It is, therefore, very clear that by increasing the starting concentration, process efficiency increases with initial reaction rate. It is also worth mentioning the substantial improvement in treatment time needed for 95% degradation, considering that, for example, the time needed to treat 750 mg L
1 is very similar to that for 75 mg L
1 but the volume would be ten times smaller after NF. Therefore, the solar treatment plant could also be ten times smaller. Combination of NF and solar photo-Fenton produced two important advantages in process efficiency. First, concentrating contaminants by NF produced a clean effluent (permeate) and a concentrated stream with higher concentrations of contaminants. By applying solar photo-Fenton to 800

Fenton photo-Fenton

–1

H2O2 consumption (mg L )

20

15

–1

ΣC (μg L )

600

400

–1

C0=75 μg L

10

pre-concentration 4 times pre-concentration 10 times 200

5

0

0 –25

0

25

50

75

100

125

illumination time (min)

Figure 5.8 Profile of solar photo-Fenton and NF/solar photo-Fenton degradation of 75 mg L
1 of ECs pre-concentrated 4 and 10 times and hydrogen peroxide required.

Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 175

this stream, hydrogen peroxide consumption and treatment time were both reduced compared with direct treatment of the water by photo-Fenton, as hydrogen peroxide was used more efficiently and the reaction rate was higher.

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Approaches to Water and Wastewater Treatment for Removal of Emerging Contaminants 177 36. Nghiem, L.D., Schafer, A.I., and Elimelech, M. (2006) Role of electrostatic interactions in the retention of pharmaceuticals active contaminants by a loose nanofiltration membrane. Journal of Membrane Science, 286, 52–59. 37. Le-Minh, N., Khan, S.J., Drewes, J.E., and Stuetz, R.M. (2010) Fate of antibiotics during municipal water recycling treatment processes. Water Research, 44, 4295–4323. 38. Van der Bruggen, B., Verliefde, A., Braeken, L. et al. (2006) Assessment of a semi-quantitative method for estimation of the rejection of organic compounds in aqueous solutions in NF. Journal of Chemical Technology and Biotechnology (Oxford, Oxfordshire: 1986), 81, 1166–1176. 39. Nghiem, L.D., Schafer, A.I., and Elimelech, M. (2005) Pharmaceutical retention mechanisms by nanofiltration membranes. Environmental Science & Technology, 39, 7698–7705. 40. Kimura, K., Amy, G., Drewes, J.E. et al. (2003) Rejection of organic micropollutants (disinfection by-products, endocrine disruptuting compounds, and pharmaceuticals active compounds) by NF/ RO membranes. Journal of Membrane Science, 227, 113–121. 41. Al-Rifai, J.H., Khabbaz, H., and Sch€afer, A. (2011) Removal of pharmaceuticals and endocrione disrupting compound in a water recycling process using reverse osmosis systems. Separation and Purification Technology, 77, 60–67. 42. Nghiem, L.D. and Schaffer, A.I. (2012) Critical risk points of nanofiltration and reverse osmosis processes in water recycling applications. Desalination, 187, 303–312. 43. Simon, A., Price, W., and Nghiem, L. (2006) Effects of chemical cleaning on the nanofiltration of pharmaceutically active compounds (PhACs). Separation and Purification Technology, 88, 208–215. 44. Guo-Dong, K., Cong-Jie, G., Wei-Dong, C. et al. (2007) Study on hypochlorite degradation of aromatic polyamide reverse osmosis membrane. Journal of Membrane Science, 300, 165–171. 45. Nghiem, L.D. and Hawkes, S. (2007) Effects of membrane fouling on the nanofiltration of pharmaceutically active compounds (PhACs): Mechanisms and role of membrane pore size. Separation and Purification Technology, 57, 176–184. 46. Westerhoff, P., Moon, H., Minikata, D., and Crittenden, J. (2009) Oxidation of organics in retentates from reverse osmosis wastewater reuse facilities. Water Research, 43, 3992–3998. 47. Dialynas, E., Mantzavinos, D., and Diamadopoulos, E. (2008) Advanced treatment of the reverse osmosis concentrate produced during reclamation of municipal wastewater. Water Research, 42, 4603–4608. 48. Zhang Pagilla, Y.K. (2010) Treatment of malathion pesticide wastewater with NF and photoFenton oxidation. Desalination, 263, 36–44. 49. Abdelmelek, S.B., Greaves, J., Ishida, K.P. et al. (2011) Removal of pharmaceuticals and personal care products from RO retentate using AOPs. Environmental Science & Technology, 45, 3665–3671. 50. Miralles-Cuevas, S., Arqu
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178 Transformation Products of Emerging Contaminants in the Environment 56. BCC (2008) Membrane bioreactors: Global markets, in BCC Report MST047B, June. 57. Meng, F., Chae, S.R., Drews, A. et al. (2009) Recent advances in membrane bioreactors (MBRs): Membrane fouling and membrane material. Water Research, 43, 1489–1512. 58. Santos, A., Ma, W., and Judd, S.J. (2011) Membrane bioreactors: Two decades of research and implementation. Desalination, 273, 148–154. 59. Clara, M., Strenn, B., Ausserleitner, M., and Kreuzinger, N. (2004) Comparison of the behaviour of selected micropollutants in a membrane bioreactor and a conventional wastewater treatment plant. Water Science and Technology, 50, 29–36. 60. Soares, A., Murto, M., Guieysse, B., and Mattiasson, B. (2006) Biodegradation of nonylphenol in a continuous bioreactor at low temperatures and effects on the microbial population. Applied Microbiology and Biotechnology, 69, 597–606. 61. Ivashechkin, P., Corvini, P.F.X., and Dohmann, M. (2004) Behaviour of endocrine disrupting chemicals during the treatment of municipal sewage sludge. Water Science and Technology, 50, 133–140. 62. Cirja, M., Ivashechkin, P., Sch€affer, A., and Corvini, P.F.X. (2008) Factors affecting the removal of organic micropollutants from wastewater in conventional treatment plants (CTP) and membrane bioreactors (MBR). Reviews in Environmental Science and Biotechnology, 7, 61–78. 63. Hai, F.I., Tessmer, K., Nguyen, L.N. et al. (2011) Removal of micropollutants by membrane bioreactor under temperature variation. Journal of Membrane Science, 383, 144–151. 64. Siegrist, H. and Joss, A. (2012) Review on the fate of organic micropollutants in wastewater treatment and water reuse with membranes. Water Science and Technology, 66, 1369–1376. 65. Sahar, E., Ernst, M., Godehardt, M. et al. (2011) Comparison of two treatments for the removal of selected organic micropollutants and bulk organic matter: Conventional activated sludge followed by ultrafiltration versus membrane bioreactor. Water Science and Technology, 63, 733–740. 66. Dolar, D., Gros, M., Rodriguez-Mozaz, S. et al. (2012) Removal of emerging contaminants from municipal wastewater with an integrated membrane system, MBR-RO. Journal of Hazardous Materials, 239–240, 64–69. 67. Comninellis, C., Kapalka, A., Malato, S. et al. (2008) Advanced oxidation processes for water treatment: advances and trends for R&D. Journal of Chemical Technology and Biotechnology (Oxford, Oxfordshire: 1986), 83, 769–776. 68. Pignatello, J.J., Oliveros, E., and MacKay, A. (2006) Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. Critical Reviews in Environmental Science and Technology, 36, 1–84. 69. Shannon, M.A., Bohn, P.W., Elimelech, M. et al. (2008) Science and technology for water purification in the coming decades. Nature, 452, 301–310. 70. Herrmann, J.M. (2010) Photocatalysis fundamentals revisited to avoid several misconceptions. Applied Catalysis B: Environmental, 99, 461–468. 71. Pelizzetti, E. and Minero, C. (1994) Metal oxides as photocatalyst for environmental detoxification. Comments on Inorganic Chemistry, 15, 297–337. 72. Goswami, D.Y. (1997) A review of engineering developments of aqueous phase solar photocatalytic detoxification and disinfection processes. Journal of Solar Energy Engineering-transactions of the ASME, 119, 101–107. 73. Alfano, O.M., Bahnemann, D., Cassano, A.E. et al. (2000) Photocatalysis in water environments using artificial and solar light. Catalysis Today, 58, 199–230. 74. Malato, S., Blanco, J., Vidal, A., and Richter, C. (2002) Photocatalysis with solar energy at a pilotplant scale: an overview. Applied Catalysis B: Environmental, 37, 1–15. 75. Colina-M
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6 Transformation Products of Emerging Contaminants Formed during Advanced Oxidation Processes Ioannis K. Konstantinou1, Maria Antonopoulou1 and Dimitra A. Lambropoulou2 1

Department of Environmental and Natural Resources Management, University of Patras, Greece 2 Environmental Pollution Control Laboratory, Department of Chemistry, Aristotle University of Thessaloniki, Greece

6.1 Introduction Throughout the last decades, the contamination of water resources by organic micropollutants, their metabolites and transformation products (TPs) has received increasing scientific attention. Of particular concern are the so-called “emerging contaminants” (ECs) which include a wide spectrum of compounds belonging to different chemical groups, such as pharmaceuticals, personal care products (PCPs), pesticides, veterinary products, endocrine disruptors (EDs), industrial compounds/by-products, surfactants, fire retardants, fuel additives, food additives and engineered nano-materials, among others [1,2]. Their widespread occurrence in various environmental matrices, in wastewaters and even in drinking water [2,3], has been recognized as a major environmental problem with consequences that are not yet fully understood [4]. Conventional wastewater treatments often achieve incomplete removal of such persistent contaminants, resulting in their discharge into surface water [5]. In the last few years, advanced oxidation processses (AOPs), such as ozonation (O3), ozone in comination with hydrogen peroxide (O3/H2O2) and/or UV light (UV/O3, UV/O3/H2O2), UV photolysis (Ph), Fenton (F) reagent, Photo- and Electro- Fenton (Ph-F, El-F), photocatalysis with TiO2 (Ph-TiO2) and polyoxometalates (POM), electrochemical oxidation (El-Ox), ultrasonication (US), catalytic wet air oxidation (CWAO), elecron beam radiolysis (Rad), gammaTransformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

180 Transformation Products of Emerging Contaminants in the Environment

radiation (GR) etc. [6–8], based on the generation of highly reactive radical species, have attracted tremendous attention for the removal of a broad range of organic pollutants from water. This is due to their potential to destroy hazardous contaminants and not simply transfer them to another phase, as do conventional treatment methods [9]. However, apart from the achievement of pollutant abatement, the identification of the TPs formed during transformation processes is of increasing relevance in evaluating environmental impact, since various intermediates are formed en route to complete mineralization. Some of these TPs can be toxic and, in some cases, more persistent than the parent compound. Therefore, full structural elucidation of AOP TPs requires advanced analytical methods that combine high separation efficiency with a maximum of structural information. Gas chromatography-mass spectrometry (GC-MS) and liquid chromatography mass spectrometry (LC-MS) techniques are among the most frequently used methods for thermally stable and polar thermally labile TPs, respectively [10,11]. Several studies have been reported in the literature in the last decades regarding the removal of a great number of ECs by different AOPs. However, the number of studies that report the degradation and mineralization rates in different aquatic media outweigh by far those that identify the TPs and the reaction mechanisms involved in these processes. The identification of TPs and their ecotoxicological effects on the environment have recently started to attract the interest of the scientific community. However, the extent of the problem has not yet been adequately evaluated and there is too little reliable information on their potential environmental risk. Actually, acute tests are mostly used to evaluate their effects. In most cases, the Microtox test was employed to assess the overall toxicity of the solution, that is, the sum of compounds formed and not only those that have been identified. With the Microtox method, the comparison was carried out by monitoring the bioluminescence of the marine bacterium Vibrio fischeri as a function of illumination time [10]. The highly challenging task regarding monitoring and identification of TPs is fundamental and is the key to maximizing the overall process efficiency. Thus, our attention in this chapter has been mainly focused on providing an overview of the identified TPs, the reaction types and mechanisms of selected categories of ECs, during their treatment by various AOPs. The contaminants were classified in terms of the characteristic structural groups. A great effort has been made to classify the data in generalized reaction pathways for each chemical family. Furthermore, for the classes that were feasible, the major TPs are depicted in generalized structural groups, while for the remainder, the chemical names of the identified TPs are presented in tables. TP formation has been observed at least once for the reported processes in the corresponding figures.

6.2 Pesticides Water pollution by pesticides has emerged as a major problem in many countries because of the high persistence of these organic contaminants in aquatic environments and the potential detrimental health impacts [12]. The first general category discussed in this chapter includes representative groups of widely used and frequently detected pesticides, namely organophosphorus (OP) insecticides, anilide herbicides (ANHs), phenylurea herbicides (PUHs) neonicotinoid insecticides (NCIs) and the individual compound glyphosate, in a broad range of environmental water samples, including, wastewaters, surface water and groundwater. 6.2.1 Organophosphorus Insecticides 6.2.1.1 Aliphatic Organophosphorus Insecticides Several investigations reported the effective decomposition of aliphatic organophosphorus compounds by different AOPs, through a series of TPs and reaction pathways which are

Transformation Products of Emerging Contaminants 183

presented explicity in Table 6.1. The more common TPs were the more toxic oxon derivatives [13–19] and different monoalkyl, trialkyl and dialkyl phosphorothioate or phosphate esters [14,17–20]. 6.2.1.2 Aromatic Organophosphorus Insecticides Details of the TPs and mechanisms of the transformation of four organophosphorus pesticides (OPPs) possessing different aromatic rings (chlorpyrifos (CPF), diazinon (DZN), azinphos methyl (AZM), and methidathion (MTD)) are presented in Table 6.2. As common
S to P

O (oxidative desulfuration) and hydrolysis of the phospathways, oxidation of P
phorothioate ester group occur. CPF is subjected to dechlorination, while MTD and AZM are subjected to the ring opening of the triazolidine and benzotriazinic moiety, respectively. Some additional pathways were, however, recognized [14,24–32]. 6.2.1.3 Aromatic (Benzene Ring) Organophosphorus Insecticides In this section, the principal organic TPs and the transformation pathways followed during the transformation of six widely used aromatic OP insecticides, having the chemical structure depicted in Table 6.3, by various AOPs are discussed. The transformation pathways and the major TPs are presented in Table 6.4 and in Figure 6.1, respectively. Oxidant attack of the  OH radical on the P

S bond and subsequent substitution of sulfur by oxygen occurred for the six insecticides using all the mentioned AOPs, resulting in the formation of the oxon derivatives, which are considered to be more toxic than the parent compounds [18,19,33– 48]. Notably, these oxon analogues have been reported to exhibit higher inhibition of cholinesterase activity compared to their parent compounds [35]. The second major critical route corresponding to the rupture of the P

O bond (hydrolysis of the ester group) resulted in the formation of the respective phenols for each insecticide (e.g., 4-nitrophenol for methyl and ethyl parathion, 2,4-dichlorophenol for dichlofenthion) and different trialklyl and dialkyl phosphorothioates and phosphates [18,19,33–48]. Subsequent successive dehalogenation, loss of different groups and/or hydroxylation of the aromatic ring yielded hydroxy and dihydroxy phenol as well as benzoquinone derivatives [33,37,42,46]. The nature of the TPs showed that the substituents were not only abstracted from the aromatic ring but also added to them with 3,5-dimethyl-4 methylthiophenol and 2,4-dinitrophenol being identified [34,38,39,42,43,46]. The S-alkyl isomers, which are considered to exhibit a very different biological activity than their O-alkyl precursors, were also identified for three OPPs (fenitrothion, dichlofenthion, and methyl parathion) [18,19,33,35,38]. The transformation of -NO2 substituent to amino group was identified in the case of nitro-OPs while sulfoxide and sulfone intermediates have been detected for OPPs bearing-SCH3 group, like fenthion [33,35,45]. 6.2.2 Anilide Herbicides (ANHs) Several studies have shown the effective decomposition of ANHs (Table 6.5) by different AOPs, through a series of various aromatic TPs following several transformation pathways, presented in Table 6.6. Hydrolytic dechlorination, ring and alkyl chain hydroxylation and oxidation, N-dealkylation and amide bond cleavage reactions can be invoked to justify the formation of the TPs depicted in Figure 6.2 [49–57]. Cyclization TPs with

Transformation Products of Emerging Contaminants 187 Table 6.3 Molecular structure of the studied aromatic OPs. X1

S R1O R2O

P

R3

O

X2 X3

OPs

R1

R2

R3

X1

X2

X3

Fenthion (F) Fenitrothion (FNT) Dichlofenthion (DF) Bromophos methyl (BM) Parathion ethyl (PE) Parathion methyl (PM)

Me Me Et Me Et Me

Me Me Et Me Et Me

Me Me H H H H

H H Cl Cl H H

SMe NO2 Cl Br NO2 NO2

H H H Cl H H

mainly quinoline and indole rings, among others, were also detected for the majority of the studied ANHs by different AOPs [34,52,55,57]. Propionic acid was generated due to the breakdown of the aromatic ring in the transformation of alachlor by direct ozonation and the O3/H2O2 process. Formic, acetic and oxalic acids were generated from either breakdown of the aromatic ring or dealkylation and further oxidation of the side chains. Cleavage of the chloroacetyl group led to the formation of monochloroacetic acid, as well [55]. Pignatello and Sun [49] reported also the formation of serine as the final TP of metolachlor resulting from the photoassisted Fenton reaction [49]. 6.2.3 Phenylurea Herbicides (PUHs) The transformation of phenylurea herbicides (PUHs) (diuron, isoproturon, fenuron, monuron, chlortoluron, linuron) presented in Table 6.7 leads to the formation of a great number of TPs. The major TPs and the transformation pathways followed are depicted in Figure 6.3 and Table 6.8, respectively. Partial or complete dealkylation of the substituted N-alkyl chains is one of the dominant processes [60–86]. The dealkylation mechanism was suggested to proceed via decarboxylation reactions and the formation of the corresponding alcohol, aldehyde and acid derivatives [68–71,82]. It is worth mentioning that N-di-demethylated and formylated photoproducts of both isoproturon and chlorotoluron were reported to be more toxic for Vibrio Fischeri than the parent PUHs [75,87]. Sequential dealkylation leads to dealkylated ureas, followed by hydrolysis to the corresponding aniline derivatives, [61,64,66,70,82,83] with chloroanilines being highly toxic compounds with potential mutagenic and oncogenic activities [88]. Hydroxylation of the aromatic ring and/or alkyl groups of the urea moiety are also significant processes. In the case of isopropturon, OH addition to the N atom attached to the aromatic ring and on the isopropyl group as well as dehydrogenation were also observed [71–73]. The repeated OH attack on the aromatic ring formed phenolic-type derivatives and quinonidal structures, leading to the opening of the aromatic ring [68,82]. As final TPs, carboxylic acids, such as oxalic, acetic and formic acids, have been reported [66]. At the same time, another route proceeds

Transformation Products of Emerging Contaminants 191 R3

X1 R1O R2O

R3

X1

O

X1

S

P

R1O

X2

O

R2O

P

X2

O

X3

R1/R1'

X2

O

X3

BM, F, FNT, DF, PM, PE Ph-TiO2, Ph, UV/POM, US/Fer/UV, UV/H2O2, Ph-F, US, Fenton, El-F, US/TiO2, El-Ox (Ti/Pt anode)

R3

X3 F, FNT, PM, PE Ph-TiO2, Ph, UV/POM, UV/H2O2, US

X1

R3

S O R1O R2O

P

R1O

O

R1 (R2)

P

O

+

R (R )

1 2 R2O BM,DF, PM, PE Ph-TiO2,Ph,UV/H2O2, Ph-F, US

+

F, FNT,PM, PE Ph-TiO2,Ph, Ph-F, US

R1O

X3

P

S

R1 (R2)

R3

R3

S R1O

' P O R1 R2O FNT,DF,PM, PE Ph-TiO2, UV/POM, Ph, Ph-F, US DF,FNT,PE Ph-TiO2, UV/H2O2, El-F

R2O

X2

BM,F, FNT, DF,PM, PE Ph-TiO2,Ph, UV/POM, US/Fer/UV, UV/H 2O2, Ph-F, US/TiO2, US, Fenton, El-F, El-Ox (Ti/Pt anode)

+

O

HO

(R2') + HO

OH

FNT,PM Ph-TiO2,UV/POM, El-F, El-Ox, (Ti/Pt anode),

+

X2

OH HO

X3

FNT,PM, PE Ph-TiO2, El-F, US

O R1O R2O

P

O

R1' (R2')

DF,FNT,PM, PE Ph-TiO2, Ph, UV/H2O2 El-F, El-Ox (Ti/Pt anode), US

R3 OH

HO

OH

PM Ph-TiO2

O

O FNT,PM, PE Ph-TiO2,UV/POM, El-F, El-Ox (Ti/Ptanode), US

R1', R2' 90% of the BDE 209 was excreted in the faeces within 72 h, a very small amount (99% of the bile BDE 209 was present as metabolites [78]. Rats could metabolize BDE 209 to faecal metabolites via oxidative debromination, leading to the formation of mono-OH- and ortho-MeO-BDEs [57,78]. Oxidative metabolism of BDE 209 would imply the formation of the arene-oxide intermediate, but its formation is unlikely on the fully brominated rings and, therefore, debromination probably precedes oxidation [57]. The lower levels of BDE 209 reported in biota are not necessarily explained by its metabolism in organisms, but are rather due to its relatively lower absorption rate compared to the other PBDEs. Another important aspect to be addressed is the toxicity associated with PBDE exposure which may arise from the formation of HO-PBDE. In vitro studies using HLMs showed that several HO-PBDEs are more active than PBDEs in their ability to compete with thyroid hormones for binding to serum transport proteins [79], to inhibit hepatic thyroxine metabolism [62], and to inhibit aromatase activity [80].

18.3 Transformation Products of HBCDs As an additive FR, HBCD easily reaches the environment through production, usage as FR in a wide variety of household consumer products and recycling of HBCD-containing materials [81]. The commercial HBCD mixture consists of three major isomers, a-, b- g- in a ratio of 10, 10, and 80% of the mixture, respectively. Although its use is not restricted, HBCD has been included in the PBT (Persistent, Bioaccumulative and Toxic) list of the European Chemical Substance Information System. Despite its small contribution to HBCD global production and use, a-HBCD is the major stereoisomer found in biota. This finding has driven the research in the direction of elucidation of the potential pathways which would explain such behaviour. 18.3.1 Degradation of HBCDs in Abiotic Matrices Since HBCD is used as an FR, the thermal stability and susceptibility to undergo photolytic degradation, and also the nature/toxicity of the generated products, become important for

558 Transformation Products of Emerging Contaminants in the Environment

safe use in consumer products. In general, two mechanisms were evidenced for the HBCD degradation: debromination with elimination of either HBr or Br2 and further generation of tetra-bromocyclododecene, di-bromocyclododecadiene, and cyclododecatriene; isomerisation processes in which stereoisomer interconversion takes place without the loss of any other secondary products (Table 18.1). Thermal gravimetric analyses were performed to study the decomposition of relatively pure HBCDs [9,10]. Thermal degradation showed an auto-accelerating behaviour, possibly catalysed by organic intermediates formed in the decomposition process. The thermal stability was influenced by sample purity and a limited effect of isomeric composition of the HBCD sample was noticed on the decomposition temperatures. Additionally, HBCDs followed the same degradation pathway when exposed to moderate heat (10  C min1), independently of the atmosphere in which they were placed, since the formed products were not different. Due to thermal stress, around 35% of the total HBCD was degraded, from which about 75% of bromine (by weight) was released as HBr, while 25% was involved in the formation of high-molecular-weight aromatic and polyaromatic brominated decomposition compounds [9,10]. These results indicated that hexa-, penta-, and tetrabrominated polyaromatic structures were not the primary products of HBCD decomposition [9,10]. Other interesting findings were obtained when temperature was tested as a factor for the isomerisation of HBCDs from various products where they are added, like expanded (EPS) or extruded polystyrenes (XPS). Early experiments suggested that HBCDs isomerise under thermodynamic control (190  C) to an equilibrium mixture of 78, 14, and 8% of a-, b-, and g-HBCDs [82]. Recent experiments investigated the influence of the materials in which HBCDs are added and the subsequent isomerisation processes. Therefore, by increasing the temperature (140 to 160  C) at which EPS and XPS previously spiked with HBCDs were kept, the proportions of g-HBCD decreased over time from 80 to 21% in exposed EPS, whereas those of a- and b-HBCDs increased from 6 to 46% and from 13 to 30%, respectively, with d-HBCD remaining at 1–3% [83]. Such results are consistent with previous data which showed that under kinetic control (110–140  C) enantiomerically pure HBCDs isomerise regio- and stereoselectively with (þ)g-HBCD converting to (þ)a-HBCD and ()g-HBCD converting to ()a-HBCD [84]. Such isomerisation processes were observed only in EPS, but not in XPS. Probably, the XPS sample was already rich in a-HBCD before thermal exposure and proportions increased only slightly from 43 to 54%, whereas those of b- and g-HBCDs decreased from 36 to 25% and from 21 to 19%, respectively, with d-HBCD remaining constant at 1%. Therefore, while the HBCD patterns of EPS and XPS differed considerably before exposure, they become similar upon thermal treatment, resembling more and more the pattern of equilibrated HBCD mixtures [82,83]. Increased temperatures (>110  C) during fabrication processes have already been reported to influence the isomerisation of HBCDs, leading to higher percentages of a-HBCD [11]. In total, results on the isomerisation processes which take place in products or in pure compounds may be informative for the HBCD profile measured in various matrices (e.g. indoor dust), where, in general, a-HBCD tends to have a higher contribution [85]. These results may have further implications for the human exposure of HBCDs via dust ingestion, since a-HBCD has already been shown to possess the highest bioaccumulation potential and longest half-life compared to b- and g-isomers [27]. Recently, photodegradation of HBCD in aqueous solutions containing methanol as co-solvent was reported [86]. Although the experiment was not exhaustive in measuring the generated TPs, it was concluded that UV radiation may degrade HBCD through a

Transformation Products of Brominated Flame Retardants (BFRs) 559

debromination mechanism, while hydroxy-products of the debrominated HBCD were proposed as intermediates of photolysis. Additionally, degradation was directly proportional to the HBCD initial concentration and debromination was also more efficient at acidic or alkaline pH values than at neutral pH [86]. As they bind strongly to solid particles (dust, soil, sediment or sewage sludge), HBCDs were detected in several matrices, following the general rule: the closer to the emission source, the higher the expected levels. Sequential degradation of 14 C-HBCDs was tested with activated sludge, digester sludge, soil or freshwater aquatic sediments in microcosms [87]. Substantial transformation of a-, b-, and g-[14 C] HBCDs was observed in wastewater sludge and in freshwater aquatic sediment microcosms, each prepared under aerobic and anaerobic conditions. Based on the identified TPs, the proposed mechanism implied debromination via dihaloelimination with the subsequent formation of a double bond [87]. Microorganisms naturally occurring in aquatic sediments and anaerobic digester sludge might thus mediate complete debromination of HBCDs. Contradictory findings were reported when these processes were tested against the hypothesis that different half-lives should be expected for individual HBCD stereoisomers, to justify the selective bioaccumulation of a-HBCD. While almost a double half-life was reported for a-HBCD degradation under anaerobic conditions in digester sludge compared to b-, and g-HBCDs in one particular study [88], no significant difference in their degradation was reported for almost similar degradation conditions [87]. However, significant differences in tested concentrations might have influenced these findings [87]. In indoor dust, an important matrix for human exposure, photo-degradation and/or isomerisation of HBCDs have been reported to occur, possibly via elimination of HBr. Using separation (GC or LC) and detection (MS) techniques, the TPs were identified as pentabromocyclododecenes (four isomers) and tetrabromo-cyclododecadienes (two isomers) [89]. 18.3.2 Biotransformation Pathways for HBCDs A hypothesis was recently launched that prolonged exposure of a biological system to increased loads of environmental contaminants can trigger the evolution of microorganisms to gain the ability to metabolize such pollutants [90]. For example the bacterial strain Sphingobium indicum B90A, isolated from HCH contaminated soils, expresses several enzymes that break CCl bonds. When tested in vitro for the possibility of breaking down CBr bonds, the haloalkane dehalogenase LinB expressed by the same bacterial strain could transform all HBCDs, with (þ)b- and (þ)g-HBCDs being degraded fastest. The initially formed pentabromocyclododecanols (HO-PBCDs) were further transformed to tetrabromocyclododecadiols (di-HO-TBCDs), while at least seven mono- and five di-hydroxylated products were distinguished by LC-MS after these degradation experiments [90]. However, it remains to be proven if such in vitro transformations can also be observed in vivo. Additional in vitro testing with liver microsomes (rat, polar bear, beluga whale, ringed seal) evidenced some important directions to be further investigated in the in vivo HBCD metabolism: formation of several phase I metabolism products (HO-compounds) through oxidative metabolism, dehydrobromination or debromination, stereospecific metabolism (bioisomerisation) with isomeric interconversion (Table 18.2). Research in this field evidenced a change in the diastereoisomeric HBCD patterns from the domination of g-HBCD in the technical mixture [91] and sediments [92] to the domination of a-HBCD in biota [81]. Although HBCD standard mixtures are mostly racemic, the

560 Transformation Products of Emerging Contaminants in the Environment

composition of biotic samples often showed a non-racemic pattern [27,93]. Additionally, the calculated total human intake of HBCDs seems to be high compared to the levels found in human tissues, despite their bioaccumulative characteristics, and recent reports concerning an average a-, b- and g-HBCD bio-accessibility of 77% from the human gastro intestinal tract after dust ingestion [94]. There is still a debate on the cause of these findings which might involve either physico-chemical influences or different metabolic pathways. Later, one particular study applied an in vitro hepatic assay using microsomes extracted from different animal species (polar bear, beluga whale, ringed seal, rat) and tested for aand g-HBCD in vitro depletion/metabolite formation [30]. In all cases, debromination of HBCD isomers appeared to be the predominant metabolic pathway in the in vitro microsomal assays, although no other potential metabolites, for example HO-HBCDs, were investigated. Dehydrobromination was reported to take place through the loss of the [–HBr] and [2HBr], but [3HBr] products were also detected in some of the incubation extracts. As expected, a-HBCD was by far the most predominant parent isomer relative to b- or g-HBCD. In the case of the St. Lawrence River beluga, a-HBCD typically comprised >99% of the total HBCD burden [30]. However, other studies showed that, besides debromination, oxidative metabolism is one of the most important biotransformation pathways for HBCDs [31,32]. By using subcellular rat liver fractions, phase I metabolites of a- and g-HBCDs were formed, mediated by the CYP P450 enzyme system. The major metabolites were reported to be OH-HBCDs, followed by intermediates like HO-PBCDs and also PBCDs as minor metabolites [32,33]. The formation of the hydroxylated phase I metabolites after incubation of rat liver microsomes with enantiomers of a-, b-, and g-HBCDs was also confirmed by other studies [31], while the degradation of all HBCD isomers followed first-order kinetics. Similar patterns of mono-HO-HBCDs were found in environmental samples (e.g. pollack, mackerel muscle and herring gull eggs) [31]. The formation of the mono- and di-HO metabolites of HBCDs, together with the presence of PBCDs and TBCDs, was tested in three wildlife species (tern egg, seal, and flounder), and also in Wistar rats exposed to 30 and 100 mg HBCD/kg bw/day for 28 days [34]. While mono-HO-HBCDs and PBCDs were detected in tern eggs from the Western Scheldt and in the blubber of harbour seals (Wadden Sea), no HO-HBCDs could be detected in the tissue of flounder (Wadden Sea) [34]. Human milk samples collected from the UK were analysed for HBCDs and TPs [27]. Enantioselective enrichment of ()-a-HBCD (average EF ¼ 0.29) was observed, indicating the potential for enantioselectivity associated with HBCD absorption, metabolism and/or excretion. Although the PBCD and TBCD derivatives formed via stepwise elimination of HBr were also detected, it is not yet clear whether these compounds originate from in vivo biotransformation of HBCDs and/or as a result of their intake via ingestion of indoor dust [27]. The interconversion from g- to a-HBCD through bioisomerisation was also tested to explain the higher contribution of a-HBCD to the total HBCDs measured generally in biota samples. Therefore, absorption, distribution, metabolism, and excretion parameters of g-HBCD with respect to dose and time following a single acute exposure and repeated exposures in adult female C57 BL/6 mice was investigated [28]. After exposure, g-HBCD was rapidly metabolized and eliminated in urine and faeces. In vivo stereo-isomerisation was observed: g-HBCD was converted to b-HBCD in liver and brain tissues and into a-HBCD and b-HBCD in fat and faeces. Polar metabolites in the blood and urine were a major factor

Transformation Products of Brominated Flame Retardants (BFRs) 561

in determining the initial whole-body half-life (1 day) after single exposure. Elimination, both whole-body and from individual tissues, was biphasic. Initial half-lives were approximately 1 day, whereas terminal half-lives were up to 4 days, suggesting limited bioaccumulation potential for g-HBCD [28]. However, in a similar exposure study, a-HBCD was excreted in the faeces as parent compound and metabolites, whereas urine only contained metabolites [29]. Initial half-lives for a-HBCD were estimated at 1–3 days, while longer terminal half-lives of 17 days were observed, suggesting a higher potential for its bioaccumulation compared to g-HBCD [29]. Although the toxicological literature of the TPs of HBCDs, for example HO-HBCDs, TBCDs or PBCDs, is scarce, higher binding affinities of PBCDs to human transthyretin receptor (hTTR) than the parent HBCDs were reported. Interestingly, two trans-PBCD isomers showed higher binding affinities to hTTR than T4 [95].

18.4 Transformation Products of TBBPA Tetrabromobisphenol A ([TBBPA, 4,40 -isopropylidenebis(2,6-dibromophenol)]) is the most widely used and distributed BFR worldwide [57]. TBBPA can be used as an additive BFR and as an alternative to octa-BDE in television casings, printer components, fax machines, photocopiers, coffee makers and plugs/sockets [96]. TBBPA is non-volatile [57] and degrades very slowly photochemically ([15,16]). TBBPA has a log Kow of 4.75 at a pH of 7.6 [97]. The widespread use of TBBPA and its detection in dust, sediments and biota has led to increasing concerns regarding its effects on wildlife and humans [96]. Since it was measured in humans, TBBPA may be a matter of concern to human health [98]. Therefore, it is imperative to know the environmental fate of the parent molecule. Several studies on rats, aquatic organisms, and/or microorganisms have indicated that metabolic Phase II conjugation and debromination of TBBPA occur in biota and, thus, the literature addressed in the following paragraphs is focused on its environmental degradation and metabolism (Tables 18.1 and 18.2). 18.4.1 Degradation of TBBPA in Abiotic Matrices In the direct photolysis of TBBPA with UV irradiation and oxidation by MnO, some intermediate products (phenols and ketones) have been identified ([15,16,99–101]). Eriksson et al. [15,16] investigated the photochemical transformation of TBBPA in an aqueous medium, and proposed that the primary photochemical reaction was the cleavage between one of the benzene rings and the isopropyl group. The reductive degradation of TBBPA over iron–silver bimetallic nanoparticles was assumed to be governed by debromination [102] and the same transformation mechanism was also confirmed by an anaerobic reductive reaction [103]. Generally, TBBPA might be degraded under simulated light irradiation, and it can be catalysed by the presence of compounds like bismuth oxybromide. The transformation mechanisms involve debromination, hydroxylation or demethylation, processes which lead to the mineralisation of TBBPA (Figure 18.3, from [17]). The photodegradation of TBBPA in the environment may result from direct photolysis as well as reaction with sunlight-generated reactive oxygen species, such as singlet oxygen. Previous studies have shown that there are environmental sources of singlet oxygen, such as the humic acids [99] which could contribute to the degradation of TBBPA.

562 Transformation Products of Emerging Contaminants in the Environment Br

Br CH3

OH

OH CH3

I

Br

–Br HO

OH

Br II

III

OH

OH CH3

OH

OH

Br

CH3

CH3

HO

OH A

Br CH3

OH OH

HO

OH

Br O

C

OH CH3 Br

CH3

OH CH3

Br

OH

CH3

B

Br

D

C

OH HO

CH3

E

OH

OH

OH HO

Br

OH

CH3

CH2OH

OH CH2OH

HO

G

Br

OH

–Br OH F

OH Br

CH2OH HO CH2OH Br

H

OH

OH Br Aromatic ring opened products

CO2+H2O

Figure 18.3 Proposed photocatalytic degradation pathway for tetrabromobisphenol A (TBBPA). Identified possible reaction intermediates of TBBPA by bismuth oxybromide and UV–VIS conditions. Compounds A, B, C and D were formed by debromination or the direct attack of
OH. Compound E was formed by demethylation through the route of TBBPA ! A ! E. Compound G was produced by the pathway of TBBPA ! A ! E ! G through hydroxylation. Compound F and H were generated by the route of TBBPA ! B/C ! F or TBBPA ! B/C ! F ! H, with demethylation reaction. Reproduced with permission from Xu et al. (2011) Copyright (2011) Elsevier Ltd.

The combustion of domestic products containing TBBPA may lead to formation and release of polybrominated dibenzo-p-dioxins and furans (PBDD/Fs). Large amounts of brominated and mixed chloro-bromo dioxins and furans can be formed in accidental fires where BFRs are present [104]. Brominated light hydrocarbons, as well as phenolic derivatives, were observed during thermal decomposition (combustion and pyrolysis) of TBBPA [105].

Transformation Products of Brominated Flame Retardants (BFRs) 563

The authors also reported at least six tetra- to hexa-BDD/Fs. TBBPA decomposition was carried out at temperatures between 180 and 270  C, by using thermogravimetry and a laboratory-scale batch reactor. The results confirmed that TBBPA processing at temperatures above the melting point (180  C) may cause the release of TBBPA in the environment due to evaporation [106]. Lin et al. [107] investigated the debromination of TBBPA by nZVI in methanol:water (1:1) solutions. More than 86% of TBBPA was debrominated within 16 h in a solution containing 3.0 g/L of nZVI at pH 7.5. A higher loading of nZVI and acidic conditions facilitated the debromination process, while coexisting Ca2þ and Naþ species inhibited the reaction. The debromination of TBBPA yielded tri-BBPA, di-BBPA, BBPA, and BPA and released Br ions. This study suggests that nZVI may be potentially employed to debrominate TBBPA in soil and sediment. At low and neutral pH, TBBPA is virtually insoluble in water and, therefore, soil mobility is expected to be minimal. However, at higher pH, the solubility of TBBPA increases significantly. The biodegradation of TBBPA was measured in various soil matrices under aerobic conditions [108]. An 80% reduction of TBBPA was observed over 45 days of anaerobic degradation of a highly contaminated soil [109]. TBBPA converted into bisphenol-A (BPA) by reductive debromination under anaerobic conditions. Voordeckers et al. [110] examined bio-debromination of TBBPA in anoxic estuarine sediments. Complete debromination of TBBPA to BPA was also observed under both methanogenic and sulfate-reducing conditions. During the anaerobic stage, TBBPA is reductively debrominated to BPA, which is completely mineralized under bacterial aerobic conditions [109]. A novel bacterium, Ochrobactrum sp. T, capable of simultaneous debromination and aerobic mineralisation of TBBPA, (Figure 18.4, from [111]) was isolated from sludge collected from an electronic waste recycling site [111]. Degradation and debromination efficiencies of 92 and 87%, respectively, were achieved within 72 h. Eight metabolic intermediates were identified during the biodegradation of TBBPA. This study was the first to propose a one-step process for TBBPA debromination and mineralisation by a single bacterial strain. Another study reported that TBBPA debromination under anaerobic conditions was achieved using a bacterial and archaeal consortium, in which strains of Pelobacter carbinolicus and Sphaerochaeta sp. seemed to play an important role [112]. Degradation of TBBPA by Trametes versicolor CCBAS 612 was probably the first example of using a white rot fungus for this purpose [113,114]. O-methylation is thought to be an alternative pathway to TBBPA degradation, but the process is not yet fully understood. 18.4.2 Biotransformation Pathways for TBBPA The absorption of TBBPA has been investigated in some aquatic species, mostly fish. The bluegill sunfish (Lepomis macrochirus) rapidly absorbed TBBPA, at 9.8 mg/L [115], and equilibrium was achieved in less than 3 days. Upon removal from dosed water, TBBPA was rapidly eliminated from all tissues. The whole body half-life of TBBPA was calculated to be less than 24 h. TBBPA was not found in mussels in Osaka Bay, Japan, but the di-MeO derivative of TBBPA was detected at concentrations up to 5 ng/g [116]. The di-MeO-TBBPA metabolite has also been associated with TBBPA in sediments taken both upstream and downstream from a plastics manufacturer [117]. The levels of both compounds were similar upstream (24–34 ng/g), but increased dramatically downstream. The di-MeO-TBBPA metabolite was present at 1500 ng/g, while the TBBPA was present at 270 ng/g. No conclusions

564 Transformation Products of Emerging Contaminants in the Environment OH Br

OH

O

OH

Br Br

Br Br

Br

Br

Br +

+ (A) H C 3

CH3 (B)

Br

OH

OH

ute

OH

Br

(C)

OH

1

Br

C

OH Br

Br

Br

Br

Ro

Br

(E)

(D)

Br

OCH3

O Br

OH Br

OH

Br O Route II

Br

Br OH

CO2 +H2O +Br–

Br

OH

Br OH (F)

OH (G)

(H)

Figure 18.4 Proposed aerobic tetrabromobisphenol A (TBBPA) biodegradation pathway by the isolated bacterial strain (Ochrobactrum sp. T.) The eight degradation intermediates were identified. Two distinct pathways occurred during biodegradation: Route I (oxidation route): intermediates (A) and (B) are produced directly from the decomposition of oxidative TBBPA, leading to intermediate (C) and 2,6-dibromophenol (not detected) after dehydrogenation and decomposition. Then, intermediate (C) is further converted to 2,4,6-tribromophenol (2,4,6-TBP) (D). Intermediate (E) also could be produced by the methylation reaction via intermediate (B). The ring cleavage product is further oxidized to CO2 and Br-1 finally. Route II (reductive debromination): intermediate BPA (G) is produced by the complete reductive debromination of TBBPA via TriBBPA (F). With prolonging degradation time, one benzene ring of BPA is further cleaved to produce intermediate (H). Reproduced with permission from An et al. Copyright (2011) Elsevier Ltd.

were drawn as to whether the metabolite formed bacterially or by metabolism in an aquatic species. Fini et al. [38] observed the extensive and rapid uptake of 14 C-TBBPA in tadpoles, with >94% of 14 C-TBBPA being metabolized within 8 h. Four metabolites (TBBPA glucuronide, TBBPA-glucuronide-sulfate, TBBPA-sulfate, and TBBPA-disulfate) were identified in tadpole extracts. These metabolites are identical to the TBBPA conjugates characterized in mammals, including humans. Most radioactivity (>75%) was associated with the sulfated conjugates.

Transformation Products of Brominated Flame Retardants (BFRs) 565

In an early study, Brady [118] administered a single oral dose (6.5–7.5 mg/kg) of C-TBBPA to rats and concluded that it was poorly absorbed from the gastrointestinal (GI) tract. However, because the dose was administered orally, it was unclear whether the compound was poorly absorbed from the intestines or TBBPA was excreted unchanged in the bile. An intraperitoneal dose of TBBPA was administered to female rats in high amounts (250 and 1000 mg/kg) [42]. Urine and faecal excretion of 14 C-TBBPA was followed for 72 h. Cumulative urine excretion was very low (90%), however, approximately 10% of the faecal 14 C-TBBPA was the debrominated tribromobisphenolA (TriBBPA). The formation of TriBBPA was concluded to have occurred in the rat liver and been excreted into the faeces via the bile, but the alternate possibility that TriBBPA was formed by microbial action in the gut could not be excluded. A study using both conventional and bile duct cannulated male Sprague–Dawley rats concluded that 14 C-TBBPA was readily absorbed from the GI tract [119]. Nearly 92% of a single oral dose of TBBPA (2.0 mg/kg) was excreted in the faeces of conventional rats. The half-life of TBBPA in all rat tissues was estimated to be less than 3 days [119]. TBBPA was distributed into most of the tissues, but cumulative residues were only 2% of the administered dose (2.0 mg/kg), and no one tissue contained more than 1% of the dose [119]. Following the oral dose in the male rat, the overwhelming majority (>90%) of the extractable radioactivity in faeces was found to be TBBPA [119]. No metabolites of TBBPA could be detected or characterized in any of the faecal extracts. Three major metabolites of TBBPA have been identified as polar conjugates in rat bile. The glucuronide ether, the diglucuronide ether, and the glucuronide ether–sulfate ester diconjugate of TBBPA were also evidenced [119]. Parent TBBPA was the only compound detected in rat faeces. The conjugated metabolites were apparently deconjugated back to the parent compound by intestinal microorganisms, and subsequently eliminated in the faeces. Conversely, most studies on the metabolic fate of TBBPA have been performed in mammals, the main metabolites identified being sulfate and glucuronide conjugates in rats [39–41]. Schauer et al. [41] also reported a predominance of Phase II biotransformation pathway products, concluding that TBBPA was completely metabolized in rats and humans into sulfates and glucuronides and then excreted in urine. The half-life of TBBPA in blood serum of occupationally-exposed Swedish workers [120] was estimated at 2.2 days, and TBBPA was thus rapidly depleted. A single oral dose of 0.1 mg/kg TBBPA was administered to five human subjects [41]. TBBPA-glucuronide and TBBPA-sulfate were identified as metabolites of TBBPA in blood and urine of the human subjects. In blood, TBBPA-glucuronide was detected in all human subjects, whereas TBBPA-sulfate was only present in two individuals. Maximum plasma concentrations of TBBPA-glucuronide (16 nmol/L) were obtained within 4 h after administration. In two individuals where TBBPA-sulfate was present in blood, maximum concentrations were obtained at the 4 h sampling point; the concentrations rapidly declined below the limit of detection (LOD) after 8 h. Parent TBBPA was not present in detectable concentrations in any of the human plasma samples. TBBPA-glucuronide was slowly eliminated in urine to reach the LOD in 124 h after administration. Absorption of TBBPA from the GI tract and rapid metabolism of the absorbed TBBPA by conjugation result in a low systemic bioavailability of TBBPA. Biotransformation studies of TBBPA by human and rat subcellular liver fractions have shown that TBBPA undergoes oxidative cleavage at the quaternary carbon atom to give brominated phenols, including 2,6-dibromohydroquinone [121]. 14

566 Transformation Products of Emerging Contaminants in the Environment

18.5 Transformation Products of NBFRs “Novel” brominated flame retardants (NBFRs) are defined as BFRs which are new to the market or newly observed in the environment [122]. These chemicals have been produced in lower volumes than the major BFRs, usually either as alternatives for them or with a very specific intended use and, not so frequently, as replacements for the FRs that have been banned or phased out of use. 18.5.1 Decabromodiphenyl Ethane (DBDPE) DBDPE is a high production volume chemical produced in the US and China that has been proposed as a replacement for BDE 209 [123]. As concerns mount about the gradual debromination of BDE 209 to lower and more toxic brominated congeners [26], the use of DBDPE has consequently increased [124]. Abiotic compartment : Similar to BDE 209, DBDPE also undergoes thermal degradation (Table 18.1) at a comparable temperature, but besides debromination or intermolecular ring closure, it also generates bromotoluenes [12]. Two nona-brominated degradation products of DBDPE were observed even in the technical mixture [125]. When exposed to strong UV radiation (125 W), DBDPE degraded to a number of hepta- to nona-BDPEs [13]. Under natural sunlight conditions, this degradation was not observed in plastic matrices [8]. Biotic compartment : In a study on rats, Wang et al. [13] compared peaks corresponding to the main photodegradation products of DBDPE with peaks from an extract of DBDPEexposed rat liver tissue. They concluded that debromination alone is not the main biotransformation pathway (Table 18.2). By using GC-EI/MS, two metabolites were tentatively identified: a methyl sulfone substituted nona-BDPE and the ethyl sulfone analogue. However, definite identification was not possible due to lack of standards for the above compounds. Apparently, reductive debromination was followed by phase I and phase II (glutationation þ transformation to methylsulfone) metabolism. 18.5.2 1,2-Bis(2,4,6-tribromophenoxy)ethane (BTBPE) BTBPE has been produced since the mid-1970s, in a high production volume in the US and a low production volume in the EU [126]. In the US, it was marketed by the Great Lakes Chemical Corporation as FF-680 with the intended use as additive FR for thermoplastic and thermoset resin systems (e.g. ABS, HIPS, polycarbonate, etc.). As of 2005, Chemtura used the FF-680 as a replacement for the octa-BDE formulation [127]. Abiotic compartment : The main decomposition products of BTBPE found through rapid removal by pyrolysis were TBP and vinyl 2,4,6-tribromophenyl ether [14] (Table 18.1). However, a prolonged contact with a heat source can also produce hydrogen bromide, ethylene bromide, polybrominated vinyl phenyl ethers, poloybrominated diphenyl ethers and dibenzodioxins. Most of them are even more toxic and reactive than BTBPE and, therefore, the follow up of these TPs is mandatory. Biotic compartment : The elimination of BTBPE from rats was studied by administering one single oral dose of 14 C-BTBPE and subsequently estimating the amount of radioactivity remaining in tissues and eliminated in faeces [37]. Most of the BTBPE (94%) was found in faeces within 72 h. Tissue retention was minimal, in the most lipophilic tissues, and the metabolites were excreted in urine, bile and faeces, at very low levels. The most significant

Transformation Products of Brominated Flame Retardants (BFRs) 567 Br Br

Br OCH2CH2O

Br

Br

OH

Br III, Isomer I -Br

Br

epoxide hydralase dihydrodial dehydrogenase

Br

Br

Br

OCH2CH2O

arene axide ring opening

OH Br

Br

OCH2CH2O

Br

arene axide ring opening

Br

Br

Br

Br

Br

Br

Br

BTBPE

Br Br

VI

Br HO

Br

OCH2CH2O

-2Br

arene axide ring opening

Br

Br

Br OH

HO OCH2CH2O

OCH2CH2O Br

Br

Br Br

OH

Br

OCH 2CH 2OH V

O

O Br

Br

Br

Br

Br

Br

Br

Br OH

OH

+

either cleavage

CYP P450 axidation

-Br

I

Br OCH2CH2O

Br

OCH2CH2O

Br

CYP P450 axidation

arean axide ring opening

Br

O

-Br Br

OH

Br

Br

IV

Br

III. Isomer 2

Figure 18.5 Proposed pathway for the metabolism of 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE) in male rats based on the metabolism identified by mass spectrometry in faeces. Oxidation, oxidative debromination and ether cleavage are the favoured biotransformations. Reproduced with permission from Hakk et al. (2004) Copyright (2004) Elsevier Ltd.

TPs were the faecal ones (Table 18.2). The predominant compounds were those resulting after hydroxylation, debromination, and cleavage on either side of the ether linkage. By correlating the observed metabolites with the information available on the main biochemical pathways in rats, the authors have proposed a pathway for the metabolism of BTBPE in rats (Figure 18.5, from [37]). 18.5.3 2-Ethylhexyl2,3,4,5-tetrabromobenzoate (TBB) and Di(2-ethylhexyl)tetrabromophthalate (TBPH) These chemicals are produced by the Chemtura Chemical Corporation and used together in two FR formulations, namely Firemaster 550 (TBB:TBPH in a ratio of 4 : 1, by mass) [128] and Firemaster BZ-54 (2.5 : 1 ratio, by mass) [36]. Firemaster 550 was first introduced as a replacement for the penta-BDE mixture [128]. TBPH is also used in the DP-45 formulation, also marketed by Chemtura. In the US, this compound is classified as a high production volume chemical [129]. Abiotic compartment : TBB and TBPH are prone to photodegradation in organic solvents (toluene, methanol and tetrahydrofuran) and the photodegradation TPs were the

568 Transformation Products of Emerging Contaminants in the Environment

debrominated analogues, with two or three bromine atoms [5]. Some TPs of TBPH also lost both alkyl side chains (Table 18.1). These chemicals degrade photolytically more slowly than the PBDEs, suggesting that their persistence in the environment might be higher. Biotic compartment : TBB and TBPH have shown a slightly different metabolism during in vitro studies (Table 18.2); the first compound was rapidly metabolized in rat microsomes and HLMs to 2,3,4,5-tetrabromobenzoic acid (TBBA) [35]. However, for TPBH, no significant loss of the initial amount added at the beginning of the experiment was observed, and no metabolites were detected. After addition of NADPH, no significant difference in the formation rate of TBBA was observed, so a different metabolic pathway than oxidation was involved. Roberts et al. [35] hypothesized that the enzymes involved in TBB cleavage were carboxylesterases and tested this hypothesis successfully by incubating TBB with porcine carboxylesterases. The same class of enzymes also cleaved one of the ester bonds in TBPH, forming mono(2-ethylhexyl) tetrabromophthalate (TBMEHP). However, when Bearr et al. [36] exposed fathead minnows to Firemaster BZ-54, 10 potential metabolites were observed. Only two of them were identified: 2-ethylhexyl dibromobenzoate and 2,3,4,5-tetrabromo methylbenzoate. No TBBA was detected. The authors hypothesized that TBBA undergoes rapid methylation by methyltransferase enzymes, thus preventing the accumulation of TBBA.

18.6 Concluding Remarks and Future Perspectives To better explain the accumulation profiles of BFRs in wildlife and humans, the lower absorption potential of the higher brominated compounds together with species differences in their metabolism are important factors explaining the common findings in biota. Additionally, debromination of the higher brominated compounds, for example BDE 209, leading to more bioaccumulative and toxic compounds should be considered as important aspects when addressing the regulatory policies for such chemicals [130]. The findings of the above-mentioned studies have suggested that in many cases, TPs may represent a higher risk in terms of toxicity than the parent products. Elucidating the transformation or metabolic pathways of BFRs would help to better explain the common findings for the congener-specific profiles in environment/biota, while indications on the lower brominated compounds, recognized as more bioaccumulative and toxic, should be considered as important aspects when addressing the regulatory policies for such chemicals. Since unregulated chemicals may degrade in the environment and biota to restricted or banned substances, regulation policies should be better extended also over the first group. Additional restriction criteria for their use in specific consumer products should also consider conditions which favour the degradation to regulated compounds.

Acknowledgments ACD was financially supported through a postdoctoral fellowship from the Research Scientific Foundation-Flanders (FWO). ACI acknowledges the PhD scholarship funding through the Marie Curie ITN INFLAME. GM thanks the University of Antwerp (UA) for a post-doctoral fellowship (GOA project). AC acknowledges financial support from the FWO and UA.

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574 Transformation Products of Emerging Contaminants in the Environment 92. Janak, K., Covaci, A., Voorspoels, S. and Becher, G. (2005) Hexabromocyclododecane in marine species from the Western Scheldt Estuary: Diastereoisomer- and enantiomer-specific accumulation. Environmental Science & Technology, 39, 1987–1994. 93. Eljarrat, E., Guerra, P., Martinez, E. et al. (2009) Hexabromocyclododecane in human breast milk: levels and enantiomeric patterns. Environmental Science & Technology, 43, 1940–1946. 94. Abdallah, M.A.E., Harrad, S., Collins, C. and Tilston, E. (2009) Preliminary assessment of bioaccessibility of HBCDs from human GIT following indoor dust ingestion using a physiologically based extraction test (PBET). Organohalogen Compounds, 71, 100–104. 95. Weber, M., Durmaz, V., Becker, R. and Esslinger, S. (2009) Predictive identification of pentabromo-cyclododecene (PBCD) isomers with high binding affinity to hTTR. Organohalogen Compounds, 71, 247–250. 96. Covaci, A., Voorspoels, S., Abdallah, M.A.E. et al. (2009) Analytical and environmental aspects of the flame retardant tetrabromobisphenol- A and its derivatives. Journal of Chromatography A, 1216, 346–363. 97. Kuramochi, H., Kawamoto, K., Miyasaki, K. et al. (2008) Determination of physicochemical properties of tetrabromobisphenol A. Environmental Toxicology and Chemistry/SETAC, 27, 2413–2418. 98. Cariou, R., Antignac, J.P., Zalko, D. et al. (2008) Exposure assessment of French women and their newborns to tetrabromobisphenol-A: Occurrence measurements in maternal adipose tissue, serum, breast milk and cord serum. Chemosphere, 73, 1036–1041. 99. Han, S.K., Bilski, P., Karriker, B. et al. (2009) Oxidation of flame retardant tetrabromobisphenol A by singlet oxygen. Environmental Science & Technology, 42, 166–172. 100. Horikoshi, S., Miura, T., Kajitani, M. et al. (2008) Photodegradation of tetrahalobisphenol-A (X¼Cl, Br) flame retardants and delineation of factors affecting the process. Applied Catalysis BEnvironmental, 84, 797–802. 101. Lin, K., Liu, W. and Gan, J. (2009) Reaction of tetrabromobisphenol A (TBBPA) with manganese dioxide: kinetics, products, and pathways. Environmental Science & Technology, 43, 4480–4486. 102. Luo, S., Yang, S.G., Wang, X.D. and Sun, C. (2010) Reductive degradation of tetrabromobisphenol A over iron–silver bimetallic nanoparticles under ultrasound radiation. Chemosphere, 79, 672–678. 103. Arbeli, Z. and Ronen, Z. (2003) Enrichment of a microbial culture capable of reductive debromination of the flame retardant tetrabromobisphenol-A, and identification of the intermediate metabolites produced in the process. Biodegradation, 14, 385–395. 104. S€oderstr€om, G. and Marklund, S. (1999) Fire of a flame retarded TV. Organohalogen Compounds, 41, 469–472. 105. Ortu~no, N., Font, R., Molto, J. and Conesa, J.A. (2011) Thermal degradation of tetrabromobisphenol A: emission of polybrominated dibenzo-p-dioxins and dibenzofurans and other organic compounds. Organohalogen Compounds, 73, 511–514. 106. Marsanich, K., Zanelli, S., Barontini, F. and Cozzani, V. (2004) Evaporation and thermal degradation of tetrabromobisphenol. A above the melting point. Thermochimica Acta, 421, 95–103. 107. Lin, K., Ding, J. and Huang, X. (2012) Debromination of tetrabromobisphenol A by nanoscale zerovalent iron: kinetics, influencing factors, and pathways. Industrial & Engineering Chemistry Research, 51, 8378–8385. 108. World Health Organization (WHO) (1995) Environmental Health Criteria, no. 172, Tetrabromobisphenol A and Derivatives. International Programme on Chemical Safety. 109. Ronen, A. and Abelovitch, A. (2000) Anaerobic-aerobic process for microbial degradation of tetrabromobisphenol A. Applied and Environmental Microbiology, 66, 2372–2377. 110. Voordeckers, L.W., Fennell, D.E., Jones, K. and Haggblom, M.M. (2002) Anaerobic biotransformation of tetrabromobisphenol A, tetrachlorobisphenol A, and bisphenol A in estuarine sediments. Environmental Science & Technology, 36, 696–701. 111. An, T., Zu, L., Li, G. et al. (2011) One-step process for debromination and aerobic mineralization of tetrabromobisphenol-A by a novel Ochrobactrum sp. T isolated from an e-waste recycling site. Bioresource Technology, 102, 9148–9154.

Transformation Products of Brominated Flame Retardants (BFRs) 575 112. Iasur-Kruh, L., Ronen, Z., Arbeli, Z. and Nejidat, A. 2010. (2010) Characterization of anenrichment culture debrominating tetrabromobisphenol A and optimization of its activity under anaerobic conditions. Journal of Applied Microbiology, 109, 707–715. 113. Uhnakova, B., Petrickova, A., Biedermann, D. et al. (2009) Biodegradation of brominated aromatics by cultures and laccase of Trametes versicolor. Chemosphere, 76, 826–832. 114. Uhnakova, B., Ludwig, R., Peknicova, J. et al. (2011) Biodegradation of tetrabromobisphenol A by oxidases in basidiomycetous fungi and estrogenic activity of the biotransformation products. Bioresource Technology, 102, 9409–9415. 115. Nye, D.E. (1978) The bioaccumulation of tetrabromobisphenol, in the bluegill sunfish. Santa Clara, CA: Stoner Laboratories; Report to Velsicol Chemical Corporation, Chicago, submitted to WHO by the Brominated Flame Retardant Industry Panel. 116. Watanabe, I., Kashimoto, T. and Tatsukawa, R. (1983) Identification of the flame retardant tetrabromobisphenol A in the river sediment and the mussel collected in Osaka. Bulletin of Environmental Contamination and Toxicology, 31, 48–52. 117. Sellstrom, U. and Jansson, B. (1995) Analysis of tetrabromobisphenol-A in a product and environmental samples. Chemosphere, 31, 3085–3092. 118. Brady, U.E. (1979) Pharmacokinetic study of tetrabromobisphenol A (BP-4) in rats. Report from Athens, Georgia, University of Georgia (Report to Velsicol Chemical Corporation, Chicago, submitted to WHO by the Brominated Flame Retardant Industry Panel).  119. Hakk, H., Larsen, G.L., Bergman, A. and Orn, U. (2000) Metabolism, excretion, and distribution of the flame retardant tetrabromobisphenol-A in conventional and bile-duct Cannulated rats. Xenobiotica; The Fate of Foreign Compounds in Biological Systems, 30, 881–890. 120. Hagmar, L., Sjodin, A., Hoglund, P. et al. (2000) Biological half-lives of polybrominated diphenyl ethers and tetrabromobisphenol A in exposed workers. Organohalogen Compounds, 47, 198–201. 121. Zalko, D., Prouillac, C., Riu, A. et al. (2006) Biotransformation of the flame retardant tetrabromobisphenol A by human and rat sub-cellular liver fractions. Chemosphere, 64, 318–327. 122. Covaci, A., Harrad, S., Abdallah, M.A.E. et al. (2011) Novel brominated flame retardants: A review of their analysis, environmental fate and behaviour. Environment International, 37, 532–556. 123. World Health Organization (WHO) (1997) Environmental Health Criteria, no.192, Flame Retardants: A General Introduction. 124. Watanabe, I. and Sakai, S. (2003) Environmental release and behavior of brominated flame retardants. Environment International, 29, 665–682. 125. Kierkegaard, A. (2007) PhD Thesis: PBDEs in the environment. Time trends, bioaccumulation and the identification of their successor, decabromodiphenyl ethane. Stockholm University. 126. European Chemical Substance Information System (ECSIS) (2012) accessed in European Commission, Joint Research Centre, European Chemicals Bureau, available at http://esis.jrc.ec. europa.eu/. 127. Hoh, E., Zhu, L. and Hites, R.A. (2005) Novel flame retardants, 1,2-Bis(2,4,6-tribromophenoxy) ethane and 2,3,4,5,6-pentabromoethylbenzene, in United States’ environmental samples. Environmental Science & Technology, 39, 2472–2477. 128. Stapleton, H.M., Allen, J.G., Kelly, S.M. et al. (2008) Alternate and new brominated flame retardants detected in U.S. house dust. Environmental Science & Technology, 42, 6910–6916. 129. Health and Environmental Horizons (HEH) (2004) High production volume (HPV) challenge program test plan for phthalic acid tetrabromo bis 2-ethylhelxyl ester (CAS#26040-51-7). Prepared for the Brominated Phthalate Ester Panel, American Chemical Council. 130. European Chemicals Agency (ECHA) (2012) Proposals to identify Substances of Very High Concern current consultations: Bis(pentabromophenyl) ether (DecaBDE) http://echa.europa.eu/ proposals-to-identify-substances-of-very-high-concern.

19 Transformation Products of Alkylphenols Montserrat Cortina-Puig1, Gabino Bol
ıvar-Subirats2, Carlos Barata2 and Silvia Lacorte2 1

2

Escola Universitaria Salesiana de Sarri a, Spain Department of Environmental Chemistry, IDAEA-CSIC, Spain

19.1 Alkylphenols: Types, Properties and Uses Alkylphenols (APs) are starting materials for the production of many surfactants and antioxidant stabilizers used in commercial products like detergents, plastics and textiles. They are used in the synthesis of alkylphenol ethoxylates (APEs) and are also generated by the decomposition of these substances in the environment. APEs are one of the most widely used classes of nonionic surfactants in domestic and industrial products [1]. APs consist of one or more alkyl chains bound to a phenol. Their general structural formula is shown in Figure 19.1. The alkyl chain can be linear, as depicted in the figure, or branched, the most usual. The position of the hydroxy group on the aromatic ring may vary, although the predominant positional isomer is the para-isomer. APs are formed through the acid-catalyzed Friedel-Craft alkylation of a phenol with an olefin [2]. The olefins used vary from single species, such as isobutylene, to complicated mixtures, such as propylene tetramer (dodecene). The alkene reacts with phenol to produce monoalkylphenols, dialkylphenols and trialkylphenols. APs range generally from C1 to C16 alkyl groups. Most common members of the APs category are 4-cumylphenol, 2-sec-butylphenol, 2-tert-butylphenol, 4-tert-butylphenol, 4-sec-butylphenol, 2,4-di-tert-butylphenol, 2,6-di-tert-butylphenol, 2,4-di-tert-pentylphenol, 4-tert-amylphenol, 4-hexylphenol, 4-heptylphenol, 4-tert-octylphenol, 4-octylphenol, 4-nonylphenol, 4-dodecylphenol and 2,3,6-trimethylphenol. Among them, the most Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

578 Transformation Products of Emerging Contaminants in the Environment HO

n

Figure 19.1 Chemical structure of APs (n denotes the number of C atoms in the alkyl chain).

representative are nonylphenols (NPs), with a C9-alkyl group and octylphenols (OPs), with a C8-alkyl group. The OPs are manufactured by the catalytic reaction of phenol and diisobutylene to produce predominantly 4-(1,1,3,3-tetramethylbutyl)phenol isomers. On the other hand, NP is produced by the reaction of phenol with branched nonene. Commercial synthesis results in a mixture of various branched NP isomers rather than a discrete chemical structure [3]. NP is the most commercially prevalent of the APs family, representing approximately 85% of the APs market. Long-chain APEs are prone to metabolize in the natural environment, leading to the formation of APEs with short ethoxylate chain length (typically 1–3), APs and carboxylated APEs, which are more toxic and persistent than their precursor [4]. The environmental fate of these metabolites is of significant interest because of their potential endocrine disrupting effects [5]. For instance, NP ethoxylates are easily transformed, under anaerobic conditions, to their main metabolite, NP which is approximately 10 times more toxic than its ethoxylate precursor [6] and is known to disrupt normal hormonal functioning in the body [7,8]. Of all APE metabolites, OP is the strongest xenoestrogen and can induce significant effects in fish at concentrations of 3 ppb [9]. On the other hand, NP is active at concentrations of 8.3 ppb [10]. Since 1998, 4-NP and 4-OP are listed by OSPAR as chemicals for priority action [11] and both APs are present in the Water Framework Directive (Directive 2000/60/EC) of substances and substance groups prioritized for action to achieve good quality surface water [12,13]. The aim of these regulations is to ensure that emissions and losses of these substances to surface waters must be reduced to zero within the next two decades. The fate of APs in different environmental media (surface waters, groundwater, air, sediments, soils and biota) is controlled predominantly by their physicochemical properties, which are summarized in Table 19.1 [14]. APs are generally solids, with the exception of the ortho-substituted APs and 4-dodecylphenol, which are liquids. APs have low to moderate water solubility, except for 2,3,6-trimethylphenol, which has a high water solubility. Vapor pressures range from low to moderate. They are expected to have low to moderate mobility in soil. Volatilization of APs from water and moist soils is considered low to moderate, based on their Henry’s constants. APs are expected to have a range of persistence from moderate to high, with persistence increasing with increasing degree of branching, and bioaccumulation potential ranges from low to high, since some compounds can be metabolized and excreted as glucuronides. APs are produced in high volumes and are basically used as intermediates for the chemical industry in the production of APEs, phenolic oximes (mineral ore extraction chemicals) and phenolic resins (plastics). In particular, 2-sec-butylphenol is used as a chemical intermediate in the synthesis of insecticides, herbicides and as a polymerization inhibitor. 2-tert-Butylphenol is used as a starting material for the synthesis of flavor and fragrance chemicals, antioxidants, insecticides, and phenolic resins. Para-substituted APs are used in fragrances,

580 Transformation Products of Emerging Contaminants in the Environment

demulsifiers, biocides, oil field chemicals, surfactants, tackifier resins, ink resins and stabilizers, lube oil additives, as a co-solvent, as well as intermediates in the synthesis of antioxidants, phenolic resins and oil additives. Di- and trisubstituted APs are used as antioxidants, stabilizers for polyols, PVC, polyurethane, adhesives, functional fluids and also as intermediates in the manufacture of phenolic antioxidants, UV stabilizers, surfactants, and fuel additives.

19.2 Transformation of Alkylphenols and Identification of Transformation Products APs can be transformed following different techniques, namely biodegradation, photodegradation and sonolysis, with biodegradation being the most widely used. These processes are generally studied under laboratory conditions to define kinetics and eventually to identify the TPs formed through gas chromatography (GC), liquid chromatography (LC) or high performance liquid chromatography (HPLC) coupled to mass spectrometry (MS).

19.2.1 Biodegradation Biodegradation is the transformation of a substance into new compounds through biochemical reactions or by the actions of microorganisms such as bacteria and is the dominant transformation process of APs in the environment. Biodegradation has been studied extensively using soils, sludge and bioreactors and different bacteria strains. Several strains of sphingomonad bacteria isolated from activated sludge, including Sphingomonas sp. strain TTNP3 [15,16] and Sphnigobium xenophagum Bayram [17–19], can degrade branched 4-NP and utilize it as a sole carbon source. Both strains metabolize various NP isomers by type II ipso substitution. Growth on NP appears to be limited to those isomers that contain fully substituted alpha carbons on the alkyl side chain. Different biodegradation procedures and analytical methods have been used to identify the TPs of 4-NP (Table 19.2). Tanghe et al. [15] used a Sphingomonas sp. strain diethyl ether extract for injection in the GC-MS with an ion trap mass analyzer. 2,4,4-Trimethyl-2-pentanol was identified as a TP of OP. Grabriel et al. [19] used HPLC-MS and tandem mass spectrometry (MS/MS) in the negative mode using a turbo ion spray ionization to obtain structural information about the UV-active NP metabolites, studying their fragments. 2-Methyl-octan-2-ol, 3,6-dimethyl-heptan-3-ol and 2,3,5-trimethyl-hexan-2-ol were identified by GC-MS using a quadrupole mass analyzer (Fisons Instruments) [18]. Sphingomonas sp. (S-strain) was also found to be efficient in the degradation of NP [23]. In this case, TPs were analyzed by GC-MS in electron ionization. More than 95% of the NP was degraded within 10 d and aromatic compounds other than NP were not identified, suggesting that the phenolic part of NP was completely degraded. The potential of S-strain for bioremedial applications was also examined. S-strain cells immobilized on chitosan or alginate beads retained their NP-degrading activity in flask-scale experiments. Furthermore, the chitosan-bound cells in a lab-scale bioreactor were found to be persistent for repeated use, suggesting that S-strain could be applicable to the treatment of NP-contaminated wastewater. Other NP-degrading bacteria are Pseudomonas sp. strain JC1 [20] and Bacillus sphaericus strain CT7 [24]. In both cases the degradation was followed by GC-MS with electron

Transformation Products of Alkylphenols 583

ionization and an ion trap mass analyzer and the identified intermediate was 40 -aminoacetophenone. Other strains of sphingomonas bacteria can degrade OPs. Namely, Sphingomonas sp. strain PWE1, isolated from activated sludge, was found to be able to transform OP to hydroquinone [25], according to GC-MS characterization. Several Pseudomonas strains can degrade medium-length n-APs and utilize it as a sole carbon source. They include Pseudomonas sp. strain HBP1 [21], Pseudomonas sp. strain MS-1 [12], Pseudomonas putida MT4 [26] and Pseudomonas veronii strain INA06 [27]. Pseudomonas sp. strain HBP1, isolated from municipal sewage sludge, was found to be able to grow on 2-sec-butylphenol as the sole carbon and energy source [21]. Strain HBP1 could degrade 2.0 mM 2-sec-butylphenol within 40 h and 3-sec-butylcatechol, 2-hydroxy-6-oxo-7methylnona-2,4-dienoic acid and 2-methylbutyric acid were found as intermediates. These metabolites were analyzed by GC with a flame ionization detector (FID) and GC-EI-MS with electronic ionization and chemical ionization with methane as the reagent gas and ion trap mass analyzer. Pseudomonas sp. strain MS-1, isolated from freshwater sediment, was another bacterium that could use 2-sec-butylphenol as the sole carbon and energy source [12]. Within 30 h, this bacterium could completely degrade 1.5 mM 2-sec-butylphenol in basal salt medium, with concomitant cell growth. 3-sec-butylcatechol, 2-hydroxy-6-oxo-7-methylnona-2,4-dienoic acid and 2-methylbutyric acid were found as intermediates. These metabolites were analyzed by GC-MS. The collected culture was acidified with 1 N HCl, shaken for 3 min with an equal volume of 1:1 (v:v) ethyl acetate:n-hexane, and centrifuged (9600 g, 4
C, 10 min); the organic layer was then extracted. The extract was dried under a nitrogen flow and subjected to TMS using a BSTFA–acetonitrile solution at 60
C for 1 h. Pseudomonas putida MT4, isolated from a mixture of activated sludge samples from four sewage treatments plants, was able to degrade 1.0 mM 4-n-butylphenol in 20 h [26]. The TP 4-n-butylcatechol was detected by GC-MS. Studies showed that the recombinant strain degraded 4-n-APs with an alkyl chain of C7, although the wild-type MT4 was unable to assimilate 4-n-APs with an alkyl chain of C5. The wide range of NP isomers with a highly branched alkyl chain did not overlap with those of known NP-degrading Sphingomonas strains, which could degrade NP isomers with a highly branched alkyl chain preferentially, but not APs with a medium-length alkyl chain [16,18,23]. Therefore, MT4 could contribute to the total removal of APs in soils and aquatic environments in combination with such Sphingomonas strains. Pseudomonas veronii strain INA06 isolated from activated sludge, was found to degrade 4-n-amylphenol and 4-n-hexylphenol as the sole source of carbon [27]. It could also cometabolically degrade linear 4-NP. This bacterium was not able to degrade 4-n-heptylphenol or larger chain 4-n-APs as the sole source of carbon. Sphingobium fuliginis strains [13] and Spirodela polyrrhiza [28] strains are the only two microbes known to grow on 4-tert-butylphenol as a sole carbon and energy source. Sphingobium fuliginis strain TIK-1, isolated from the rhizosphere sediment of the emergent aquatic plant Phragmites austrils, was able to completely degrade 1.0 mM 4-tertbutylphenol in a basal salts medium within 12 h [13]. 4-tert-butylcatechol and 3,3dimethyl-2-butanone were identified as internal metabolites. Actually, studies showed that strain TIK-1 cell could degrade 4-APs with various sized and branched side chains (ethyl, n-propyl, isopropyl, n-butyl, sec-butyl, tert-butyl, n-pentyl, tert-pentyl, n-hexyl, n-heptyl, n-octyl, tert-octyl, n-nonyl and branched nonyl) via a meta-cleavage pathway,

584 Transformation Products of Emerging Contaminants in the Environment

but not 2- or 3-APs. The metabolites of 4-tert-butylphenol and other APs were analyzed by GC-MS, as described in [12]. Spirondela polyrrhiza strain OMI, isolated from the rhizosphere Spirodela polyrrhiza, was also able to degrade 4-tert-butylphenol and almost the same range of APs as Sphingobium fuliginis strain TIK-1 [28]. 4-tert-Butylphenol metabolites were analyzed by GC-MS. The culture sampled at each sampling time was acidified with 2 N HCl to pH 2–3, shaken for 3 min with an ethyl acetate: n-hexane mixture (2:1, v/v) and centrifuged (378  g, 10 min) and the organic layer was collected. For trimethylsilylated (TMS) derivatization, the extract (100 mL) was dried under nitrogen flow and then TMS at 60
C for 1 h using 100 mL of BSTFA–acetonitrile solution (1:1, v/v). GC-MS analysis with two oven programs was used to analyze the metabolites with and without TMS derivatization. Seven metabolites were identified and three structures were determined: 4-tert-butylcatechol, 3,3-dimethyl-2-butanone and 2,2-dimethyl-propionic acid. Biodegradation of 2,6-di-tert-butylphenol was accomplished using two effective bacterium strains, F-1-4 and F-3-4, which were isolated from an acrylic fiber wastewater [29,30]. Strains were immobilized in calcium alginate gel through different immobilization methods, namely immobilization of single, homogeneously mixed, and stratified strains. Results indicated that the stratified immobilization yielded the highest degradation efficiency and 91% of 2,6-di-tert-butylphenol could be biodegraded after 12 days of incubation. Transformation of 4-n-OP and 4-n-NP by ammonia-oxidizing bacteria Nitrosomonas europea ATCC 19 718 was investigated by Sun et al. [31]. TPs of the reaction between nitrite and 4-n-OP and 4-n-NP were nitro-n-OP and nitro-n-NP, respectively, which had less estrogenicity than their parent compounds. All these compounds were identified by UPLC and a quadrupole-time of flight (Q-TOF) mass spectrometer. There are few studies on the biodegradation of APs under anaerobic conditions. Shibata et al. [32] reported that 4-n-propylphenol was degraded anaerobically in a paddy soil, whereas APs with long and branched alkyl chains (4-n-butylphenol, 4-secbutylphenol, 4-tert-butylphenol and 4-tert-OP) were poorly degraded over 224 days of incubation. Later the same research group demonstrated the anaerobic biodegradation of 4-n-butylphenol and 4-sec-butylphenol in a paddy soil supplemented with nitrate [33]. This anaerobic microbial degradation occurred via oxidation of the alpha carbon in the alkyl chain. Although the supplementation of nitrate enhanced dramatically the degradation of APs in the paddy soils, aerobic degradation of 4-tert-butylphenol, 4-tert-OP and 4-n-OP was still not observed. In both studies, the microorganism involved in the anaerobic degradation was not elucidated. For this reason, Shibata and Katayama [22] isolated and characterized the microorganism responsible for the biodegradation of 4APs under nitrate-reducing conditions. This microorganism was the anaerobic bacterium Thauera sp. strain R5, which could co-metabolically transform 4-APs to the corresponding metabolites (ketone, alkene and alcohol derivatives) with an oxidized alpha carbon in the alkyl chain, accompanied by nitrate reduction. In the presence of 4hydroxybenzoate as a carbon source, the possible anaerobic metabolic pathways of 4-nbutylphenol by strain R5 were proposed as follows: dehydrogenation of 4-n-butylphenol to 1-(4-hydroxyphenyl)-1-butene; hydroxylation to 1-(4-hydroxyphenyl)-1-butanol; and then dehydrogenation to 40 -hydroxybutyrophenone. In the degradation of 4-APs with linear and medium length alkyl chains (C6), 1-(4-hydroxyphenyl)alkanones were observed. 1-(4-hydroxyphenyl)alkenes were detected in the degradation of linear APs with medium or long chain lengths of up to eight plus the secondary alkyl chain. These

Transformation Products of Alkylphenols 585

results suggested that the degradability of 4-APs by strain R5 might decrease with increasing chain length and bulky size. GC-MS was used to detect the four metabolites with retention times of 8.5, 9.4, 10.4 and 10.7 min and identified as 4-isopropenylphenol, 4-(1-butenyl)phenol, 40 -hydroxypropiophenone and 40 -hydroxybutyrophenone by comparison of their spectra with those of authentic standards, the fragment patterns and the molecular ion peaks. Some research groups have studied the degradation of NP in soils amended with sewage sludge. Hesselsoe et al. [34] studied the degradation of 4-NP in homogenous and nonhomogeneous mixtures of soil and sewage sludge and demonstrated that degradation was dependent on oxygen availability. The mineralization of 14 C-labeled NP in both homogenized and nonhomogenized sludge–soil mixtures was investigated and the results showed that NP was degraded within 38 d in aerobic homogenized mixtures, whereas in nonhomogenous mixtures containing sludge aggregates its degradation was generally not completed within 3 months. Mortensen and Kure [35] studied the degradation of NP in spiked soils and in soils treated with organic waste products, including both anaerobic and aerobic sludge, compost and pig manure. They showed that more complete degradation of NP was observed in sludge amended soil compared with compost-treated soil. They also studied the influence of plant growth on the degradation and demonstrated that when NP was added as waste to the soil, plant growth significantly stimulated the degradation. Mineralization of 14 C-labeled NP was also used in wheat cell cultures [36]. After incubating 1 ppm NP in the cell culture, more than 90% of the applied radioactivity could be extracted from the cells. Hydrolysis of metabolites resulted in two major TPs which were less polar than the parent metabolite substance but more polar than 4-NP. Both the metabolites and the TPs of the hydrolysis were analyzed by GC-EI-MS using a double-focusing Kratos MS 50 fast scan mass spectrometer. 4-n-NP derivatives were dissolved in methanol containing 0.1% ammonia and injected in a LC-MS/MS with electrospray. The identified TPs were glucose, glucuronic acid, aglycon, 1,5-diacetyl-2,3,4,5-tetramethylglucitol, and 1,2,3-triacetyl-3,4,6-trimethylglucitol. The same degradation method was used to study the biotransformation of 4-tert-OP in isolated rat hepatocytes [37]. After a 15 min incubation of hepatocytes with 30 mM 14 C-labeled 4-tert-OP, over 97% of the 4-tert-OP was metabolized to a complex mixture of metabolites, which were detected by radio-HPLC (Figure 19.2a) and identified by GC-MS (Figure 19.2b). The initially formed metabolites were identified as products of hydroxylation of the aromatic ring to form catechols and methylated catechols, as well as glucuronide conjugates of the catechol metabolites or parent phenol. These TPs were further metabolized by hydroxylation of the alkyl chain, followed by glucuronide conjugation of the alkoxy group. The identified TPs were 2-hydroxy-4-(10 ,10 ,30 ,30 -tetramethylbutyl)phenol (I), 4-(10 ,10 ,30 ,30 -tetramethylbutyl)phenoxy-b-glucuronide (II), 2methoxy-4-(10 ,10 ,30 ,30 -tetramethylbutyl) phenoxy-b-glucuronide (III), 2-hydroxy-4(10 ,10 ,30 ,30 -tetramethylbutyl)phenoxy-b-glucuronide (IV), 2-hydroxy-5-(10 ,10 ,30 ,30 -tetramethylbutyl) phenoxy-b-glucuronide (V), 2-methoxy-4-(20 -b-glucuronoxy-10 ,10 ,30 ,30 -tetramethylbutyl)phenol (VI), 4-(20 -b-glucuronoxy-10 ,10 ,30 ,30 -tetramethylbutyl) phenol (VII), 2-methoxy-4-(2 0 -hydroxy-1 0 ,1 0 ,3 0 ,3 0 -tetramethylbutyl)phenoxy-b-glucuronide (VIII), 4-(20 -hydroxy-10 ,10 ,30 ,30 -tetramethylbutyl) phenoxy-b-glucuronide (IX) and 2b-glucuronoxy-4-(20 -hydroxy-10 ,10 ,30 ,30 -tetramethylbutyl)phenol (X). The proposed pathways for the degradation of tert-OP in rat hepatocytes as well as the structures of the metabolites are shown in Figure 19.2c.

586 Transformation Products of Emerging Contaminants in the Environment (a)

(b) V

1.4

Relative abundance (%) 100

Radioactivity (kCPM)

1.2

237 MeO

237

90

TMSO

80

t-OP

1.0

x10 308

70

308 amu

60

0.8

50 0.6

40

IV b

0.4

II

VII X IX VI VIII

0.2

III

73

293

30 20

I

10 0

0

50 0

100

150

200

250

300

350

m/z

10 20 30 40 50 60 70 80 90 100 110 120 130 140 150 Retention time (min)

(c) HO t-OP

HO GlcUAO

OH

HO II

HO

a

OH

OH

HO

GlcUAO I

HO GlcUAO

HO

IV

MeO

GlcUAO

HO

HO

IX

V

MeO OGlcUA

HO

HO

VII

a

OH

MeO

GlcUAO

GlcUAO

HO III

OH X

MeO

MeO HO

GlcUAO VI

OH

GlcUAO VIII

Figure 19.2 (a) Radio-HPLC profile of radioactive metabolites. Reprinted with pemission from [37] Copyright (2000) Elsevier Ltd. (b) GC-MS spectra of the trimethylsilyl ether of the agliycone of metabolite III, Reprinted with pemission from [37] Copyright (2000) Elsevier Ltd. (c) Proposed pathways of tert-OP in rat hepatocytes.

Transformation Products of Alkylphenols 587

19.2.2 Photodegradation The photodegradation process is gaining importance in the study of the fate of APs in the environment. Different photodegradation techniques have been used to degrade APs: (i) photocatalysis, (ii) photochemical degradation, and (iii) direct photolysis. The main oxidizing reagents involved in these methods are ozone, oxygen, and hydrogen peroxide (H2O2). These reagents, and sometimes water, can produce highly reactive species, such as hydroxyl radicals (OH) and singlet oxygen (1 O2 ), under appropriate conditions. The main reducing reagents involved are electrons or zero valent metals, such as iron. These species can be introduced by addition or by in situ reactions. In photodegradation processes, UV radiation is used as the energy source to directly decompose ozone and hydrogen peroxide or, through a photosensitizer, to degrade oxygen or water to create reactive radicals. UV radiation itself can sometimes be used to directly photolyze pollutants. Combinations of UV and oxidizing reagents are flexible. Photocatalysis is the main photodegradation technique used to degrade APs. The main factors that influence the photocatalytic degradation of APs are the presence of photosensitizers and the light intensity. A good photosensitizer should be chemically and biologically inert, photocatalytically active, easy to produce and use, and activated by sunlight. Semiconductors (e.g., TiO2, ZnO, Fe2O3, CdS and ZnS) can act as photosensitizers for light-reduced redox processes due to their electronic structure, which is characterized by a filled valence band and an empty conductance band [38]. When a photon with energy of hn matches or exceeds the bandgap energy, Eg, of the semiconductor, an electron, ecb  , is promoted from the valence band into the conductance band, leaving a hole, hvb þ , behind. This positive hole may react directly with the surface-sorbed organic molecules or indirectly oxidize organic compounds via formation of OH radicals, which are the most potent oxidizing agents. Among the different photosensitizers, TiO2 is considered the most efficient semiconductor for environmental applications. It has some significant advantages over other semiconductors, such as its low cost, its high chemical stability and its capacity to generate highly oxidizing holes and electrons. The generated holes can produce OH and superoxide (O2) radicals, which are strong oxidizing agents. Therefore, they play an important role in the photocatalytic degradation processes: TiO2 þ hy ! TiO2 ðhybþ þ ecb Þ 

hybþ þ H2 OðadsÞ ! OHðadsÞ þ Hþ 

(19.1) (19.2)

hybþ þ OH ðadsÞ ! OHðadsÞ

(19.3)

O2 þ ecb ! O 2

(19.4) 

þ (19.5) O 2 þ 2H $ 2 HO2 Then these active species oxidize organic compounds present in solution, eventually mineralizing them to CO2, H2O and simple inorganic ions. The reaction describing the process for the implementation of this semiconductor in the degradation of APs is given by the following equation: TiO2 ;hyEg

AP þ O2 ! CO2 þ H2 O þ inorganic matter

(19.6)

One of the major products of this reaction is CO2, as the result of oxidation of organic compounds.

588 Transformation Products of Emerging Contaminants in the Environment

Tanizaki et al. [39] used a porous and nanostructured TiO2 immobilized on a photoreactor to photocatalytically degrade amyl-, hexyl-, heptyl-, octyl-, 4-tert-octyl- and NP. With the photocatalyst, about 90% of their initial concentration was decomposed within 60 min of UV irradiation. They found that the longer the alkyl chain length of the APs, the faster the decomposition rate. They also showed that the decomposition rate of straight-chained APs was faster than that of branched ones. TiO2/UV was also used to investigate the photocatalytic destruction of 2,4,6-trimethlyphenol in aqueous suspensions [40]. The authors found that the more diluted the initial concentration of the AP, the faster the degradation. Davezza et al. [41] applied TiO2/UV photocatalysis to treat soil washing extracts containing 4-tert-butylphenol in the presence of non-ionic (C12E8, C12E23 and Brij 35) and anionic (SDS) surfactants and some of their binary mixtures. It was proven that Brij 35 allowed faster photocatalytic treatment of the soil washing wastes. An important disadvantage of TiO2 is that it only works under UV light irradiation (l 380 nm) and this wavelength only corresponds to 8% of the sunlight energy. Furthermore, TiO2 is usually used as a powder in aqueous suspensions and the recovery of the material from the solution at the end of the process is very difficult and expensive. In order to increase the efficiency of the photocatalytic process in TiO2, reduced forms of TiO2, such as TiOx (x < 2), have been used [42]. NP was completely degraded in 30 min under visible light by using a 350 mm thick layer of TiO1.8 grown on a soda lime substrate. Ardizzone et al. [43] combined TiO2/UV photocatalysis and ozonation to successfully degrade 4-cumylphenol to CO2. 4-Cumylphenol (0.2 mM) was completely mineralized after 90 min of the combined treatment. The success of the process was largely due to the presence of sandwiched TiO2 layers deposited onto an Al lamina. The stable and purposely rough TiO2 surface offered a large number of adsorption sites for both the pollutant and ozone, thus promoting synergistic, highly oxidizing reaction steps. The possible formation of ortho/ para bisphenols as oxidation intermediates was suggested as the main starting step leading to ring opening intermediates and finally to CO2. Wu et al. [44] developed a spiral photoreactor system for the efficient degradation of 4-tert-OP in a water solution with TiO2 used as photocatalyst. A single layer TiO2 film prepared with 13.6% of TiO2 precursor was proven to be the optimal photocatalytic reaction condition, under which 90% of 4-tert-OP could be degraded in 30 min when the initial concentration was 10 mg/L. The intermediate products of 4-tert-OP degradation were probably some hydroxy group containing substances which could be decomposed further in the reaction sequences. There were 3 TPs identified with by LC-MS: 4-(2,4,4-trimethylpentan2-yl)benzene-1,2-diol, 4-(3-hydroxy-2,4,4-trimethylpentan-2-yl)phenol and 4-(2,3,3-trimethylphentan-2-yl)benzene-1,3-diol. BiVO4 was also reported as a photocatalyst to degrade 4-n-NP under irradiation from a solar simulator [45]. The photocatalytic activity of BiVO4 to degrade 4-n-NP was comparable to that of TiO2 in an air saturated solution and, further, became greater in an O2 saturated solution under solar light. However, the CO2 mineralization yield was quite low since the photooxidation power was not sufficient to oxidize intermediate products. The same photocatalyst was applied to the degradation of propyl-, butyl-, pentyl-, hexyl-, heptyl-, octyl- and NP under solar light [46]. Degradation rates became faster with increasing alkyl chain length. The amount of adsorption on the BiVO4 surface was much larger for longer hydrophobic APs, showing the suitability of BiVO4 to degrade the endocrine disruptors OP and NP. BiVO4 was shown to be also effective for the degradation of branched 4-NP. The

Transformation Products of Alkylphenols 589

analyses of TPs were carried out on a GC-MS instrument (Hewlett-Packard 6890/5973) equipped with a chiral capillary column. Later the same research group used silver fineparticles loaded on BiVO4 to efficiently degrade 4-n-NP and 4-n-OP under visible light irradiation [47]. By using silver as a co-catalyst, the electron–hole charge separation at the photocatalyst/co-catalyst interface became very efficient, improving the photocatalytic activity. Furthermore, the metallic silver particles contributed to an increase in the reaction rates and CO2 mineralization yields for the degradation of NP. A common TP was identified as cis, cis-4-alkyl-6-oxo-2,6-hexadienoic acid. Babaei et al. [48] described the modeling of NP degradation by the UV–vis/ZnO process. In this regard, a full factorial experimental design was performed to study the main variables affecting the degradation process. ZnO was proposed as an alternative to TiO2 since the pathway of OH radical generation is similar to that of UV/TiO2. Thus, it can be described by Equations 19.7 and 19.2–19.5: ZnO þ hy ! ZnOðhybþ þ ecb Þ

(19.7)

The main advantage of ZnO in comparison to TiO2 is that it absorbs over a larger fraction of the UV spectrum and the corresponding threshold of ZnO is 425 nm. They concluded that the degradation rate constant decreased with an increase in the initial concentration of NP while it increased with ZnO loading until a concentration of 0.5 g/L. The degradation rate constant also increased with the pH up to pH10, after which a significant decrease was observed. Trivalent ion was also chosen as a photosensitizer because Fe(III)-aquo complexes are very efficient at forming hydroxyl radicals. Fe(III)-aquo complexes are known to undergo photolysis upon irradiation with wavelength l > 300 nm, yielding Fe(II) and OH radicals. Among Fe(III)-aquo complexes, Fe(OH)2þ is photolyzed with the highest quantum yield [49], according to the following reaction. hy

FeðOHÞ2þ ! Fe2þ þ OH

(19.8)

Brand et al. [50] investigated the degradation of 4-OP photoinduced by Fe(III) in water/ acetonitrile solution (95 : 5). In this case, OH radicals did not react with the AP, but with acetonitrile, leading to acetonitrile-derived radicals. The photoproducts were identified by LC-ESI-MS. The first photoproduct was confirmed with m/z 222 for the molecular peak and m/z 151 for the major fragment, the 4-(1,1,3,3-tetramethylbutyl)pyrocatechol. The second and the third photoproducts were benzoquinone and 4-hydroxyacetophenone and were identified by comparison with the authentic standards. The photoproducts IV–VII came from the oxidation on the different sites of 4-OP alkyl chain. Dimeric photoproducts were not identified. Direct photolysis using UV-solar simulating light was also employed to degrade APs. Neamtu and Frimmel [51] showed that the degradation of OP in an aqueous solution by using direct photolysis was much slower than when the photodegradation took place in the presence of H2O2, due to the generation of OH radicals through the H2O2 photolysis. Actually, after 8 h of irradiation, the estrogenic activity of OP remained approximately constant for experiments conducted using the UV solar simulator, whereas it practically disappeared at the beginning of the reaction when 50 mM H2O2 was added. They also studied the effect

590 Transformation Products of Emerging Contaminants in the Environment

of some common water constituents, such as bicarbonate, nitrate, Fe(III) ions and dissolved natural organic matter (DNOM) on the photodegradation of OP. The results confirmed that the degradation of OP was also increased in the presence of NO3 and DNOM and was decreased in the presence of HCO3. On one hand, NO3 is able to form OH radicals, therefore the degradation rate of photolysis increases. On the other hand, the light absortion by DNOM led to the formation of reactive species, such as 1 O2 , O2, OH and peroxyl (ROO) radicals, which reacted with APs. Phenol, 1,4-dihydroxybenzene and 1,4-benzoquinone were found as the dominant intermediate products. Similar results were obtained by Xu et al. [52]. They found that the half-life of OP under simulated solar irradiation conditions ranged from 0.6 to 2.5 days, depending on the initial concentration of the precursors. They also studied the influence of humic acid (HAC), one of the most important fractions of humic substances, which represent the most widespread pool of natural recalcitrant organic carbon occurring in the biosphere, constituting the most stable fraction of organic matter in soils, sediment and waters. HAC produces and competes for OH radicals, absorbs light and can inhibit photolysis by binding APs [53]. Bledzka et al. [54] also used direct photolysis to degrade OP. The results obtained with this method were compared to those obtained with the advanced oxidation process (H2O2/ UV) and those obtained with photosensitized oxidation using 1 O2 . Although efficient degradation was found with photolysis at 254 nm, the rate of the process was very slow. The H2O2/ UV system was found to be the most efficient method to degrade OP. The degradation of NP was also evaluated using simulated solar UV-irradiation in the absence and presence of DNOM, HCO3, NO3 and Fe(III) ions [51]. The results indicated that the oxidation rate increased in the presence of H2O2, Fe(III) and DNOM. In this case, the estrogenic activity of NP aqueous solution after 193 h of irradiation in the absence of initial H2O2 was comparable with that after 7 h of irradiation for experiments in the presence of 50 mM H2O2. Phenol, 1,4-dihydroxybenzene and 1,4-benzoquinone were also found as the dominant TP. 19.2.3 Sonolysis Recently, there has been interest in the use of ultrasound to decompose organic compounds. The chemical effects of ultrasound derive from acoustic cavitation: the formation, growth and implosive collapse of bubbles in a liquid. During the adiabatic compression stage of oscillating or collapsing bubbles, high temperatures of several thousand K and pressures of several hundred bars are reached in a short period of time. Water molecules under such extreme conditions undergo thermal dissociation to yield H atoms or OH radicals: H2 O ! H þ OH

(19.9)

O2  ! 2 H

(19.10)

H þ O2 ! OOH

(19.11)

O þ H2 O ! 2 OH

(19.12)



H þ O2 ! OH þ O

(19.13)

Organic solutes in the vicinity of a collapsing bubble or partitioned into the gas phase of the bubble undergo thermal decomposition and/or react with the reactive radicals.

Transformation Products of Alkylphenols 591

The sonolytic degradation of 4-cumylphenol in aqueous solution was investigated by Chiha et al. [55]. First, they studied the influence of operating parameters for the sonication process. They found that the degradation rate increased proportionally with increasing ultrasonic power from 20 to 100 W and temperature in the range 20–50
C. The most favorable conditions for the degradation were observed in acidic media. The dependence of the 4cumylphenol degradation rate on the presence of various saturating gases showed the following order: argon > air > nitrogen. Actually, the degradation was inhibited in the presence of nitrogen gas due to the free radical scavenging effect in the vapor phase within the cavitation bubbles. The ultrasonic degradation of 4-cumylphenol was clearly promoted in the presence of bromide and bicarbonate ions. Finally, experiments conducted using pure and natural water demonstrated that the sonolytic treatment was more efficient in the natural water compared to pure water. Sonolysis of butyl-, pentyl-, octyl- and NP in water was studied at a sound frequency of 200 Hz with an acoustic intensity of 6 W cm2 under argon, oxygen and air atmospheres [56]. All the APs were effectively degraded, although their sonolytic degradation rates were dependent on the length of the alkyl chain. Over 90% degradation of APs occurred within 30 min sonication, except for NP (within 100 min sonication). The addition of Fe(II) and Fe (III) as catalysts during sonication resulted in a remarkable enhancement of degradation. On the other hand, sonolysis was also used to accelerate enzymatic hydrolysis of APs in fish bile [57]. Commonly the hydrolysis time was reduced from 16 h in sonolysis to 20 min in enzymatic hydrolysis.

19.3 Occurrence of Alkylphenol Transformation Products in the Environment Phenolic compounds in the environment can arise from the degradation of natural substances as well as from anthropogenic activities, such as human settlements, industrial and agricultural activities. APs have been detected in water, sediments, soil, biota and air. The partitioning, transport and fate of APs in the environment depend on the physicochemical properties of each compound. Processes such as sorption, volatilization, degradation and metabolization are main factors affecting the fate of APs in the environment. The presence of both APs and their TPs is responsible for the toxicological effects toward aquatic and terrestrial ecosystems. These are directly related to the concentration of each compound in each environmental matrix and to its specific toxicity. Therefore, it is essential to determine the partitioning of APs in the ecosystem, the interactions between the different environmental compartments, and the degradability of each compound. This information, considered as a whole, will permit better evaluation of the most dangerous compounds and, therefore, allow appropriate actions for their containment to be taken. 19.3.1 Water Water is the main vehicle of transport of APs in the environment. APs have been detected worldwide in all types of waters, including surface, coastal, groundwater, rainwater and lakes. Site specific industrial/urban and agricultural use of AP-containing products contribute to the dispersion of APs. Because APs are high production volume chemicals with high detrimental impact, they are considered as a main group of contaminants for which actions

592 Transformation Products of Emerging Contaminants in the Environment

have to be undertaken to minimize their environmental presence. Among APs, most of the studies have been performed on OP and NP, because of their high use, accumulation [58], estrogenic effects [9] and, since 1998, because they are legislated. Whereas APs are distributed worldwide, the concentration detected in waters is geographically dependent, and legislation actions, water policy and water treatment facilities play an important role in the residues detected at each site. Furthermore, the monitoring of APs is basically restricted to OP and NP, and very little information is available on the fate of these common APs in the water ecosystem. APs were monitored in rivers and estuaries in England and Wales during 1993 [59]. Dissolved NP (and total extractable NP) ranged from 0.2 (0.5) to 53 (180) mg/L, the highest levels being in the river Aire. Concentrations of NP were one order of magnitude lower in estuaries than in rivers, reflecting dispersion and dilution processes. Out of four sampled estuaries, both NP and OP were detected only in the Tees estuary at 5.2 and 13 mg/L, respectively [59]. Matthiessen et al. [60] reviewed the biological effects of contaminants in British estuaries and coastal waters for the last 100 years. Until 1970, the major pollution impacts on estuarine organisms were probably attributed to poorly treated sewage and oxygen depletion whereas, since the installation of wastewater treatment plants (WWTPs), the release of toxic micropollutants became a serious environmental problem and NP was identified in the Tyne and Tees estuaries as contributing to toxic effects in the copepod Tisbe battagliai. In Italy, the nature, origin and trend of phenolic compounds in the river Po, whose water is used for drinking purposes, was evaluated. Among several methyl and methylethyl phenols, NP isomers showed a decrease in concentration from 158 mg/L in 1994 to 7.5 mg/L in 1996, although the frequency of variation was invariable. This decrease in concentration was explained by the substitution of this surface active agent with new products, such as alcohol polyethoxylates. In France, four rivers representing rural, agricultural, urban and industrial watersheds were monitored in 1999–2000 to determine the presence of APs in water and sediment. NP was detected from 0.006 to 0.55 mg/L and OP from 0.001 to 0.0077 mg/L, and the highest levels were found in urban and industrial sites. Sediment concentrations varied from 0.0022 to 2.87 mg/g for NP and from 0.001 to 0.491 mg/g for OP, the highest concentration being in urban sites. In some sites, estrogenic activity was induced at those concentrations and, thus, could produce toxic effects toward aquatic organisms [61]. Also in France, NP and OP, together with their mono-, di- and tri-ethoxylated compounds, were analyzed in surface and groundwater in the Rh^one-Alpes region in March–April 2007. With solid phase extraction (SPE) and LC-MS/MS, only NP mono and di-ethoxylated and OP di-ethoxylated were detected at 0.07–1.41 mg/L, whereas NP and OP were not identified [62]. In Japan, Tsuda et al. [63] monitored the presence of 4-NP and 4-tert-OP in water and fish from eight rivers flowing into Lake Biwa. NP was detected in waters at 0.11– 3.08 mg/L all year round and in all 48 analyzed samples, whereas 4-tert-OP was detected at lower levels (nd–0.09 mg/L) and at lower frequency (23 out of 48 samples). NP was detected in fish at levels up to 63 ng/g ww and OP at 6 ng/g ww. For NP, bioconcentration factors were lower in field fish than in laboratory experiments, whereas for 4-tert-OP, these values were maintained. In another study, stir bar sorptive extraction-thermal desorption-GC-MS (SBSETD-GC-MS) with in situ and in tube silylation was used to determine the presence of APs in Tama river water, Japan. The method was optimized to determine 4-tert-buthylphenol, 4-npentylphenol, 4-n-hexylphenol, 4-n-heptylphenol, 4-tert-OP, 4-n-NP and 4-NP. The most ubiquitous detected compound was NP, identified upstream, midstream and downstream of the river at constant concentrations of 0.055–0.059 mg/L while tert-OP was detected from 0.012 mg/L in upstream waters and 0.018 mg/L in downstream waters. The other compounds

Transformation Products of Alkylphenols 593

were not detected above the calculated LOD of 0.0002–0.005 mg/L [64]. In the river HsinDian Creek, in Taiwan, 4-tert-OP was not identified using HPLC-MS/MS in water, while in sediments and fish mean concentrations were 287 and 36.3 ng/g ww. On the other hand, NP was detected in water at 1.026 mg/L and at 817 and 238 ng/g ww in sediments and fish, and was the most ubiquitous detected compound [65]. In China, the behavior of APs was studied in the Jialu river, Henan Province, which annually receives 726 kg of NP and 30.2 kg of OP from the urban discharges from the city of Zhengzhou [30]. The concentration of NP ranged from 0.075 to 1.52 mg/L and that of OP, from 0.021 to 0.063 mg/L. Dilution and biodegradation accounted for the main decline in AP concentrations along the river, although adsorption and air–water exchange also contributed to the elimination of APs. The decay half-lives for NP and OP were 1.6 and 2.4 d, respectively, and about 70.2% of total NP and 24.1% of total OP were finally eliminated from the water phase in a downstream direction, indicating that the downstream channel served as a sink for APs in the study area [30]. In Rumania, industrial wastewaters dealing with paper, detergent, electricity and heat, refinery, wood and sewage of different cities, contained 4-tert-OP from 0.011 to 0.069 mg/L, 4-OP from 0.018 to 0.045 mg/L and 4-NP from 0.064 to 0.698 mg/L, and river concentrations were of the same order, 0.076 and 0.156 mg/L for 4-tert-OP, from 0.024 to 0.091 mg/L for 4-OP and from 0.343 to 0.4.4 mg/L for 4-NP [66]. In that study, APs were analyzed by HPLC using a mixed-mode stationary phase column (C18/SCX). In the Polish river waters monitored in 2009, NP was always detected from 0.18 to 0.53 mg/L while OP was only detected in 4 out of 11 rivers at limit of quantification (LOQ) 0.04 mg/L. In this study, a SPE method using polytetrafluoroethylene/teflon (PTFE) microspheres as sorbent, followed by HPLC-fluorescent detection was developed, obtaining very good recoveries and sensitivity for the low concentration analysis of APs in water [67]. In Japanese rivers, 4-tert-buthylphenol, 4-n-pentylphenol, 4-n-hexylphenol, 4-n-heptylphenol, 4-OP and 4-NP were analyzed by HPLC with coulimetric-array detection, reporting high efficiency and selectivity [68]. NP and OP were detected in drinking water at 0.04 and 0.05 mg/L, while in river waters, buthylphenol, penthylphenol, hepthylphenol, OP and NP were detected at 0.03–0.17 mg/L. In Norway, Boitsov et al. [69] analyzed produced water from nine oil installation plants. SPE and GC-MS were used to quantify 49 known APs and P another set Pof 35 peaks corresponded to “non-identified” AP peaks. The concentration of C2 APs to C9 APs ranged from 0.001 to 1098 mg/L and the most ubiquitous compounds were C2 APs, such as dimethylphenol and ethylphenol. Lower concentrations were found with increasing alkyl chain length, and NP was the least detected compound. The levels of low-weight APs varied greatly among locations but only moderate differences were observed in the levels of long-chained APs. In addition, unidentified APs were semi-quantified at levels from 0.021 to 6.58 mg/L [69]. The study reports that the release of these compounds, especially long-chain para-substituted, can act as endocrine disruptors for marine biota. In the NE Spain, wastewater treatment plants (WWTPs) and river outflows were identified as a major source of organic micropollutants to coastal waters, with discharges of 8800 g/d and 17 030 g/d. Among other contaminants, OP, NP and NP monoethoxylated were identified as the major contaminants, with concentrations from 0.013 to 0.061 mg/L for OP and NP in river water and 0.395 and 1.021 mg/L for OP and NP in wastewater effluents, respectively [70]. In that study, water was analyzed unfiltered to capture the total concentration of APs (dissolved þ particulate). In the Thermaikos gulf, Northern Aegean Sea, Greece, NP, 4-tert-OP, n-OP and other endocrine disrupting chemicals (EDCs) were detected in seawater (1.7–201 ng/L), suspended

594 Transformation Products of Emerging Contaminants in the Environment

particulate matter (61–1578 ng/g dw) and in sediments (6–2695 ng/g), and mussels only accumulated NP at 27.3–79.4 ng/g dw. Rivers, raw wastewater from tanneries and sewage submarine emissaries discharging into the gulf were the main contributors of APs to the Thermaikos gulf. In addition, a high risk for sediment biota was observed for NP [71]. It is not surprising that 4-NP has also been detected in rain and snow in urban, suburban and rural areas in Germany and Belgium. In this study, 16 isomers of NP were quantified by GC-MS. The mean concentration of 4-NP in rain water and roof runoff was 0.253 mg/L, being the levels in suburban areas > rural > urban. In addition, lower levels were detected in summer rain than in winter rain. In snow, 4-NP was detected with a mean value of 0.242 mg/ L, with levels in urban areas > suburban, and not detected in rural areas. The presence of NP in rain and snow points to wet deposition as a source of 4-NP in the environment [72]. WWTPs receive high loads of APs through urban and industrial effluents, and runoff. Within the WWTP, APs are eliminated from water during primary and biological treatment, although if treatment is not fully effective they can be released by the effluent. In a typical biological treatment WWTP, the elimination efficiency is 70–90% [73], although a high amount can accumulate in sludge which, in turn, may contain high levels of APs [74]. Sludge is used in many countries as organic fertilizer and can be a source of APs to agricultural soils or groundwater. Monitoring of APs in WWTP can be performed in two ways. One way is to determine the concentration in the effluents, with the aim to determine the potential contribution of APs to received waters. Another way is to determine the mass balance of contaminants by monitoring the influent, the effluent and the sludge, to determine the efficiency of the WWTP and to define the partitioning and behavior of APs within the WWTP. These different approaches have been used depending on the objectives to be achieved. Blackburn et al. [59] determined NP at 0.1–5.4 mg/L in 15 English sewage effluents and indicated that these levels were low, presumably as a result of the UK voluntary bans on APEs in domestic detergents [59]. In a Geneva (Switzerland) municipal sewage wastewater receiving 130 000 m3/d [75], the total concentration of free 4-APs ranged from 1 to 6.8 mg/L, which was composed of 18 differently branched 4-NPs derivatives identified with both GC and LCMS, while 4-tert-OP, 2-NP, 2,4-dinonylphenols, various branched 4-OP and 4-decylphenols could not be detected in wastewater. In Spain, NP, NP monoethoxylate and NP diethoxylate were studied in a sewage treatment plant receiving 0.268 m3/s [76]. Using simultaneous steam distillation/solvent extraction and GC-MS and isotope dilution quantification, NP was detected at 57.6 mg/L in the influent, at 72.73 mg/L in the pre-treatment effluent, at 58.18 mg/ L in the primary effluent, and finally at 0.65 mg/L in the plant effluent, with a very high removal efficiency. In contrast, very low levels of 4-NP, 4-OP, 4-tert-OP were detected in urban wastewaters from Granada, Spain, where 4-NP was detected in only one sample at 0.1 mg/L while 4-OP was never detected using solid phase extraction followed by GC-MS with BSTFA/TMCS derivatization [77]. These concentrations are very low compared to wastewaters from more industrialized or urbanized areas. In a WWTP from Australia with a capacity of 5000 habitants-equivalent or 1300 m3/day, 4-tert-OP was detected at 0.0026– 0.0044 mg/L, NP at 0.070–0.120 mg/L and 4-cumylphenol at 0.0023 mg/L. Anaerobic sludge contained 10.9 ng/g 4-tert-OP, 50.8–406 ng/g NP and 0.78 ng/g 4-cumylphenol [78], and these levels were maintained in the aerobic and anoxic zones of the sludge bioreactor, except for 4-cumylphenol that increased slightly. These levels are low, compared to other WWTP with higher urban and industrial pressure. In that study, a SBSE coupled to GC-MS was used for the determination of APs in wastewaters and sludge [78]. Very few studies report the presence of APs other than OP and NP and, to our knowledge, no study reports the

Transformation Products of Alkylphenols 595

biodegradation of APs within a WWTP. Ko et al. [79] reported the comparison of laserinduced fluorescence and GC-MS to determine phenolic endocrine disrupting compounds in treatment effluents and rivers from South Korea during the period 2002–2004. 4-tert-buthylphenol was detected in 3 sewage samples out of 11 analyzed at 25.7–29.3 mg/L; 4-n-hexylphenol was detected in one sewage water sample at 101.6 mg/L; 4-n-heptylphenol was detected in 9 samples at 79.5–399 mg/L, 4-n-OP was detected at 2 sites and 4-tert-OP in one sewage sample at 185.9 mg/L. Finally, NP was the compound detected in the highest concentrations and in most river and sewage waters at 44.2–244.8 mg/L and 57.4–299 mg/L, respectively [79]. The concentrations and impact of APs is correlated to the anthropogenic activities of a given site, although there is generalized evidence that APs, especially OP and NP, are widespread water contaminants. To control the occurrence of APs, NP and OP have been included in the list of the 33 priority hazardous substances of the European Union Water Framework Directive (directive 2000/60/EC). In addition, Directive 2003/53/EC has also established restrictions on the use, production and marketing of these compounds with the aim to phase out these contaminants. As a result, the levels of APs have been declining in environmental matrices after bans on the use of APs. 19.3.2 Air The occurrence of APs and their TPs in air is an issue of concern due to potential adverse effects on humans and the ecosystem. Dachs et al. [7] reported for the first time the occurrence of NPs in the atmosphere of the lower Hudson river estuary (USA). By sampling atmospheric particulate matter and gas phase samples with a Hi-Vols (flow rate of 0.5 m3/min for 24 h) using quartz fiber filters and polyurethane foam, respectively, and analysis by GC-EIMS, NPs were detected in the atmosphere from 0.1 to 51.4 ng/m3 for the aerosol phase and from 0.2 to 68.6 ng/m3 for the gas phase [7]. In a following study, tert-OP was detected at mean concentration of 0.038 and 0.21 ng/m3 and in the aerosol and gas-phase, respectively, whereas NP mean concentrations were 5.4 ng/m3 in the aerosol and 6.9 ng/m3 in the gas phase. In that study, local sources rather than long-range transport accounted for the main atmospheric input of APs and seasonal trends in the concentration of OP and NP were observed with higher gas-phase concentrations during summer compared to fall [80]. APs were also identified in a mountain site in Fichtelgebirge (NE Bavaria, Germany), 700 m asl, and lower concentrations were observed compared to an urban site [80]. OP was detected at 0.02–0.16 ng/m3 in the gas phase and at 0.3–4.2 pg/m3 in the aerosol; NP was detected at 0.15–1 ng/m3 in the gas phase and at 1.7–117 pg/m3 in the aerosol. In this study, the method was developed to identify the composition mixture of NP-isomers [81,82] in the atmosphere near industrial areas. Xie et al. [83] reported mean concentrations of tert-OP of 0.14 and 0.007 ng/m3 in the vapor and particle phase, respectively, and for NP mean levels of 0.22 and 0.04 ng/m3, in the vapor and particle phase, respectively, in suburban areas of the North sea, and reported the declining trends from suburban areas to the open sea. They conclude that the atmosphere is a significant pathway for the transport of APs in the environment. In a following study, the air–sea exchange fluxes of NP and tert-OP in coastal areas of the North Sea were investigated. NP and tert-OP were detected in the vapor phase of the air sampled at the NE German coast and levels decreased 3–5 times towards the open sea. Atmospheric particulate matter of 10 mm (PM10), collected with active air samplers, can also be used to determine the

596 Transformation Products of Emerging Contaminants in the Environment

occurrence of APs in air. In Thessaloniki, Greece, APs were determined in PM10 in 2 sites, an urban-traffic (1 000 000 inhabitants) site and an industrial site with oil refining, petrochemical, fertilizer and cement production, metal smelting and scrap metal incineration. 24-h sampling collecting was used and mean concentrations of 20–117 ng/m3 of PM10 in urban-traffic site were detected and 32–237 ng/m3 in the industrial site. In the urban site, tert-OP was detected from 0.01 to 0.12 ng/m3 and NP from 1.6 to 16.5 mg/m3. In the industrial site, OP ranged from bdl to 0.02 and NP ranged from 2.5 to 10.9 ng/m3. The levels of APs encountered in the airborne particulates were attributed to wastewater discharges, aeration of sewage tanks, and water-air exchange from polluted surface waters and no differences were observed between urban and industrial sites [84]. Not surprisingly, APs have also been identified as indoor air pollutants. APs, among another 89 contaminants, were the main contaminants detected in indoor air and dust from 120 homes from Cape Cod (USA). 24-h indoor air sampling of 85% 1 Trimethylamine 4.20 75-50-3 2 Acetaldehyde 4.656 75-07-0 3 Methyl formate 5.06 107-31-3 4 Tetramethylhydrazine 5.66 6415-12-9 5 Formaldehyde 6.90 2035-89-4 dimethylhydrazone 6 2-Butanone 7.14 78-93-3 7 Acetonitrile 7.57 75-05-8 8 Acetaldehyde 9.30 7422-90-4 dimethylhydrazone 9 1,10 ,4,40 -Tetramethyltetrazene 15.60 6130-87-6 10 Pyrazine 18.45 290-37-9 11 Dimethylaminoacetonitrile 18.90 926-64-7 12 1-Methyl-1H-pyrazole 19.10 930-36-9 13 N-Nitrosodimethylamine 20.63 62-75-9 14 1,3-Dimethyl-1H-pyrazole 20.77 694-48-4 15 N,N-Dimethylformamide 20.90 68-12-2 16 1,4-Dimethyl-1H-pyrazole 21.65 1072-68-0 17 Dimethyl cyanamide 21.81 1467-79-4 18 1,5-Dimethyl-1H-pyrazole 22.40 694-31-5 19 N,N-Dimethylacetamide 22.80 127-19-5 20 1-Methyl-1H-1,2,4-triazole 25.60 6086-21-1 21 1,3-Dimethyl-1H-1,2,4-triazolea 26.16 16778-76-0 22 1-Formyl-2,226.50 3298-49-5 dimethylhydrazineb 23 1-Methyl-1H-imidazole 27.60 616-47-7 24 Formamide 29.56 75-12-7 25 Benzyl alcohol 29.95 100-51-6 26 1H-Imidazole 34.07 288-32-4 27 1H-1,2,4-Triazole 35.60 288-88-00

Significant ions, m/z (relative abundance, %) 58(100); 59 (46); 42 (35); 30 (13) 44(100); 43(51); 42(13); 41(5) 60 (100); 43(4) 88(100); 73 (89); 44(73); 42(63) 72(100); 71(71); 42(61); 57(15) 43(100); 72(26); 57(4); 40(3) 41(100); 40(52); 39(19); 38(11) 86(100); 44(51); 42(48); 85(34) 116(100); 43(53); 72(42); 42(32) 80(100); 53(32); 52(10); 51(9) 83(100); 84(59) 58(51); 42(51) 82(100); 81(46); 54(22); 42(13) 74(100); 42(37); 43(15); 41(4) 96(100); 95(86); 81(21); 68(14) 73(100); 44(42); 42(19); 72(9) 95(100); 96(75); 68(20); 42(15) 69(100); 70(53); 42(13); 53(8) 96(100); 95(57); 68(14); 53(12) 44(100); 87(75); 43(45); 72(24) 83(100); 56(29); 84(5); 40(4) 97(100); 56(48); 42(11); 98(5) 59(100); 43(73); 42(36); 88(6) 82(100); 81(20); 54(17); 42(14) 45(100); 43(55); 44(27); 42(15) 79(100); 108(93); 107(67); 77(65) 68(100); 41(26); 40(14); 67(9) 69(100); 42(28); 40(4); 41(4)

Preliminary identification of compounds with 50 to 85% probability match with NIST MS library 28 Dimethyldiazene 4.264 503-28-6 43(100); 58(38); 42(22); 57 (4) 29 Ethanamine, 2,20 -oxybis[N,N12.67 3033-62-3 58(100); 42(26); 72(26); 115(5) dimethyl-] 30 Methyldimethylcarbamate 18.31 7541-16-4 72(100); 88(67); 103(63); 44(27) 31 Propiolactam 20.46 930-21-2 71(100); 42(59); 43(19); 41(3) 32 4-Amino-3-penten-2-one 21.28 1118-66-7 84(100); 99(95); 42(61); 43(32) 33 2-Methyl-2H-tetrazole 22.55 16681-78-0 56(100); 57(7); 41(6); 43(6) 34 Methanal-N-methyl-N22.72 61748-05-8 59(100); 57(40); 43(23); 86(20) formylhydrazone 35 3-Methyl-1H-1,2,4-triazole 34.90 7170-01-6 83(100); 42(56); 56(34); 44(25) 36 4-Methyl-4H-1,2,4-triazole 39.24 10570-40-8 83(100); 42(21); 59(7); 55(6) a

Compound identified by analysis of its mass spectrum. Identified by comparison with spectrum of synthesized compound. Note: MS data was obtained in the range of m/z between 34 and 150.

b

630 Transformation Products of Emerging Contaminants in the Environment

NIST’05 and the Wiley7th edition MS libraries search, 27 of them being identified with very high probability. In total, 13 new metabolites were reliably identified. 21.2.2 Examination of Fall Sites Analysis of soil from the sites where carrier rockets fell for the presence of TPs was complicated for a long time by lack of proper analytical methods. Until 2007, samples taken from fall sites in Central Kazakhstan were rarely analyzed for the presence of N-nitrosodimethylamine, 1,10 ,4,40 -tetramethyl-2-tetrazene and 1-methyl-1H-1,2,4-triazole. The latter of these TPs was detected in the majority of samples from these sites while the tetrazene was detected for only a very short period of time after the fuel spill. N-nitrosodimethylamine was mainly detected in sites treated with chemical oxidant. Examination of three selected fall sites in 2009–2010 included massive screening of soil samples for the presence of possible TPs of 1,1-DMH [12]. The sites originated from Proton rocket launches in 2003, 2006, and 2008. Analyses were carried out according to the method described in ref. [9] based on SPME-GC-MS with some modifications. Soil samples (2.0 g) were weighed into 20 mL vials and extracted using 85 mm Carboxen/ polydimethylsiloxane (CAR/PDMS) fiber at 40  C for 30 min without agitation. Desorption was carried out in a split/splitless GC inlet equipped with 0.7 mm direct SPME liner (Supelco, Bellefonte, PA) at 240  C for 10 min. For identification of the TPs, MS detection was done in full scan mode. The results of identification of 1,1-DMH transformation products obtained by analyses of soil samples from the three fall sites included in the present study [12] are given in Table 21.2. The obtained data reveal that the main products of 1,1-DMH transformation (pyrazoles, triazoles, and N-nitrosodimethylamine) were detected at all the sites where discharge of rocket fuel was registered. Formation of other TPs probably depends on the characteristics of the fall sites (chemical and mechanical composition of the soil, moisture content, season, the amount of spilled fuel, etc.). Many of the TPs have further been identified as being more or less readily biodegradable [13]. Thus, with time, many TPs have degraded and are consequently not detected. It is worth noting that the TPs detected in the soil at the oldest of the examined fall sites (FP 3, fall in 2003) were also found at the other siteses (FP 1, fall in 2008 and FP 2, fall in 2006). Dimethylaminoacetonitrile was detected only in soil at the latest landing site (FP 1, fall in 2008). Comparison of the results of the qualitative identification of 1,1-DMH TPs in the laboratory and field investigation (Table 21.3) confirmed most of the results obtained in earlier performed laboratory experiments [9,11,13]. However, in the course of examination of the fall sites, the presence of trimethylamine, acetaldehyde, methyl formate, tetramethylhydrazine, the dimethylhydrazones of formaldehyde and acetaldehyde, and tetramethyltetrazene in soil was not stated, probably due to their relatively low stability (previous laboratory experiments have shown that these compounds may be preserved in soils from the fall sites for up to 1 month only) [1], high volatility, as well as the long period of time between the discharge of the residual rocket fuel and the sampling described here (2–9 years). To confirm the formation of other products of 1,1-DMH transformation, it is necessary to examine further fall sites characterized by different soils, seasons for the falls, amount of discharged fuel and oxidant, and so on.

Identification (Quantitative Determination and Detection) 631 Table 21.2 Products of 1,1-dimethylhydrazine transformation qualitatively identified in soils from fall sites examined in 2009 and 2010. No

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32

RT, min

4.491 11.692 12.039 13.587 13.985 14.323 14.928 15.450 15.630 15.934 16.455 16.724 17.284 17.425 18.373 18.797 18.880 19.580 19.602 19.650 19.706 20.188 20.584 20.773 21.395 21.450 21.652 23.290 23.335 24.164 25.357 26.528

Compound

acetonitrile dimethylaminoacetonitrile 1-methyl-1H-pyrazole 1,3-methyl-1H-pyrazole N-nitrosodimethylamine N,N-dimethylformamide 1,4-dimethyl-1H-pyrazole dimethylcyanamide 5-amino-1,3-dimethyl-1H-pyrazole 1,5-dimethyl-1H-pyrazole 2-methyl-2H-tetrazole 3,4,5-trimethyl-1H-pyrazole 1,3,5-trimethyl-1H-pyrazole 2,5-dimethyl-2H-tetrazole 2-methyl-1,2,3-oxidiazole [1,3,4]-oxadiazole 4-acetyl-3-methyl-1H-pyrazole 4-benzylamine N-methyl-N-nitro-methanamine 4-acetyl-1,5-dimethyl-1H-pyrazole tetrazoboroline, 5-ethyl-1,4-dimethyl1-methyl-1H-1,2,4-triazole 1-ethyl-1H-1,2,4-triazole 1,3-dimethyl-1H-1,2,4-triazole 2-methyl-propanenitrile-2,20 -azobis 1-formyl-2,2-dimethylhydrazine 4-methyl-4H-1,2,4-triazole 4,5-dimethyl-4-imidazoline-2-one 1H-pyrazole 1-methyl-1H-1,2,4-triazole-3-amine 1-methyl-2-amino-1H-imidazole 2,4-dimethylbenzeneamine

2009

2010

FP 1

FP2

FP 1

FP 2

FP 3

þ þ þ þ þ þ þ þ þ þ  þ þ  þ  þ þ  þ þ þ þ þ þ þ þ    þ 

þ  þ þ þ  þ  þ þ þ þ þ        þ þ  þ   þ þ þ þ  

þ þ þ þ þ þ þ   þ þ þ þ þ þ þ      þ þ þ   þ  þ   

þ  þ þ þ þ þ   þ þ þ þ þ  þ     þ þ þ þ   þ þ þ þ þ þ

þ  þ þ þ þ þ   þ  þ  þ þ þ   þ   þ þ þ   þ þ þ þ  

21.2.3 Possible Mechanisms of Formation of the Main Transformation Products 1,1-Dimethylhydrazine is a highly reactive compound and, as shown above, its transformation leads to the formation of many products of different classes. Study of the actual mechanisms in such a complex system becomes especially challenging. There are many discussions going on between chemists concerning the mechanisms, however, experiments remain the only way to describe them. In ref. [14], an attempt was made to describe the mechanisms of formation of the main TPs of 1,1-DMH in water and soil. In the course of the experiments, aqueous extracts from different soils were prepared, spiked with pure UDMH and analyzed frequently during 90 days by

632 Transformation Products of Emerging Contaminants in the Environment Table 21.3 Comparison of the results of the quantitative identification of 1,1-dimethylhydrazine transformation in laboratory and authentic samples. #

Compound

CAS No.

Compounds detected in laboratory and authentic samples 1 acetonitrile 2 pyrazine 3 dimethylaminoacetonitrile 4 1-methyl-1H-pyrazolea 5 N-nitrosodimethylaminea 6 1,3-dimethyl-1H-pyrazolea 7 N,N-dimethylformamide 8 1,4-dimethyl-1H-pyrazolea 9 dimethylcyanamide 10 1,5-dimethyl-1H-pyrazolea 11 N,N-dimethylacetamide 12 1-methyl-1H-1,2,4-triazolea 13 1,3-dimethyl-1H-1,2,4-triazolea 14 1-formyl-2,2-dimethylhydrazine 15 2-methyl-2H-tetrazole 16 3-methyl-1H-1,2,4-triazole 17 4-methyl-4H-1,2,4-triazolea

75-05-8 290-37-9 926-64-7 930-36-9 62-75-9 694-48-4 68-12-2 1072-68-0 1467-79-4 694-31-5 127-19-5 6086-21-1 16778-76-0 3298-49-5 16681-78-0 7170-01-6 10570-40-8

Compounds detected only in authentic samples 18 5-amino-1,3-dimethyl-1H-pyrazole 19 3,4,5-trimethyl-1H-pyrazole 20 1,3,5-trimethyl-1H-pyrazole 21 2,5-dimethyl-2H-tetrazole 22 2-methyl-1,2,3-oxadiazole 23 [1,3,4]-oxadiazole 24 4-acetyl-3-methyl-1H-pyrazole 25 p-toluidine 26 N-methyl-N-nitro-methaneamine 27 4-acetyl-1,5-dimethyl-1H-pyrazole 28 tetrazoboroline, 5-ethyl-1,4-dimethyl29 1-ethyl-1H-1,2,4-triazole 30 2-methyl-propanenitrile-2,20 -azobis 31 N-ethylidene-1-pyrrolidineamine 32 4,5-dimethyl-4-imidazoline-2-one 33 1H-pyrazole 34 1-methyl-1H-1,2,4-triazole-3-amine 35 1-methyl-2-amine-1H-imidazole

3524-32-1 11072-91-9 1072-91-9 16681-78-0 3451-51-2 288-99-3 105224-04-2 106-49-0 4164-28-7 21686-05-5 20534-01-4 16778-70-4 78-67-1 60144-27-6 1072-89-5 288-13-1 49607-51-4 6646-51-1

Compounds detected only in laboratory samples 36 trimethylamineb 37 acetaldehydeb 38 methyl formateb 39 tetramethylhydrazineb 40 formaldehyde dimethylhydrazoneb 41 2-butanone 42 acetaldehyde dimethylhydrazoneb 43 1,10 ,4,40 -tetramethyl-2-tetrazeneb 44 1-methyl-1H-imidazole

75-50-3 75-07-0 107-31-3 6415-12-9 2035-89-4 78-93-3 7422-90-4 6130-87-6 616-47-7

Identification (Quantitative Determination and Detection) 633 Table 21.3 (Continued) #

Compound

CAS No.

45 46 47 48 49 50 51 52 53 54

formamide benzyl alcohol 1H-imidazole 1H-1,2,4-triazole dimethyldiazene ethaneamine, 2,20 -oxibis [N,N-dimethyl-] methyldimethylcarbamate propiolactam 4-amino-3-pentene-2-one methanal-N-methyl-N-formylhydrazone

75-12-7 100-51-6 288-32-4 288-88-00 503-28-6 3033-62-3 7541-16-4 930-21-2 1118-66-7 61748-05-8

Note: a Compounds detected in soils of all fall sites where the rocket fuel spillages were detected. b Unstable or volatile compounds, whose presence in real samples can be determined only in the immediate sampling and analysis after spillage of 1,1-dimethylhydrazine.

a precise and highly optimized for the experiment method, based on HPLC with diode array detection. The collected data are represented in Figure 21.1. In the course of the experiments, it appeared that degradation of 1,1-DMH occurs by a series of consecutive and parallel chemical reactions. Formaldehyde dimethylhydrazone is

(b)

80 60

C, mg/L

C, mg/L

(a)

1

40 2

20 0

C, mg/L

(c)

3 4

0

20

40 60 t, days

80

100

80 60

1

40 20 0

2

0

3 4

20

40 60 t, days

80

100

80 60 1

40 20 0

2 3 4

0

20

40 60 t, days

80

100

Figure 21.1 Degradation of spiked 1,1-DMH in aqueous extracts from soils. (a) distilled water; (b) aqueous extract from sand; (c) aqueous extract from heavy loam. 1, 1,1-DMH; 2, formaldehyde dimethylhydrazone; 3, 1-formyl-2,2-dimethylhydrazine; 4, 1-methyl-1H-1,2,4-triazole.

634 Transformation Products of Emerging Contaminants in the Environment H C N

N H3C

H3 C N

H3C

NH2

HC N

H3 C

N

CH2

CH3

N

H 3C N H 3C

H N

O C H

Figure 21.2 Proposed scheme of transformation of 1,1-dimethylhydrazine [14].

the main intermediate product of 1,1-DMH transformation and the formation of the final products – 1-methyl-1H-1,2,4-triazole and 1-formyl-2,2-dimethylhydrazine (Figure 21.2). These results were later confirmed [1] when studying the transformation of 1,1-DMH in soils using GC-MS of the headspace above spiked soils. In addition, it was suggested that formation of 1,3-dimethyl-1H-1,2,4-triazole involves a further intermediate – acetaldehyde dimethylhydrazone. However, this suggestion requires further confirmation. In the opinion of researchers working in this field, oxygen is the main oxidative reagent in the transformation processes of dimethylhydrazine in water and soil [1,15]. However, many compounds present in soil and water can enhance chemical transformations in the system; some of them can act as catalysts. Such properties were confirmed for ions of the transition metals (Mn2þ, Fe3þ and Cu2þ) [16].

21.3 Distribution and Fate of Transformation Products of 1,1-Dimethylhydrazine in Soil at Fall Sites The first stage of a “Proton” rocket has six fuel tanks with 1,1-DMH and six boosters. After separation of the first stage, the weight of the boosters is significantly higher than the weight of the nearly empty fuel tanks. Consequently, during landing of the first stage, the rocket boosters are the first to touch the land, forming an 8–10 m crater. As the first stage has six boosters, the result may be several large (more than 50 cm) and numerous small spots of fuel at and around the epicenter of the fall site [12]. Examination of a 100  100 m2 square at the epicenter of the fall site showed that epicenter TPs of 1,1-DMH are present only in the crater formed as a result of the rocket landing [12]. Most metabolites can be detected only at the sites of obvious fuel discharge. The obtained patterns of horizontal distribution (Figure 21.3) show the highest concentrations of MTA at the sites of obvious discharge of rocket fuel with gradual decrease with distance. Using these data it was concluded that horizontal migration of the TPs in soil is very slow. It can be caused by a serious deformation of the soil during a rocket landing, leading to a decrease in pore sizes that can be used as a pathway for migration. Also, migration can be inhibited by a high affinity of TPs to the soil, together with high water solubility and low Henry’s Law constant values (see Chapter 29). The results of a detailed study of distribution of 1,1-DMH TPs along the soil profile showed that TPs can migrate down to a depth of approximately 120 cm (Table 21.4). However, the migration rate appears rather slow. After 2–3 years following the landing studied, the highest concentrations of 1,1-DMH TPs are registered, as a rule of thumb, at a depth of

Identification (Quantitative Determination and Detection) 635

Figure 21.3 Distribution of 1-methyl-1H-1,2,4-triazole at fall site 1 at different depths: (a) 0–10 cm; (b) 10–20 cm; (c) 20–30 cm; (d) 30–40 cm.

20–60 cm, however, this value can vary significantly, depending on the type, humidity and physical properties of the soil, the features of the relief, and other conditions. In the surface layer, again as a rule of thumb, only semi-volatile TPs are registered, probably as a result of fast biodegradation in combination with the high evaporation rate of the volatile TPs [12]. Taking into account the high water solubility of all 1,1-DMH metabolites, they can easily migrate down the soil profile with rain or water from melting snow reaching groundwaters, especially in the spring. The high stability of TPs in deeper soil horizons can be explained by lower access to oxygen, that is considered as the main oxidant in the transformation of 1,1-dimethylhydrazine [15], and much lower microbial activity.

21.4 Analytical Methods Applied in the Monitoring Scientists and engineers were challenged with a lack of robust analytical methods and sampling techniques for analysis of environmental samples for the content of 1,1-DMH and its

636 Transformation Products of Emerging Contaminants in the Environment Table 21.4 The response of 1,1-dimethylhydrazine TPs in soil samples taken from different depths in FP2 (point w3). RT, min

Compound

12.03 13.61 14.02 18.78 19.08 20.10 20.50 20.70

1-methyl-1H-pyrazole 1,3-methyl-1H-pyrazole N-nitrosodimethylamine [1,3,4]-oxidiazole 3-methyl-1H-pyrolizidine 1-methyl-1H-1,2,4-triazole 1-ethyl-1H-1,2,4-triazole 1,3-dimethyl-1H-1,2,4triazole 4-methyl-4H-1,2,4-triazole N-ethylidene-1pyrodinamine 1H-pyrazole 1-methyl-1H-1,2,4-triazole3-amine

21.63 23.09 23.33 24.16

Peak area, 106 arbitrary units 0– 20

20– 40

40– 60

60– 80

10 5 2 0.2 0.3 9 0.6 1

14 5 3 0.5 0.3 8 0.9 0.9

12 42 3 0.2 0.3 4 0.5 0.6

0.4 0.1 0.4 0.06 0.07 1 n/d 0.1

0.2 n/d n/d n/d n/d 0.1 n/d 0.06

n/d n/d n/d n/d n/d n/d n/d n/d

n/d n/d n/d n/d n/d n/d n/d n/d

1 0.2 0.08 0.7

2 2

2 1

0.7 0.2

n/d n/d

n/d n/d

n/d n/d

0.1 0.1

1 1

1 0.9

0.3 0.3

n/d 0.1

n/d n/d

n/d n/d

n/d 1 n/d 0.07 n/d 2 0.1 0.4

0.7 0.5

80– 100

100– 120

120– 140

140– 160

Note: The maximum response of each compound is indicated in bold.

TPs. Determination of 1,1-DMH and its TPs in environmental samples is a quite complex task, primarily due to the reactivity and on-going chemical and biological processes in the samples. Such analytical methods are especially needed, since 1,1-DMH is considered to be very reactive, readily converting to TPs upon contact with air, water and soil. Environmental factors driving these processes are only little known. Analytical methods should consider and minimize chemical transformations, especially during sample preparation. Inefficiency and, in many cases, the absence of the available methods for analysis of environmental samples contaminated with TPs of 1,1-dimethylhydrazine were the main factors limiting study of the fate of rocket fuel residuals at the sites of its spill. Research during the first years was mainly focused on methods for determination of 1,1-DMH [9]. A huge development in this direction started when the GC-MS technique was used [7]. 21.4.1 Methods of Quantitative Determination 21.4.1.1 N-Nitrosodimethylamine NDMA is the most toxic of the known TPs of 1,1-dimethylhydrazine. A clean-up level for NDMA in groundwater established by the US EPA is 0.7 ng/L [17]. Maximum acceptable concentration (MAC) for NDMA in water varies from country to country. For Russia and Kazakhstan, the MAC value is set to 1 mg/L [18]. Detection of NDMA at such levels requires highly sensitive methods. A standard method of the US EPA [19] is based on solid phase extraction (SPE) of NDMA from water, subsequent elution, concentration, and determination by large-volume injection capillary gas chromatography with negative chemical ionization tandem mass spectrometry.

Identification (Quantitative Determination and Detection) 637

The detection limit of the method is 1 ng/L, however it can vary for different inlets and mass spectrometers. Separation is performed on a Restek Rtx 5SIL MS 30 m  0.25 mm, 1.0 mm column. Other columns with dimensions not less than the indicated values can be used. The method is very sensitive, but due to its complex instrumentation not widely applicable. Certified methods for determination of NDMA in water [20] and soil [21] were developed by the group of O.A. Shpigun and A.D. Smolenkov from M.V. Lomonosov Moscow State University. The methods are based on HPLC with UV detection at 240 nm. Sample preparation for the soil is based on basic distillation of NDMA, for water, sample preparation is not required. At an injection volume of 100 mL, the detection limit for soil is 0.02 mg/kg, and for water it is 0.01 mg/L. However, the validity of the results obtained using the method for determination of NDMA in soil is questionable as it was proved that basic distillation causes many chemical reactions to proceed, leading to excessive formation of 1,1-DMH and possibly NDMA [8]. Another serious drawback of these methods is the unknown compositions of HPLC column and mobile phase – they must be purchased directly from the developer. In Kazakhstan, the price of this column is reported to be about 1.5–2 times higher than price of a normal reversed-phase column. A similar HPLC-DAD method was developed for direct quantitative determination of NDMA in aqueous samples [1] using a 75  4.6 mm, 3.0 mm particle size Zorbax XDB-Phenyl column (Agilent, USA). The reported detection limit for an injection of 20 mL was 0.1 mg/L, with the length of the chromatographic run being 2.5 min. The method was successfully applied also for aqueous extracts from soils and allowed simultaneous determination of other main TPs – 1-methyl-1H-1,2,4-triazole, 1-formyl-2,2-dimethylhydrazine and formaldehyde dimethylhydrazone. A sample chromatogram is given in Figure 21.4. In spite of the similar retention times of NDMA and MTA, they do not interfere with each other because they have absorption maxima at different wavelengths. Detection limits for MTA, FDMH and FADMH are reported to be 0.2, 0.5 and 0.2 mg/L, respectively. Recently, a few methods have been developed for the determination of NDMA in aqueous samples using a novel sampling and sample preparation method – SPME, a quite simple, reliable and easily automated technique. Application of SPME in combination with gas chromatography provided detection limits for NDMA at a level of 30 ng/L, using mass spectrometric detection in negative chemical ionization mode [22]. Further, the method was optimized for analysis of water and aqueous extracts from soil contaminated with 1,1-DMH-based rocket fuel residuals [23]. Special attention was given to the stage of desorption of analytes from the SPME fiber coating in the GC inlet to minimize analyte losses and any chemical interaction between TPs. The desorption temperature was lowered to 200  C. To accelerate NDMA desorption from the fiber coating at such low temperature and provide better peak shape, a less efficient fiber coating based on polydimethylsiloxane/divinylbenzene (PDMS/DVB) was applied instead of the CAR/PDMS coating. The extraction temperature was also lowered to 30  C. The sample volume and the amount of NaCl added were selected from the method [22]: 8.0 mL and 2.8 g, respectively. The extraction time was chosen to be 10 min. The method was optimized for automated determination of NDMA: analysis of a single sample takes 1 h, while analysis of a series of samples takes 40 min per sample. A sample chromatogram is presented in Figure 21.5. The linearity of a calibration curve plotted based on analyses of standard solutions of NDMA in triplicate was determined to be 1.0000. The error of the method was calculated to be less than 20% for the whole concentration range (2–100 mg/L).

638 Transformation Products of Emerging Contaminants in the Environment

Figure 21.4 HPLC-DAD chromatograms of a water sample contaminated with TPs of 1,1-dimethylhydrazine at different wavelengths.

200 180 Abundance

160 140 120 100 NDMA

80 60 40 10

11

12

13

14

15

t, min

Figure 21.5 Fragment of SIM (m/z 74) HS SPME-GC-MS chromatogram of NDMA sample. CNDMA ¼ 1.0 mg/L, PDMS/DVB fiber coating, extraction temperature 30  C, extraction time 10 min, sample volume 8.0 mL, vial volume 20 mL, NaCl added 2.8 g, desorption temperature 200  C; column HP-Innowax 30 m  0.25 mm, 0.25 mm film.

Identification (Quantitative Determination and Detection) 639

Quantitative determination of NDMA in soil samples is performed after extracting it with organic solvent (typically, acetone or methylene chloride) [7], however, for real samples, the sensitivity of the available methods (>10 mg/kg) is typically insufficient [12]. 21.4.1.2 1-Methyl-1H-1,2,4-Triazole In spite of its low toxicity, determination of 1-methyl-1H-1,2,4-triazole in environmental samples affected by rocket fuel spills is of huge importance due to its high stability [1] and migration potential [13]. As was shown above, it can also serve as a marker for 1,1-DMHbased rocket fuel contamination. Determination of MTA in soil samples can be performed by GC-MS after extraction with acetone or methylene chloride in a Soxhlet apparatus [24] or under ultrasound [7], followed by clean-up (if necessary) and evaporative concentration. The method [24] was optimized for determination of MTA in soil samples using a HPInnowax 30 m  0.25 mm, 0.25 mm film, column, providing the highest retention of MTA and the best peak shape (Figure 21.6). To achieve the highest sensitivity, MS detection was applied in selected ion monitoring (SIM) mode using the molecular peak of MTA at m/z 83. The detection limit was determined to be 5 mg/kg. The method was successfully applied in numerous examinations of landing sites in Central Kazakhstan and various laboratory experiments allowed detection of other TPs of 1,1-DMH – pyrazoles, triazoles and tetrazoles [12]. A method based on IC-MS for the determination of MTA, FDMH, dimethyl guanidine and dimethyl amine was described in [25]. Optimized extraction from soil was performed with methanol (for FDMH) and slightly alkaline water (for other analytes). Separation was carried out on a Nucleosil 5 SA (150  2 mm) cation-exchange column. Detection was conducted using atmospheric pressure chemical ionization (APCI) registering positively charged molecular ions. Levels of quantification made up 10, 50, 50 and 250 mg/L for FDMH, MTA, dimethyl guanidine and dimethyl amine, respectively.

Figure 21.6 Chromatogram of a sample containing MTA using optimized GC-MS method.

640 Transformation Products of Emerging Contaminants in the Environment

A further study [26] reports a GC-MS based method for simultaneous determination of FDMH and MTA in soil samples. The authors pay their highest attention to the preparation of certified reference material – soil containing an exact amount of analyte, followed by study of the efficiency of extraction of the analytes from soil using different solvents. It was shown that the most efficient method for preparation of certified soil is spiking of MTA and FDMH solutions in methylene chloride, followed by evaporation of the solvent. Adsorption of MTA by soil from MTA solution was found to be most efficient, while other solvents tested (water and methanol) showed strong retention of MTA and its losses during solvent evaporation. The most efficient solvent for ultrasonic extraction of MTA and FDMH from soil was found to be methanol, providing an analyte recovery of 58–72%, depending on the soil type. The reported concentration range for this method is 0.05–50 mg/kg. 21.4.1.3 1-Formyl-2,2-Dimethylhydrazine Determination of 1-formyl-2,2-dimethylhydrazine in environmental samples is complicated as it is one of the most polar and least volatile TPs of 1,1-dimethylhydrazine. Under certain conditions FDMH can degrade with formation of the original pollutant. As described in Section 21.4.1.2, FDMH was first detected using IC-MS [8], a technique that later became the key one for its quantitative determination in environmental samples [25]. Another method developed by the same group is based on GC-MS [26]. Both methods are described in Section 21.4.1.2. FDMH can also be determined in water and aqueous extracts from soils using the HPLCDAD method at 195 nm described in Section 21.4.1.1. 21.4.1.4 Formaldehyde Dimethylhydrazone As described above, formaldehyde dimethylhydrazone is the metabolite formed in the very first period after the spill of 1,1-dimethylhydrazine. Then FADMH is decomposed to other transformation products like MTA and FDMH. FADMH is unstable in soil, but in water it can be present for a period close to 1 month, depending on its initial concentration, water composition and climatic conditions [1]. According to QSAR-based calculations [27], FADMH possesses toxicity close to the level of 1,1-DMH. Determination of FADMH, especially in water samples is very important for studies on the fate of rocket fuel residuals. There are two methods available for quantitative determination of FADMH in water samples. One is based on HPLC-DAD at 237 nm, the method described in Section 21.4.1.1. A second method was developed using GC-MS and GC-NPD in combination with SPME [28]. The desorption temperature was shown to be an important factor affecting FADMH losses due to thermal degradation in the GC inlet. An optimal desorption of FADMH providing the best analyte peak shape was achieved at a temperature of 200  C when using 65 mm PDMS/DVB fiber coating (Figure 21.7). The optimized extraction parameters are: 1-min headspace extraction at 30  C; sample volume 2.0 mL, vial volume 20 mL, amount of NaCl added 0.70 g, pH > 8.5. Detection limits were estimated to be 1.5 and 0.5 mg/L using mass selective and nitrogen-phosphorus detection, respectively. The developed method was successfully applied to laboratory-scale experiments to quantify FADMH formed due to degradation of 1,1-dimethylhydrazine.

Identification (Quantitative Determination and Detection) 641

10000

15000

PDMS/DVB Abundance

Abundance

15000

FADMH

5000

0 2.0

2.5

3.0

3.5 4.0 time, min

4.5

5.0

CAR/PDMS

10000

FADMH

5000

0 2.0

2.5

3.0 3.5 4.0 time, min

4.5

5.0

Figure 21.7 Chromatograms of FADMH-containing aqueous samples obtained using different SPME fiber coatings.

21.4.1.5 1,10 ,4,40 -Tetramethyl-2-Tetrazene There are a set of standard methods for the determination of 1,10 ,4,40 -tetramethyl-2-tetrazene in water [29] and soil [30] based on ion chromatography with amperometric detection. Sample preparation for soil is based on alkaline distillation. Water samples do not require sample preparation. Detection is conducted on a carbon electrode at a potential of 1.3 V. With an injection volume of 500 mL, the detection limit for soil samples is 0.125 mg/kg, and 0.05 mg/ L for water samples. 21.4.2 Systematic Approach to the Control of TPs of 1,1-Dimethylhydrazine As was shown above, at present a sufficient number of powerful analytical methods are available for quantitative determination of TPs of 1,1-dimethylhydrazine. All the most important methods are listed in Table 21.5. However, the main lack of the most available methods is the time- and labor-consuming sample preparation. Most of the methods still work only for a single analyte. Laboratories performing monitoring of fall sites for the presence of TPs still experience the problem of the high cost of these studies limiting the number of samples that can be analyzed. Screening multi-residue analytical methods become more and more popular in the field of environmental analysis. They are typically cost-friendly, sensitive to a broad range of analytes and provide collection of very valuable data for further decisionmaking and quantitative analyses. Such a method was developed for the TPs of 1,1-dimethylhydrazine using GC-MS in combination with SPME [9]. The method is selective for almost the whole range of known metabolites of 1,1-DMH – from the most volatile to the semi-volatile, from very polar to mid-polar analytes. It was achieved using an 85 mm CAR/PDMS fiber coating during 1-hr soil headspace sampling. An extraction temperature of 40  C was selected to minimize chemical processes in the soil and the headspace. However, for soil samples originating from 5þ years old fall sites of rocket-carriers, this temperature can be increased to 70–80  C to enhance the desorption of analytes from soil samples. In the case of insufficient analyte responses, the extraction time can be extended to close to 24 h. Special attention in the study was given to degradation of reactive analytes in the GC inlet, initially leading to losses of target analytes and the formation of new compounds at

Identification (Quantitative Determination and Detection) 643

Abundance

(a) 5,00E+009 4,00E+009 3,00E+009 2,00E+009 1,00E+009 0,00E+000

5

0

Abundance

(b) 5,00E+009

10

4

15

20

5

6

25 t, min

30

35

40

45

50

25 t, min

30

35

40

45

50

4,00E+009 3,00E+009 2,00E+009

1

3 2

1,00E+009

7

0,00E+000 0

5

10

15

20

Figure 21.8 Comparisons of chromatograms of 1.00 g soil sample spiked with 0.67 mL of UDMH obtained by headspace SPME with 85 mm CAR/PDMS fiber using different GC inlet temperature programs: (a) 170  C hold for 0.1 min, then 1  C s1 ramp to 250  C, hold for 40 min; (b) constant temperature of 250  C. Sampling conditions: time ¼ 15 s, room temperature (20  C). Note: Peaks: 1, dimethylamine; 2, methanediamine, N,N,N0 ,N0 -tetramethyl-; 3, tetramethylhydrazine; 4, formaldehyde dimethylhydrazone (co-eluting with 1,1-DMH); 5, acetaldehyde dimethylhydrazone; 6, 1,10 ,4,40 -tetramethyl-2-tetrazene; 7, N1,N1-dimethyl-N2-(dimethylamino) formamidine.

temperatures higher than 170  C required for fast desorption from the CAR/PDMS fiber coating. The GC inlet temperature was optimized using the capability of a programmable temperature vaporization (PTV) inlet at 170  C, held for 0.1 min, then 1  C s1 ramp to 250  C where it was held for 40 min. Temperature programming resulted in a fast desorption along with minimal chemical transformation in the GC inlet. The effect of the optimized desorption temperature program on the TPs pattern in a soil sample is shown in Figure 21.8. The use of SPME resulted in high sensitivity (especially for most volatile metabolites of 1,1-DMH), speed, small labor consumption due to automation, and simplicity of use. The only limitation of the method was found to be a significant loss of sensitivity in the case of elevated moisture content of soils. However, the method was shown to be perfect for screening 1,1-DMH TPs in soils originating from Central Kazakhstan where the first stages of “Proton” rocket-carriers land [12]. The developed method was used as a key to the development of a systematic approach to control the TPs of 1,1-dimethylhydrazine in environmental samples (Figure 21.9) presented in [34]. This approach provides collection of an excessive amount of data at minimum cost. In practice, more than 30% of soil samples were rejected as uncontaminated after results collected by the screening method [12].

644 Transformation Products of Emerging Contaminants in the Environment The site being examined

Sampling

Air

Soil

Water

Analysis by the screening method (SPME/GC/MS)

Objects polluted with transformation products

NDMA

TMT

FDMH

MTA

Analysis by standard methods

Others

Uncontaminated

Report

Semi-quantitative determination by SPME-GC-MS

Report

Figure 21.9 A systematic approach to controlling environmental objects on sites of hydrazinebased rocket fuel spills.

21.5 Conclusion This chapter presents a review of the most recent data on the TPs of 1,1-dimethylhydrazine, including their identification, quantitative determination, distribution and fate in the environment. It was shown that analytical methods for monitoring TPs in the environment have drastically improved during the last 5 years. The highest success was achieved when applying mass spectrometry-based techniques – GC-MS and LC-MS. These methods provided efficient separation of many transformation products, identification of unknown metabolites together with their sensitive and selective determination. Sample preparation was mainly based on classic techniques, including organic solvent extraction and concentration. However, utilization of modern solventless methods based on headspace extraction and HSSPME provided efficient detection of the most volatile TPs that were impossible to detect using classic sample preparation techniques. The HS SPME technique was optimized for the whole range of TPs in soil; it was so efficient that, in combination with GC-MS, it provided the highest level of sensitivity for most of the known TPs. Up till now more than 50 TPs have been detected in environmental samples taken from the field and during laboratory experiments. There have been a number of attempts to study the processes of chemical transformation in soil and water. It was shown that formaldehyde dimethylhydrazone is the main intermediate product of 1,1-DMH transformation and the formation of the final products – 1-methyl-1H-1,2,4-triazole and 1-formyl-2,2-dimethylhydrazine In

Identification (Quantitative Determination and Detection) 645

spite of a huge success in this field, mechanisms of the chemical transformation of 1,1-dimethylhydrazine in the environment remain unclear, mainly due to their complexity. The high stability of TPs, especially those with cyclic structures (triazoles, pyrazoles and tetrazoles), in the environment of Central Kazakhstan where such studies were performed, can be caused by low microbial activity and limited oxygen access at deeper soil levels due to soil compression after rocket landing. This compression, together with the high affinity of TPs to soil, slows down their horizontal migration. The vertical migration of TPs in the conditions of Central Kazakhstan is rather slow. The maximum depth where transformation products were detected was 120 cm. However, due to high water solubility, the risk of reaching groundwater still exists.

References 1. Alimzhanova, M.B. (2009) Transformation products of 1,1-dimethylhydrazine and their determination using chromatographic methods. Dissertation for Candidate of Chemical Sciences Degree, Almaty (in Russian). 2. Urry, W.H., Olsen, A.L., Bens, E.M. et al. (1965) Autoxidation of 1,1-dimethylhydrazine, US. Naval Ordinance Test Station Technical Publication 3903. 3. Slonim, A.R. and Brennan, G. (1976) Hydrazine degradation in aquatic systems. Environmental Contamination & Toxicology, 16, 301–330. 4. Mathur, M.A. and Sisler, H.H. (1981) Oxidation of 1,1-dimethylhydrazine by oxygen. Inorganic Chemistry, 20, 426–429. 5. Sisler, H.H., Mathur, M.A., Jaln, S.R., and Greengard, R. (1981) Studies of chloramination of dimethylamine and 1,1-dimethylhydrazine. Industrial & Engineering Chemistry Product Research and Development, 20, 181–185. 6. Ou, L.T. and Street, J.J. (1987) Microbial enhancement of hydrazine degradation in soil and water. Environmental Contamination & Toxicology, 39, 541–548. 7. Buryak, A.K., Tataurova, O.G., and Ulyanov, A.V. (2004) Study of transformation products of unsymmetrical dimethylhydrazine on model sorbents by the method of gas chromatography/mass spectrometry. Russian Journal of Mass Spectrometry, 1-2, 147–152 (In Russian). 8. Smolenkov, A.D., Rodin, I.A., Shpak, A.V., and Shpigun, O.A. (2007) 1-Formyl-2,2-dimethylhydrazine as a new decomposition product of 1,1-dimethylhydrazine. International Journal of Environmental Analytical Chemistry, 87, 351–359. 9. Kenessov, B., Koziel, J.A., Grotenhuis, T., and Carlsen, L. (2010) Screening of transformation products in soils contaminated with unsymmetrical dimethylhydrazine. Analytica Chimica Acta, 674, 32–39. 10. Rodin, I.A., Moskvin, D.N., Smolenkov, A.D., and Shpigun, O.A. (2008) Transformations of asymmetric dimethylhydrazine in soils. Russian Journal of Physical Chemistry A, 82, 911–915. 11. Kenessov, B.N. (2008) Identification of transformation products of unsymmetrical dimethylhydrazine in soils by the method of headspace extraction in combination with chromato-mass-spectrometry. News of the National Academy of Science of Kazakhstan. Chemical Series, 5, 48–53 (in Russian). 12. Kenessov, B., Alimzhanova, M., Sailaukhanuly, Ye. et al. (2012) Transformation products of 1,1dimethylhydrazine and their distribution in soils of fall places of rocket carriers in Central Kazakhstan. Science of the Total Environment, 427–428, 78–85. 13. Carlsen, L., Kenessov, B.N., and Batyrbekova, S.Ye. (2008) A QSAR/QSTR study on the environmental health impact by the rocket fuel 1,1-dimethyl hydrazine and its transformation products. Environmental Health Insights, 1, 11–20.

646 Transformation Products of Emerging Contaminants in the Environment 14. Alimzhanova, M.B., Kenessov, B.N., Batyrbekova, S.Ye. et al. (2009) Transformation of unsymmetrical dimethylhydrazine in aqueous extracts from soils. News of the National Academy of Science of Kazakhstan. Chemical Series, 1, 87–92 (in Russian). 15. Batyrbekova, S.Ye. (2009) System analysis of environmental objects suffered from a negative impact of Baikonur cosmodrome activity. Dissertation for Doctor of Chemical Sciences Degree, Almaty (in Russian). 16. Alimzhanova, M.B., Doszhan, G., Kenessov, B.N. et al. (2009) Study of 1,1-dimethylhydrazine transformation processes in water in the presence of iron (III), copper (II) and manganese (II) cations. KazNU Bulletin. Chemical Series, 54, 139–144 (in Russian). 17. US EPA (July 20 2001) Record of Decision for the Western Groundwater Operable Unit OU-3, Aerojet Sacramento Site. 18. Adushkin, V.V., Kozlov, C.I., and Petrov, A.V. (eds) (2000) Ecological Problems and Risks of Rocket-Space Techniques Impact on the Environment, Ankil, Moscow, (in Russian). 19. US EPA (2004) Method 521 (rev. 1.0) Determination of Nitrosamines in Drinking Water by Solid Phase Extraction and Capillary Column and Solid Phase Extraction and Capillary Column Gas Chromatography, Cincinnati, Ohio, USA. 20. (2002) A Standard Method for Determination of Nitrosodimethylamine in Samples of Natural Water Using Reversed Phase Chromatography with Spectrophotometric Detection (MVI 3-02), Moscow State University, Moscow (in Russian). 21. (2002) A Standard Method for Determination of Nitrosodimethylamine in Samples of Soil Using Reversed Phase Chromatography with Spectrophotometric Detection (MVI 1-02), Moscow State University, Moscow (in Russian). 22. Grebel, J.E., Young, C.C., and Suffet, I.H. (2006) Solid phase microextraction of N-nitrosamines. Journal of Chromatography A, 1117, 11–18. 23. Kenessov, B.N. (2009) Use of solid-phase microextraction as sample preparation method for determination of N-nitrosodimethylamine in water contaminated due to hydrazine-based rocket fuel spills. KazNU Bulletin. Chemical Series, 54, 92–97 (In Russian). 24. Kenessov, B., Batyrbekova, S., Nauryzbayev, M. et al. (2008) GC-MS determination of 1-methyl1H-1,2,4-triazole in soils affected by rocket fuel spills in Central Kazakhstan. Chromatographia, 57, 421–424. 25. Rodin, I.A., Anan’eva, I.A., Smolenkov, A.D., and Shpigun, O.A. (2010) Determination of the products of the oxidative transformation of unsymmetrical dimethylhydrazine in soils by liquid chromatography/mass spectrometry. Journal of Analytical Chemistry, 65, 1405–1410. 26. Smirnov, R.S., Rodin, I.A., Smolenkov, A.D., and Shpigun, O.A. (2010) Determination of the products of the transformation of unsymmetrical dimethylhydrazine in soils using chromatography/mass spectrometry. Journal of Analytical Chemistry, 65, 1266–1272. 27. Carlsen, L., Kenessov, B.N., and Batyrbekova, S.Ye. (2009) A QSAR/QSTR study on the human health impact by the rocket fuel 1,1-dimethylhydrazine and its transformation products. Environmental Toxicology and Pharmacology, 27, 415–423. 28. Kenessov, B., Sailaukhanuly, Ye., Koziel, J.A. et al. (2011) GC-MS and GC-NPD determination of formaldehyde dimethylhydrazone in water using SPME. Chromatographia, 73, 123–128. 29. (2002) A Standard Method for Determination of Tetramethyltetrazene in Samples of Natural Water Using Ion Chromatography with Amperometric Detection (MVI 4-02), Moscow State University, Moscow (in Russian). 30. (2002) A Standard Method for Determination of Tetramethyltetrazene in Samples of Soil Using Ion Chromatography with Amperometric Detection (MVI 2-02), Moscow State University, Moscow (in Russian). 31. Yergozhin, Ye.Ye., Solomin, V.A., Lyapunov, V.V., and Tsukerman, V.G. (2000) A Method for Determination of 1,1-Dimethylhydrazine and Nitrosodimethylamine, Prelim. Patent of the Republic of Kazakhstan No. 8455, Nat. Patent. Agency of Kazakhstan, Almaty (in Russian).

Identification (Quantitative Determination and Detection) 647 32. Ponomarenko, S.A. (2009) Simultaneous Determination of Several Hydrazines and Nitrosodimethylamine Using Ion Pair Chromatography, Author’s Abstract of Dissertation for the Degree of Candidate of Chemical Sciences, Moscow (in Russian). 33. Kenessov, B.N., Sailaukhanuly, Ye., Musrepov, B.A., and Kaldarov, A.K. (2008) Determination of 1-methyl-1H-1,2,4-triazole in water samples by high-performance liquid chromatography with diode-array detection. KazNU Bulletin. Chemical Series, 1, 184–190 (in Russian). 34. Kenessov, B.N. (2010) A system of analytical control of environmental objects at sites of hydrazine-based rocket fuel spills. KazNU Bulletin. Chemical Series, 59, 403–406 (in Russian).

22 Assessment of the Occurrence and Fate of Transformation Products of Endocrine Disrupting Compounds EDCs in the Environment Vasiliki Boti, Vasilios Sakkas and Triantafyllos Albanis Laboratory of Analytical Chemistry, Department of Chemistry, University of Ioannina, Greece

22.1 Introduction Over the past few decades, accumulating evidence has suggested that many environmental contaminants can alter chemical signaling among cells. Wildlife studies from various environments worldwide implicate environmental contaminants as probable causal agents in various widespread human diseases. The increasing number of xenobiotic chemicals adversely influences wildlife by altering reproduction, growth, and survival, by changing the normal function of the endocrine system. These chemicals have been observed to mimic hormones, act as antihormones, or alter the synthesis and/or degradation of hormones. Thus, they have been termed endocrine disrupting compounds (EDCs). EDCs may be transformed by biotic and abiotic processes, when released into the environment. These transformation processes usually involve a cascade of reactions following a plethora of pathways and resulting in the formation of a number of transformation products (TPs), metabolites or degrades. In this chapter we use the term TPs. While we often know a lot about the environmental properties, fate and effects of the parent compounds, we know much less about the TPs. TPs can behave very differently from the parent compound. For example, selected TPs are much more persistent than their associated parent compound in soils, waters and sediments and some may be transported around the local, regional and Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

650 Transformation Products of Emerging Contaminants in the Environment

global environments to a different extent than the parent compound. TPs can also have very different toxicities or endocrine disruption activity than the parent compounds. The environmental risks of TPs can, therefore, be very different than the risks of the parent compound. However, the main problems concerning the TPs of the EDCs in the environment are the lack of knowledge of their potential impact on wildlife and human compared to their parent compounds, and the inadequate studies of their environmental fate and monitoring of their levels in the environment. The identification and characterization of TPs arising from a particular parent substance in a particular system can also be extremely difficult due to problems of extraction, detection at environmentally relevant levels, and quantification in the absence of standards; however, the arrival of new analytical methodologies (e.g., time-of-flight, orbitrap mass spectrometry) and the availability of expert systems for predicting transformation pathways is now making this task less daunting. This chapter attempts to provide an overview of the current studies concerning the formation, detection, occurrence and effects of the main TPs of some classes of EDCs in the environment, indicating, at the same time, the gaps in the existing knowledge and future actions that need to be taken. The chapter focuses mainly on the results of a series of monitoring studies into the occurrence in water bodies of selected TPs of the main substances included in the EU EDCs draft list, as well as their fate after their introduction into waste water treatment plants (WWTPs).

22.2 Endocrine Disrupting Compounds (EDCs) of Concern 22.2.1 Definitions and Regulatory Issues A number of definitions of EDCs were developed in the late 1990s by national and international governmental agencies. According to the US Environmental Protection Agency (EPA) task force on endocrine disruption (EDSTAC), in May 1997 [1]: “An endocrine disruptor is an exogenous chemical substance or mixture that alters the structure or function(s) of the endocrine system and causes adverse effects at the level of the organism, its progeny, the populations, or subpopulations of organisms, based on scientific principles, data, weight-ofevidence, and the precautionary principle.” However, the European scientific and regulatory community has agreed on the following definition of an endocrine compound during the Weybridge Conference [2]: “An endocrine disruptor is an exogenous substance that causes adverse health effects in an intact organism, or its progeny, consequent to changes in endocrine function.” More recently, the International Programme for Chemical Safety (IPCS – which involves WHO, UNEP and ILO) has, together with Japanese, US, Canadian, OECD and European Union experts, developed a consensus working definition for endocrine disrupters (EDs) that was also adopted as a working definition in the European Community Strategy for Endocrine Disrupters (International Programme on Chemical Safety, 2002) [3]: “An endocrine disrupter is an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub)populations.” It is often overlooked that the Community Strategy as well as the WHO/IPCS definitions also made an effort to describe what a potential endocrine disrupter should be: “A potential endocrine disrupter is an exogenous substance or mixture that possesses properties that might be expected to lead to endocrine disruption in an intact organism, or its progeny, or

Assessment of the Occurrence and Fate of Transformation Products 651

(sub)populations.” In summary, currently available definitions of “endocrine disrupter” are either neutral in terms of specifying the toxicological relevance of the effects to be described, or they introduce the idea of adversity. In May 2005 international experts and scientists from different disciplines convened in Prague to discuss European research on EDCs, known as the cluster for research on EDs (CREDO). The results summarized in the Prague Declaration on Endocrine Disruption, reinforced concerns over the long-term consequences of exposure in humans and wildlife [4]. Three pieces of European Community legislation deal explicitly with EDs: The Plant Protection Product Regulation, PPPR (1107/2009); the chemicals regulation, REACH (1907/2006) and the new Biocidal Product Regulation, BPR (currently under negotiation) [5]. The PPPR stipulates that active substances, safeners and synergists with endocrine disrupting properties that may cause adverse effects in humans cannot be approved for use unless the exposure of humans under realistic conditions of use is negligible. However, the PPPR does not detail how EDs should be defined for the purposes of this regulation. The task of developing such criteria is the responsibility of the European Commission (EC) which is mandated to present draft criteria by 14 December 2013. Under the REACH regulation, EDs may be included under the authorization scheme if they are deemed to be Substances of Very High Concern (SVHC) according to Article 57 (f). When the regulation was enacted in 2006, it was recognized that there was limited scientific knowledge about the effects of EDs. Consequently, the EC is mandated with reviewing the provisions of REACH regarding EDs (Art 138 (7)) by 1 June 2013. The negotiation of the text of the Biocidal Product Regulation (BPR) between the Council and the Parliament has been finalized. It stipulates that substances considered as having endocrine disrupting properties that may cause adverse effects in humans or which are identified in accordance with Articles 57(f) and 59(1) of REACH shall not be approved, unless the risk to humans is negligible. The EC is required to specify the scientific criteria for the identification of biocides with endocrine disrupting properties no later than 13 December 2013. EDCs encompass a diverse group of compounds, including pesticides, personal-care products (PCPs), steroids and hormones, surfactants, perfluorinated compounds (PFCs), flame retardants, industrial additives and agents, and gasoline additives, as well as some TPs. The European Commission’s Department of Environment along with a panel of experts has produced an EU-Strategy for EDs list of 66 substances with classification of high, medium or low exposure concern [6]. Generally, chemicals with hormonal activity include: Natural hormones from any animal, released into the environment, and chemicals produced by one species that exert hormonal actions on other animals, for example, human hormones unintentionally reactivated during the processing of human waste in sewage effluent, may result in changes to fish. Natural chemicals including toxins, produced by components of plants (the so-called phytoestrogens), and certain fungi. Synthetically produced pharmaceuticals that are intended to be highly hormonally active, for example, the contraceptive pill and treatments for hormone-responsive cancers may also be detected in sewage effluent. Man-made chemicals and TPs released into the environment. These include some pesticides (e.g., DDT and other chlorinated compounds), chemicals in some consumer and medical products (e.g., some plastic additives), and a number of industrial chemicals (e.g., polychlorinated biphenols (PCBs), dioxins). The hormonal activity of these chemicals, is many times weaker than the body’s own naturally present hormones, for example, nonyl phenol (a breakdown product of alkylphenol ethoxylate surfactants), has an estrogenic activity only about one-ten thousandth that of the natural hormone, estrogen.

652 Transformation Products of Emerging Contaminants in the Environment

22.2.2 Mechanisms of Endocrine Disruption The main evidence suggesting that exposure to environmental chemicals can lead to disruption of endocrine function comes from changes seen in a number of wildlife species. There is also some limited evidence in humans that adverse endocrine-mediated effects have followed either intentional or accidental exposure to high levels of particular chemicals. EDCs can act on the endocrine system to disturb the homeostatic mechanisms of the body or to initiate processes at abnormal times in the life cycle. The endocrine system is a complex network of glands, hormones and receptors. It provides the key communication and control link between the nervous system and bodily functions, such as reproduction, immunity, metabolism and behavior. The endocrine system consists of a set of glands, which secrete hormones – the chemical messengers carried by the blood. Endocrine glands include the thyroid gland, testes, and ovaries. Hormones, transported at concentrations often as low as parts per trillion or less, circulate around the body via the blood stream and modulate cellular or organ functions by binding with receptors in the target cells. Well-known hormones include estrogen and testosterone, thyroxin and insulin. Receptors in the target cells, once activated by binding of the hormone, regulate the functions and processes in the tissue through interactions with the cell’s DNA or other complex intracellular signaling processes. There are many ways in which EDCs can upset hormone signaling: (i) they may mimic the biological activity of a hormone by binding to a cellular receptor (agonistic effect); (ii) they may bind to the receptor preventing the binding of the natural hormone (antagonistic effect); (iii) they may bind to transport proteins in the blood, thus altering the amounts of natural hormones that are present in the circulation; or (iv) they may interfere with the metabolic processes in the body, affecting the synthesis or breakdown rates of the natural hormones. Up to now, because of a series of observations in both humans and wildlife the spotlight has focused on disruption to those hormones that play a major part in the control of reproduction and development. Thus, the effect of the EDCs mode of action is mainly categorized based on the main sex steroids, such as estrogens and androgens. Estrogens comprise a group of chemicals of similar structure mainly responsible for female sexual development and reproduction, with the principal human estrogen being 17b-estradiol. Testosterone is the principal human androgen, chemicals responsible for the development and maintenance of the male sexual characteristics. EDCs can cause endocrine disruption through a range of mechanisms by acting as: (i) environmental estrogens, for example, methoxychlor, bisphenol A; (ii) environmental antiestrogens, for example, dioxin, endosulfan; (iii) environmental antiandrogens, for example, vinclozolin, DDE, (iv) toxicants that reduce steroid hormone levels, for example, fenarimol, endosulfan; (v) toxicants that affect reproduction primarily through effects on the central nervous system, for example, dithiocarbamate; and (vi) others (e.g., altering thyroid hormone levels, aromatase activity).

22.3 Environmental Fate and Transformation of EDCs EDCs encompass a variety of chemical classes that are often pervasive and widely dispersed in the environment. Some are persistent, can be transported long distances across national boundaries, and have been found in virtually all regions of the world. Others are rapidly degraded in the environment or human body, or may be present for only short periods of time but at critical periods of development.

Assessment of the Occurrence and Fate of Transformation Products 653

Chemical substances enter the environment in different ways. Pesticides are released at their point of application; industrial chemicals are unintentionally released by volatilization, leaking or leaching, either during a product’s lifetime or after ultimate disposal. Natural hormones are excreted by various organisms and enter environmental compartments directly or after they have passed through wastewater treatment plants. Once a substance has passed through the environment, it can undergo different fates, such as: further distribution between the environmental compartments water, air, and soil/sediment, as well as their subcompartments, or degradation and transfer processes in the compartments and subcompartments. Most risk assessment studies and some epidemiologic studies have looked at the exposure and toxicology of a single compound. However, two other considerations must be included: the presence of TPs of EDCs and the cumulative exposure to pesticides multiresidue. In the following sections, the first topic of the fate of selected potential EDCs will be reviewed in general, or, if appropriate, in more detail. The occurrence and whether transformation or TPs of native compounds exhibit endocrine-disrupting characteristics will be considered. As useful tools for predicting and understanding the behavior of chemicals in the environment, their physicochemical properties can be used. The most important parameters include water solubility, adsorption coefficient and bioconcentration factor. The major sources of environmentally-relevant emerging contaminants (ECs) are primarily WWTP effluents and secondarily terrestrial runoffs (from roofs, pavements, roads and agricultural land), including atmospheric deposition. Once released into the environment, ECs are subject to processes (e.g., biodegradation, and chemical and photochemical degradation) that contribute to their elimination. Depending on the compartment in which synthetic chemicals are present in the environment (e.g., groundwater, surface water, and sediment) or in the technosphere (e.g., WWTPs and drinking-water facilities), different transformations can take place, sometimes producing products that can differ in their environmental behavior and ecotoxicological profile [7]. TPs include both metabolites excreted via urine or feces, and degradation products that can be formed by physico-chemical and biological processes in WWTPs or water works and/or in the environment, from parent EDCs and/or their metabolites released. It is undisputed that, in some cases, TPs of some contaminants are more persistent and can exhibit greater harmful effects than their parent compounds. For example, one study showed that, at the organism level, the only sublethal effect seen was an increase in heart rate at low concentration and a decrease at higher concentration with the use of aldicarb-sulfoxide but not with aldicarb [8]. Another study reported that the oxons of methyl-parathion, chlorpyrifos and diazinon were 15 to 10 times more toxic (to sperm DNA) than their corresponding parent compounds [9]. As another example, in vitro studies confirmed that 2,4-dichlorophenoxyacetic acid (2,4-D), a commonly used organophosphate herbicide promoting the proliferation of androgensensitive cells, is a known estrogen receptor ligand [10]. Also, the major biodegradation product of nonylphenol ethoxylates, nonylphenol, is much more persistent than the parent compound and can mimic estrogenic properties [7]. While substantial data exist regarding the occurrence and ecotoxicology of parent pharmaceuticals, much less is known about their metabolites and TPs, whose existence is even unknown in most cases. It is not necessary for EDCs to be present in high concentrations for them to have a significant effect. According to Dr. Palanza [11], two significant conceptual shifts have emerged by recent research on EDs different to traditional toxicological approaches. First, extremely low doses can cause measurable and substantial disruption in animal models and, secondly, consistent findings during the lifecycle of an organism, show that “developmental stages are

654 Transformation Products of Emerging Contaminants in the Environment

typically far more vulnerable to signal disruption than adult stages.” These facts generate the necessity for the determination of the EDCs and their TPs at very low levels in the environment, leading to new, accurate and extremely sensitive analytical techniques.

22.4 Analytical Methodology In the field of EDCs, the majority of methods applied until now have focused mainly on the determination of unchanged (parent) compounds; however, when monitoring environmental waters, the occurrence of TPs can often be higher than that of unchanged compounds. Due to their mobility in the soil–water environment, TPs can reach groundwater more easily than their parent compounds. Therefore, the most relevant TPs must be gradually incorporated into the analytical methods in order to have a wider knowledge of the water quality. The inclusion of TPs in multiresidue methods is an analytical challenge for several reasons [12]: many of these compounds are still not well known; the number of potential analytes to be investigated in water would increase drastically; they are usually more polar than the parent compounds, and their extraction (isolation from the water matrix)/preconcentration is more difficult; and the commercial availability of reference standards is rather limited. Additionally, adding more compounds in triple-quadrupole based methods would lead to an increase in the number of monitored transitions, which would produce either a decrease in sensitivity or lower peak definition. Moreover, one of the main problems in the development of quantitative methods for TPs is their higher polarity in relation to the parent compound, which makes the preconcentration step by the usual approaches troublesome. In addition, identification and determination of EDCs, and consequently their TPs, is a challenging task because of the extremely low levels at which they are present in the environment (nano- or picograms per liter). In the last 10 years, many techniques have been applied to the extraction and analysis of EDCs and their TPs in environmental samples. Regarding the extraction of water samples, the most commonly used methods are liquid–liquid extraction (LLE), solid-phase extraction (SPE) and solid-phase microextraction (SPME). LLE and SPE can be used to determine a broad range of EDCs in one analysis (see Tables 22.1–22.3). SPE methods are rapid, efficient (good recoveries and low detection limits), use less solvent than LLE and consequently have lower laboratory expenses [13]. In addition, SPE methods can be automated by using laboratory robotic systems [12,14,15] that do all or part of the sample preparation steps. The most widely used methods for the analysis of EDCs and TPs are based on gas chromatography (GC), liquid chromatography (LC), and capillary electrophoresis (CE). For GC and LC, mass spectrometry (MS)-based detection techniques have become powerful instruments because of their selectivity and sensitivity. Previously, GC-MS was most commonly used for analysis of EDCs and TPs. Concretely, GC-MS or GC-MS2 [16–20] is a very efficient technique with a high resolution power for volatile and semivolatile compounds, especially when using the selected-ion monitoring (SIM) mode, which eliminates interferences and improves detection limits. However, due to the need for derivatization [21,22], in consideration of the lower sensitivity of GC-MS when compared with LC coupled with ion trap (IT) or triple quadrupole analyzers (QqQs), and due to the physico-chemical characteristics of TPs (normally higher polarity and solubility in water than parent compounds), LC-MS or LC-MS2 has gradually become one of the most suitable techniques for analysis of TPs in environmental samples [23–26]. The excellent sensitivity and selectivity of LC-MS2 in Selected or Multiple Reaction Monitoring (SRM or

Assessment of the Occurrence and Fate of Transformation Products 659 Table 22.3 Occurrence of the main TPs of 17b-estradiol in WWTP and environmental samples worldwide. Area/Sample types

TPs of 17b-estradiol Levels (ng/L)

Spain, samples at different stages of the WWTP process 23 European Countries, groundwater China, river water

E1 E3

UK, East Sussex, UK, river water

E1

E3 E1 E1 16a-Hydroxyestrone

US, WWTP

E1 E3

Japan, WWTP

E1 E3 E1-3-sulfate E1-3-glucuronide E2-3-glucuronide E2-3-sulfate E2-3,17-disulfate E3-3-glucuronide E3-3-sulfate

USA, WWTP

E1-3-sulfate E1-3-glucuronide E2-3-glucuronide E2-3-sulfate

Korea, WWTPs M-WWTPs L-WWTPs H-WWTPs P-WWTPs

E1 E3 E1 E3 E1 E3 E1 E3

Spain, WWTPs

E1 E3

Taiwan, river water

E1 E3 E1 E3

Instrumentation

Ref.

Cpt/Cst/Cfe 15.5/13.4/11.6 22.2/18.7/16.1 4a

SPE-LC-ESI-MS/ MS

[24]

SPE-LC-MS/MS

[23]

1.0–1.6 1.6–5.2 U/S/Db >> chlorine-based oxidants. However, discrepancies have been observed, depending both on the type of toxin that is oxidized and the water quality parameters. In cyanobacteria bloom events, when GSM and MIB are also released, removal processes are required. Common disinfectants and oxidants, such as Cl2, ClO2 and KMnO4 are shown to be ineffective in their removal [95,96]. On the contrary, removal of GSM and MIB from water has been achieved using ozone [97,98], rendering ozonation an effective water treatment process, with varying degrees of success depending on the operational parameters used [99,100]. In the subsections below, degradation by-products, toxicity of treated water and reaction pathways of chemical oxidation of cyanobacterial metabolites are discussed.

688 Transformation Products of Emerging Contaminants in the Environment

23.3.2.1 Chlorination Chlorine is the most common reagent used in water treatment as both oxidant and disinfectant. Chlorine can be applied for water treatment in different forms, such as Cl2, sodium hypochlorite (NaClO), chlorine dioxide (ClO2) and chloramines (NH2Cl, NHCl2). A number of studies used free chlorine for MCs oxidation and indicated that the process was effective [101–113]. Chlorine-based weaker oxidants, such as chlorine dioxide and chloramines, have been used as alternatives to chlorine, but were found to be ineffective for MCs elimination [108,111,114]. Ho et al. studied the differences in the chlorine reactivity of four MC analogs; the oxidation of the MCs was related to the chlorine exposure, with the ease of oxidation following the trend: MC-YR > MC-RR > MC-LR > MC-LA [112]. This trend, attributed to differences in the structure of the toxins functional groups, was in agreement with differences in the reactivity of amino acids shown in previous model compound studies [115]. Tsuji et al. [116] examined the effect of sodium hypochlorite on MC–LR and -RR and were the first to report by-products of MC-LR during chlorination. Several reaction products were observed by high-performance liquid chromatography coupled to a UV detector (HPLC-UV). Using LC–MS, four isomers were identified as dihydroxy-microcystin with m/z 1029. They proposed that dihydroxy-microcystin could be formed by chlorine addition on the conjugated diene of Adda, followed by a hydrolysis forming two hydroxyl moieties [87,116]. This was in agreement with the reduction of absorbance of Adda amino acid at 240 nm and the subsequent loss of toxicity measured by PP-2A inhibition assay. A second by-product associated with m/z 1047 has been identified as the hydrated form of dihydroxymicrocystin [117]. In a more recent study, Merel et al. [107], identified four new compounds, that is, monochloro-microcystin, monochloro-dihydroxy-microcystin, dichloro-dihydroxymicrocystin, and trichloro-hydroxy-microcystin, using the LTQ-Orbitrap technology for accurate m/z measurements and tandem mass spectrometry (MS2) experiments. Identified products are presented in Figure 23.6. Merel et al. [118] have also studied the oxidation of CYN by chlorination. LC-MS analysis was used for identification of chlorination by-products. The formation of three byproducts was reported: 5-chloro-cylindrospermopsin and cylindrospermopsic acid which was

Figure 23.6 Chlorination by-products of microcystins (MCs).

Transformation Products of Hazardous Cyanobacterial Metabolites in Water 689

first identified by Banker et al. [119], and one stable chlorination by-product that has not been previously reported. The third CYN by-product characterized by its m/z ratio equal to 375.097 was related to the chemical formula C13H18N4O7S. Toxicity tests of MC solutions treated with chlorine have been carried out [101,109]. Senogles-Derham et al. using the heterozygous P53 transgenic mouse model to investigate potential cancer-inducing effects from such oral dosing solutions, demonstrated that removal of cyanotoxins by chlorination did not result in an increased incidence of cancer in mice that were genetically predisposed to the effect of genotoxic carcinogens [109]. Also the protein phosphatase 1 inhibition assay (PPIA) was applied to assess the toxicity of chlorination products of MC-LR and MC-RR [101]. Samples containing oxidation products were fractionated by HPLC and the fractions’ toxicities were tested with PPIA. The results indicate that protein phosphatase 1 inhibition resulted only from intact MC and the oxidation products were non-toxic. Similar results were obtained by experiments carried out in natural water samples, also demonstrating that reaction products were non-toxic [101]. The cytotoxicity of both pure CYN (before chlorination) and the mixture of by-products (after chlorination) on human intestinal Caco-2 cells according to MTT and NRU assays were assessed. Both of the mitochondrial and lysosomal activities measured revealed that CYN chlorination significantly decreases mixture cytotoxicity [118]. Even though they have been less investigated, saxitoxins and nodularins are also altered by chlorine. For these toxins, no by-products have been identified, but the chlorinated mixture does not show acute toxicity. On the contrary, the fact that anatoxin-a has very slow reaction kinetics suggests that this toxin resists chlorination [102]. 23.3.2.2 Oxidation with Permanganate In the 1960s, application of potassium permanganate began in the water industry having many potential uses as an oxidant, destroying micro-organisms and organic pollutants [120]. Permanganate in general attacks functional bonds with multiple bonds and may oxidize organic compounds via several reaction pathways, primarily electron exchange, hydrogen atom abstraction, hydride ion abstraction or oxygen donation. In an early study, Rositano et al. [121], and in a few more recent studies [108,122– 124], it was reported that oxidation of MC-LR with permanganate appeared to be an effective process for its removal in water samples. Oxidation of MC-LR, -RR and -YR with potassium permanganate in natural surface water was also investigated and rate constants for the reactions were determined [108]. Ding et al. [122] studied the release and removal of MCs from microcystis with different oxidants. Results showed that permanganate had a much slower reactivity than ozone. In comparison with free chlorine, permanganate had a higher reactivity for most of the MC species. The main concern using permanganate for cyanobacteria and cyanotoxin treatment is the formation of colloidal MnO2 which must be removed. Oxidation of the MC variants by permanganate seems to proceed in a way similar to that of alkenes reaction with permanganate, via hydroxylation and diol formation, due to the diene function group of Adda [123,124]. Permanganate ion reacts with the conjugate double bonds of Adda forming a cyclic hypomanganate ester intermediate, which is rapidly hydrolyzed to form dihydroxy-MC [87]. This oxidation mechanism was supported by Huang et al. who have studied the reaction of MC-LR with permanganate and have

690 Transformation Products of Emerging Contaminants in the Environment

Figure 23.7 TPs of MC-LR using KMnO4.

determined dihydroxy-MC formed, using LC-MS [106]. This was also supported by the reaction’s stoichiometry, according to which 2 moles of permanganate are consumed per mole of reacted MC-LR [123]. Chen et al. [124] investigated the oxidation of MC-RR with spectroscopic analysis. Based on the spectral changes during the course of the reaction (increase in absorbance in the wavelength ranges 200–220 and 340–360 nm), suggested the formation of intermediate products upon MC-RR oxidation. Concerning the increase of absorbance near 300 nm, it could be due to the formation of 2-methyl-3methoxy-4-phenylbutyric acid (MMPB) from the MC-RR oxidation which contains a carboxylic acid group exhibiting absorption at this area of the spectrum. Wu et al. [125] investigated the kinetics of MMPB formation by oxidation of MC-LR with permanganate–periodate. The same group, using the same technology and employing HPLC with diode array detection, indicated that a minimum 85% yield of MMPB was achieved [126]. Even though an approach to identification of TPs formed during MCs reaction with permanganate has been performed, the complete oxidative pathways have not yet been elucidated. The oxidation TPs of MC-LR reaction with permanganate are presented in Figure 23.7. Toxicity tests of MC-LR and -RR treated with permanganate were performed using the protein phosphatase 1 inhibition assay (PPIA). Results showed that residual toxicity was associated only with MCs, while reaction products were non-toxic [101]. Investigation of the toxicity of intermediate products formed in natural waters showed no interactions between the matrix components with the MCs or their oxidation products, supporting the feasibility of applying permanganate to remove MCs in natural waters. 23.3.2.3 Ozone The ozonation process is one of the most affordable treatment technologies that can be used for removal of cyanotoxins from drinking water. Ozone is a strong oxidizing agent that proceeds through direct molecular reaction pathway and/or indirect radical formation [127]. A number of studies have been performed dealing with the ozonation of cyanotoxins and they mainly focused on the investigation of doses required for toxins destruction [108,128–136], the effect of water quality parameters [108,129,131–134] and

Transformation Products of Hazardous Cyanobacterial Metabolites in Water 691

Figure 23.8 Ozonation TPs of microcystins (MCs).

the mechanistic aspects of ozone’s reactivity with MCs [131]. The majority of those focused on MC-LR degradation [108,128–134] and a few on MC-RR [133,135,137] and MC-LA [129,132]. Results indicated that ozonation as a treatment process for cyanotoxins is highly efficient but only a few studies dealt with the TPs formed [135,137]. In general, ozonation in comparison with the chlorination process produces less harmful TPs. Investigating the reaction intermediates of MC-LR and MC-RR with O3, two studies were performed in which they associated the formation of TPs to ozone doses [135,137]. When high ozone doses were applied, cyanotoxins degraded into small molecular weight TPs (below 450 amu) with linear structures, while with low ozone doses larger molecular weight intermediates were formed (510  m/z  837). The majority of the TPs identified were related to the oxidation of the Adda amino acid and its conjugated bonds. Hydroxyl addition and oxidative cleavage to smaller moieties was observed. Hydroxyl substitution was also observed on the aromatic ring of the Adda group. In addition a secondary pathway was observed, consisting of ring opening at the Mdha-Ala peptide bond [135,138]. Ozonation TPs of MCs degradation are presented in Figure 23.8. Toxicity bioassays (Microtox) were performed on the treated solutions in order to evaluate the toxicity effect of ozonation. Toxicity was found to be related to the ozone dose, since the Adda moiety, which is linked to the toxic properties of MC-LR and -RR, can be oxidized faster by higher doses of ozone [137,138]. Toxicity measurements were also performed using protein phosphatase inhibition in vitro and mice model in vivo to evaluate the safety of the TPs. It was found that ozonation eliminates the toxic properties of MC-LR and MC-RR, showing that O3 treatment of cyanotoxins is potentially a suitable detoxification technology [135,137]. Ozone, as was previously reported, being one of the strongest oxidants is capable of degrading T&O compounds. In several studies the ozone process was applied for GSM/MIB removal in combination with other methods, such as UV irradiation or H2O2. Duguet et al. [139] observed that GSM and MIB removal by ozonation was influenced by solution pH, and with addition of H2O2, degradation was 10-fold faster than with ozone alone. A number of studies [140] presented different results on the destruction of GSM and MIB with O3, attributed to the nature of the water studied in each case. On the contrary,

692 Transformation Products of Emerging Contaminants in the Environment

Figure 23.9 Some of the TPs identified during the ozonation of MIB [142].

Liang et al. [141] found that the presence of background organics have no significant effect on the ozonation of GSM/MIB. They also confirmed the significance of pH as a key factor since it is directly related to the concentration of hydroxyl radicals produced during the process. Concerning the TPs of GSM and MIB formed after application of ozone, two studies have been performed to date [142]. Qi et al. [142] applied ozonation in order to investigate the degradation of MIB, proposing a transformation pathway. Camphor (I), (Figure 23.9) was likely to be the primary TP which further oxidized to form other TPs, such as ketones (II, III), an alcohol (IV), and a carboxylic acid (V), (Figure 23.9). Those products were further degraded into other TPs, with aldehydes being the main intermediates of ozonolysis of MIB.  Additionally, hydroxyl radical experiments indicated that OH radicals accounted for the degradation of MIB. The majority of the products identified in this study were identical with those identified by Li et al. [143] using ozonation with the addition of H2O2 as the water treatment technology, indicating the synergistic action between O3 and H2O2. 23.3.3 Advanced Oxidation Processes Among all kinds of water purification techniques, oxidative degradation methods, using electromagnetic radiation have been receiving increased attention for the detoxification of the aquatic environment. They are generally referred to as advanced oxidation processes (AOPs) and mainly involve UV light in the presence of hydrogen peroxide or ozone, and UV and near-visible light in the presence of TiO2 [144–148]. These methods lead to mineralization of the organic pollutants and are mainly based on the generation of  electron/hole pairs and highly oxidizing OH radicals. In the subsections below, TPs, the toxicity of treated water and the reaction pathways of advanced oxidation of cyanobacterial metabolites are discussed. 23.3.3.1 Titanium Dioxide Photocatalysis TiO2 photocatalysis is among the most efficient emerging processes for the treatment of cyanotoxin contaminated water [94,138,149,150]. This “green” emerging process

Transformation Products of Hazardous Cyanobacterial Metabolites in Water 693

is known to efficiently perform water purification, disinfection and detoxification without necessitating the use of or resulting in the production of hazardous compounds. [146,151,152] Upon irradiation, TiO2 is activated and forms reactive oxygen species   (ROS) including hydroxyl radicals ( OH, main specie), superoxide radical anion (O2 ),  and perhydroxyl radical (HO2 ), as well as the conduction band electron (ecb) [146,153]. Though different types and variants of cyanotoxins have been treated with TiO2 photocatalysis, such as microcystins (microcystin-LR, -RR, -LA and –YR) [151,152,154–161], nodularins [162], cylindrospermopsin [163,164], only a few studies have dealt with the intermediates formed during treatment. These studies concern the hepatotoxins microcystin-LR (MC-LR) [151,154–156], nodularin [162], and cylindrospermopsin [164]. Recently, the complete photocatalytic mineralization of MC-LR to CO2 and inorganic ions (NO2, NO3 and NH4þ) has been reported under both UV-A and solar light in the presence of TiO2-based photocatalysts [161]. TiO2 photocatalysis has also been used for the removal of GSM/MIB. Lawton et al. [165] reported rapid degradation of both GSM/MIB with more than 99% removal within 60 min of illumination in the presence of suspended TiO2. Bellu et al. [166] using a pellet form of TiO2 completely removed GSM within 25 min of treatment. A preliminary study was also performed by Pemu et al. [167] using TiO2 photocatalysis of GSM in which few intermediate products were identified. Microcystin-LR MC-LR is the most studied derivative of the hepatotoxic microcystins and among the different cyanotoxins groups (cyclic peptides, alkaloids and liposaccharides). Its frequent appearance in water resources, solubility and stability comprise the major reasons [149]. Photocatalysis in the presence of TiO2 appears to degrade MC-LR effectively through the formation and decay of several intermediates prior to mineralization [161]. In a few studies the transformation pathways of MC-LR with TiO2/UV-A [151,154–156] and NTiO2/(l > 400 nm) photocatalysis have been presented [157]. The first studies of the transformation pathways of MC-LR with TiO2 photocatalysis in slurries was by Liu and coworkers [151,156]. Based on the experimental conditions (CMCLR ¼ 1000 mg/L; 1.0% w/v TiO2; 42.8 mM H2O2; pH ¼ 4.0) ten reaction intermediates (651  m/z  1029) were identified and organized into three oxidation pathways. Initially, the chain of the Adda amino acid was oxidized by the ROS, with the conjugated carbon double bonds undergoing simultaneous hydroxylation and isomerization to form the m/z ¼ 1029.5 products, (Figure 23.10). Subsequent oxidation and cleavage steps of the hydroxylated C6C7 and C5C4, resulted in the formation of smaller molecular weight ketones (m/z ¼ 835.5) and aldehydes (m/z ¼ 795.5), respectively (Figure 23.10). Further oxidation of the latter compound produced the carboxylated intermediate with m/z ¼ 811.5. The cyclic structure of MC-LR was also affected at the peptide bond between Mdha and Ala, which cleaved (C3O) to produce [M þ Na ¼ 965]. Hydroxylation of the C6C7 double bond of the Adda occurred simultaneously with its linearization to form the m/z ¼ 977 compound. Subsequent oxidation resulted in the corresponding ketone (m/z ¼ 783), aldehyde (m/z ¼ 743) and carboxylic acid (m/z ¼ 759). During the last pathways a highly oxidized linear MC-LR with the Mdha-Ala-MeAsp moities removed (m/z ¼ 617), underwent dihydroxylation of the Adda chain (m/z ¼ 651). Toxicological tests based on the inhibition of PPIA have shown reduced toxicity of the toxin solution as a whole and its fractions with treatment time. This verifies that the

694 Transformation Products of Emerging Contaminants in the Environment R1

R1

HN

HN O

CH3 OH OH

CH3

O

O

NH OH

R2 HO C

N

HN

OH

O R4

O NH O

OH NH

NH

R2

O

OH

O

O

NH OH

R2

O

O

R1 HN O

O

NH O

H3C

OH

R3 R2

O

m/z = 1029.5 C49H76N10O14 Mass: 1028.6 (a), (b), (c), (d) R1

R1 OH O

CH3

O

O

OH

O

O

O

R2

O NH

O HO

O

O

C

R2

O N

N

NH

R4

O

NH

HO

O

C

O

NH R2

O

HN O

NH

HO

R1

HN

HN

HN O

O O

NH

NH

NH

R4

R3

R3

NH

R2

O

O

R2

O

m/z = 1011.5 C49H74N10O13 Mass: 1010.5 (c), (d) R1

HN

HN

O

O O

O

H

R2

O

m/z = 835.5 C37H58N10O12 Mass: 834.4 (a), (b), (c), (d) HO

C

HO

O

O C

O N

NH

m/z = 795.5 C34H54N10O12 Mass: 794.4 (a), (b), (c), (d) HO

O

HN O

NH R2

NH O

O

NH HN

NH CH2

O

O

C NH O C

NH

NH CH3 CH3

O

O O

NH O

O

O

C NH O C

NH

NH CH

3

CH3

OH

NH HN

NH2

O

O

OH

C

R1

NH NH2

m/z = 783.4 C34H58N10O12 Mass: 782.4 (b), (c)

C HN

NH2

C34H58N10O11 Mass: 783.4 (a), (c), (d)

Figure 23.10 TPs identified during photocatalytic degradation of MC-LR as reported by: (a) Liu et al. Reproduced with permission from [156] Copyright (2003) American Chemical Society, (b) Choi et al. Reproduced with permission from [157] Copyright (2007) American Chemical Society, (c) Antoniou et al. Reproduced with permission from [154] Copyright (2008) Elsevier Ltd, (d) Antoniou et al. Reproduced with permission from [168] Copyright (2010) Elsevier Ltd.

Transformation Products of Hazardous Cyanobacterial Metabolites in Water 695

linearized TPs, as well as TPs with a hydroxylated and isomerized Adda chain, do not present any toxicity [151,156]. The other study on TiO2/UV-A is by Antoniou et al. [155], in which the authors utilized two different TiO2 photocatalytic films (mTiO2 ¼ 1.4 mg per thin film and mTiO2 ¼ 50.4 mg per thick film) for the degradation of MC-LR (Co ¼ 20 mg/L) at neutral pH and detected 21 types of TPs. The most important of these TPs are presented in Figure 23.10. The significantly higher number of TPs isolated in this study [155] compared to others [151,156] is mainly attributed to the experimental conditions. The absence of additional oxidants and neutral pH reduced the strong interaction between the toxin and the catalyst and allowed the  formed OH to interact with all the sites of the toxin [152,168]. Overall, the transformation of MC-LR was found to be initiated at four different sites: three on the Adda amino acid (aromatic ring, methoxy group and conjugated double bonds) and one on the cyclic structure (Mdha amino acid). At first, the aromatic ring underwent hydroxyl substitution of aromatic hydrogen to form the m/z 1011.5 TP, (Figure 23.10). This TP together with m/z ¼ 1029.5 (Figure 23.10) and 1063.5 had multiple peaks on the total ion chromatogram (TIC) attributed to their geometrical or constitutional isomers. In the case of m/z ¼ 1011.5, the o-, p-, and to a lesser extent the m-hydroxylated TPs were formed. A second hydroxylation of the aromatic ring followed, which yielded the m/z ¼ 1027.5 TPs. The detection of m/z ¼ 1027.5 serves as a confirmation of the first substitution, since addition of hydroxy groups into the aromatic ring is known to increase the ring’s electron density. This results in the acceleration of the reaction rates of additional electrophilic reactions, as is hydroxyl radical attack [169]. The methoxy group of the Adda chain can be completely removed to produce the m/z ¼ 965.6, DmADDA (demethoxylated-MC-LR), through an intermediate step, where a formic acid ester (m/z ¼ 1009.6) is formed. This pathway was not reported when TiO2 particles were used.  The most targeted site for OH attack was the conjugated diene bonds, because of their  location in the MC-LR structure and their susceptibility to oxidation by OH (kOH  10910 1 1 M s ) [170]. The diene bonds produced hydroxyl adducts through two different steps: hydroxyl addition (m/z ¼ 1029.5 and m/z ¼ 1063.5) and hydroxyl substitution (m/z 1011.5). The last one was the enol-MC-LR (m/z 1011.5) and initiated the complete removal of the Adda chain. The enol-MC-LR rapidly isomerized to the more stable tautomer of ketoneMC-LR (m/z ¼ 1011.5) and, following a series of oxidative induced bond cleavage steps, it transformed to a ketone-derivative (m/z ¼ 835.4), to an aldehyde-derivative (m/z ¼ 795.4), and finally to a hydroxyl-derivative m/z ¼ 783.4, (Figure 23.10). The last site where degradation of MC-LR was initiated was the Mdha amino acid found in the cyclic structure. Consecutive oxidation steps, such as the double hydroxylation of the Mdha (m/z ¼ 1029.5), (Figure 23.10), its oxidation to aldehyde (m/z ¼ 1011.5), (Figure 23.10), and cleavage of the R2CCOR bond (m/z ¼ 1015.5), were observed. In addtion, intermediates where the degradation occurred simultaneously at the Adda chain and cyclic structure, such as m/z ¼ 783.4 (Figure 23.10), were observed. Incorporation of heteroatoms, such as N and F, into the TiO2 structure reduces the band gap energy of the semiconductor or provides mid-gap energy sensitization levels allowing visible light activation [157,158,160]. The ROS formed under these conditions differ from   UV-A activation and it is believed that the degradation occurs from O2  and HO2  radicals.  The latter ones react in a similar way to OH and can also produce hydroxylated TPs. Illumination of N-TiO2 nanoparticles with visible light, only produced multiple peaks of the

696 Transformation Products of Emerging Contaminants in the Environment

m/z ¼ 1029.5 [157]. However, when UV-A radiation was utilized to activate the N-TiO2 nanoparticles, besides m/z ¼ 1029.5, the oxidative cleavage of the Adda chain through the formation of a ketone (m/z ¼ 835.5), an aldehyde (m/z ¼ 795.5) and a hydroxyl derivative (m/z ¼ 783.5), (Figure 23.10), was also observed. Similar oxidation products for MC-LR have also been reported with other AOPs, such as ultrasonically induced degradation and g-irradiation [171,172]. Nodularin Nodularins is another family of cyclic polypeptides with five amino acids whose nine congeners also inhibit the proper function of protein phosphatases (PPs) [173]. The only photocatalytic oxidation study with TiO2/UV-A on the TPs formed during treatment is by Liu et al. [162] They utilized TiO2 nanoparticles in slurry (0.1%  w/v) and identified 10 TPs (175  m/z  859). Similar to MC-LR, OH decomposed the conjugated diene structure of Adda to form dihydroxylated products of different stereochemistry (m/z ¼ 859). Further into the photocatalytic degradation, the hydroxylated C4C5 and/or C6C7 bonds of Adda cleaved to form an aldehyde (m/z ¼ 665) and ketone products (m/z ¼ 625). These TPs were then oxidized to peroxidated products (m/z ¼ 695), followed by hydrolysis of the peptide bonds, resulting in small amino acid fragments (m/z ¼ 286; m/z ¼ 175). Similar to MC-LR [151,156], toxicity studies with the PP1 enzyme were conducted and showed loss of toxicity due to photocatalytic treatment, as well as the lack of formation of toxic TPs. Cylindrospermopsin Known from the red tides in the coast line of Australia, CYN is a tricyclic-guanidine alkaloid, with cytotoxic, dermatotoxic, genotoxic, hepatotoxic in vivo, and perhaps carcinogenic properties [174], To the best of our knowledge, a study presented in the 2008 ACE Meeting [164] is the only one on the TPs of CYN with reactive oxygen species (ROS) generated with UV-A/TiO2 films. Tandem mass spectrometry was used to identify the structures of the 23 TPs observed in the TIC chromatogram. Isobaric compounds were also detected as multiple peaks in the TIC, indicating the formation of stereoisomers or structural isomers [164]. Analysis of the MS/MS data, indicated the formation of hydroxylated compounds through substitution of unsaturated carbon bonds (m/z ¼ 432). The complete publication is still pending. Geosmin and 2-Methylisoborneol Titanium dioxide was also used for photocatalytic degradation of GSM and MIB. Lawton et al. [165] reported rapid degradation of both GSM/MIB with more than 99% removal within 60 min of treatment, using suspended TiO2 and complete removal of GSM within 25 min, using a pellet form of TiO2, respectively. Bamuza-Pemu and Chirwa [167] using TiO2 photocatalyst presented some preliminary results concerning TPs formed during GSM degradation. The TPs identified (3,5-dimethylhex-1-ene, 2,4-dimethylpentan-3-one, 2-methylethylpropanoate and 2-heptanal) showed that GSM undergoes rapid ring opening and subsequent bond cleavage at multiple cites, producing acyclic saturated and unsaturated compounds, with those not being of environmental concern. In a recent study Fotiou et al. [175] using TiO2 under UV-A irradiation, effectively degraded GSM/MIB. They also presented results for the mineralization of the compounds via total organic carbon (TOC) measurements. They identified a plethora of TPs of GSM

Transformation Products of Hazardous Cyanobacterial Metabolites in Water 697

produced during the photocatalytic procedure; the majority of them were cyclic ketones which upon ring opening formed acyclic saturated and unsaturated TPs. In this study [175], TPs of MIB formed during photocatalytic degradation using TiO2 under UV-A irradiation were also investigated. As a primary TP produced during photocatalysis of MIB, was proposed d-camphor; formed with a b-scission reaction mechanism on the methyl group, leading to the ketone. Mechanisms involved in other identified TPs are hydroxyl radical oxidation of compounds, driven by electrophilic substitution reactions. The majority of those are five-membered rings revealing ring opening on the MIB molecule. 23.3.3.2 Sulfate Radical-based AOPs (SR-AOPs) Another group of AOPs is based on the generation of sulfate radicals (SO4 ) which are  strong e acceptors with comparable oxidizing abilities to OH [168,170,176,177]. SRAOPs involve the activation of the oxidants persulfate (PS) and peroxymonosulfate (PMS) with catalysts, radiation and/or heat [168,178,179]. Up to now, to the best of our knowledge, Antoniou et al. [168,176], studied first the degradation of MC-LR with SR-AOPs and also identified the corresponding TPs. Specifically, the PS/UV (300 < l < 400 nm) system at neu tral pH was utilized, to avoid photolysis of the toxin and limit the formation of OH since  they produce similar TPs to SO4 . Tandem mass spectrometry indicated the formation of nine TPs, out of which four masses (m/z ¼ 1011.5, 1027.5, 1029.5, and 1045.5) had multiple peaks in the TIC. Stable sulfate adducts were not detected, since they transform to hydroxylated TPs (through substitution or addition). The mechanistic steps included bond isomerization and oxidation, combined with bond cleavage and the oxidative cleavage of small functional groups (e.g., COOH, m/z ¼ 999.5), up to the complete removal of the Adda chain  (m/z ¼ 783.4). The electrophilic character of SO4  was evident from the multi-hydroxylation of the aromatic ring of the Adda amino acid. Toward the end of the treatment, TPs were formed with the cyclic structure and the Adda chain was greatly oxidized [168]. 

23.3.3.3 Polyoxometalate Based AOPs (POM) Polyoxometalates (POM) are acid condensation products, mainly of molybdenum and tungsten [180–182], that upon excitation with near-visible and UV light become powerful oxidizing reagents capable of destroying a great variety of organic compounds in aqueous  solutions [183,184]. OH radicals generated by reaction of POM with H2O seem to play a key role in the process [185]. Oxygen oxidizes (regenerates) the catalyst and through reductive activation may or may not participate further in the process, depending on the substrate [186]. Due to their photocatalytic performance, POM can be recognized as an advanced oxidation process. POM are at least as effective as the widely published TiO2, presenting similar behavior [185]. Photocatalytic degradation of GSM and MIB under UV and near-visible light in the presence of POM photocatalyst, SiW12O404, in aqueous solution has been studied and compared with the photodegradation by TiO2 suspensions [175]. GSM and MIB are effectively degraded in the presence of both photocatalysts. TPs identified in the case of GSM and MIB, are mainly identical in the presence of both photocatalysts, suggesting that the photodegra dation mechanism takes place via a common reagent, that is, OH radicals.

698 Transformation Products of Emerging Contaminants in the Environment

23.4 Research Gaps, Recent Trends and Future Needs Cyanobacterial metabolites are a wide range of molecules that include some of the most potent toxins known and some very strong T&O compounds. The exact purpose of their biosynthesis is in most cases not yet fully understood, while significant progress has been made in the elucidation of genetic and biochemical mechanisms involved in their production. Toxic and odorous metabolites can be released in surface water and subsequently interact with naturally existing matter and organisms through complex chemical, biological and biochemical pathways. Understanding of the fate of these metabolites in the environment should be based on robust knowledge of the numerous parameters that can be used to predict their transport, persistence, reactions and toxicity under varying environmental conditions. Evolution of analytical methods based on mass spectrometry, that is, LC-MS and relative techniques, have provided significant tools toward the identification and detection of TPs of those metabolites. However, in many cases, for example, in the natural photolytic degradation of complex molecules such as microcystins, limited data are still available and it is clear that future studies should focus on processes such as electron energy transfer and singlet oxygenation to understand the underlying mechanisms. Further studies are also needed to understand the effect of natural organic matter on the biodegradation of metabolites and the associated pathways. Regarding the effects of disinfectants and oxidants in water treatment, the most studied cyanobacterial metabolites are the microcystin class, mainly MC-LR. The TPs of cylindrospermopsin have been explored but only under removal by chlorination. Additional studies using other removal methods (ozonation, oxidation by potassium permanganate, etc.) with respect to cylindrospermopsin remain uncompleted. Various other cyanotoxins, such as anatoxin-a, nodularin, saxitoxin, have still not been studied versus conventional water oxidation treatment and their metabolites are unknown. The story is similar for AOPs, such as photocatalysis by titanium dioxide, oxidation by sulfate radicals, and UV-H2O2. Most studies regarding AOPs have focused on microcystins, specifically MC-LR. To the best of our knowledge, there are no studies on identifying the TPs for either odor compounds or cyanotoxins. Worth noting is that there is currently no literature addressing the TPs of visible-light active titanium dioxide for any cyanotoxins. Heretofore, the trend for advanced oxidation of cyanobacteria metabolites has been to focus on kinetics, not on pathways or TPs. Another important issue regarding oxidation processes for water treatment is the toxicity of the metabolites and TPs that are produced. Toxic TP formation in water is a major concern for public health and may pose restrictions to water treatment processes. There is also a need, for practical reasons, to determine the point at which the oxidation process is complete, that is, the point at which there are no toxic compounds present in the water. For example, as has been demonstrated with microcystin, complete oxidation of the toxin is not needed to decontaminate water, since from the first oxidation step non-toxic metabolites are produced. This is directly relevant to the efficiency and, therefore, the cost-effectiveness and applicability of these processes. In conclusion, even with the significant increase in published research concerning cyanotoxins and cyanobacteria over the last decade, many aspects regarding the fate of cyanobacterial metabolites in the environment and during water treatment are not fully understood. More efforts are needed to completely elucidate their complex transformation pathways so that better strategies for human health and ecosystem protection can be developed.

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706 Transformation Products of Emerging Contaminants in the Environment 139. Duguet, J.P., Bruchet, A., and Mallevialle, J. (1989) New advances in oxidation processes: the use of the ozone/hydrogen peroxide combination for micropollutant removal in drinking water, International Water Supply Association (IWSA) “Organic Micropollutants”, Barcelona, September. 140. Terashima, K. (1988) Reduction of musty odor substances in drinking water – A pilot plant study. Water Science and Technology: A Journal of the International Association on Water Pollution Research, 20, 275–281. 141. Liang, C., Wang, D., Chen, J. et al. (2007) Kinetics Analysis on the Ozonation of MIB and Geosmin. Ozone: Science & Engineering, 29, 185–189. 142. Qi, F., Xu, B., Chen, Z. et al. (2009) Efficiency and products investigations on the ozonation of 2-methylisoborneol in drinking water. Water Environment Research: A Research Publication of the Water Environment Federation, 81, 2411–2419. 143. Li, X., Huang, Y., and Wang, D. (June (2010)) Efficiency and Mechanism of Degradation of 2-Methylisoborneol(2-MIB) by O3/H2O2 in Water, Bioinformatics and Biomedical Engineering (iCBBE), 4th International Conference, Chengdu, China. 144. Legrini, O., Oliveros, E., and Braun, A.M. (1993) Photochemical processes for water treatment. Chemical Reviews, 93, 671–698. 145. Kamat, P.V. (1993) Photochemistry on nonreactive and reactive (semiconductor) surfaces. Chemical Reviews, 93, 267–300. 146. Hoffmann, M.R., Martin, S.T., Choi, W. et al. (1995) Environmental applications of semiconductor photocatalysis. Chemical Reviews, 95, 69–96. 147. Malato, S., Fern
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Transformation Products of Hazardous Cyanobacterial Metabolites in Water 707 158. Pelaez, M., De La Cruz, A.A., Stathatos, E. et al. (2009) Visible light-activated N-F-codoped TiO2 nanoparticles for the photocatalytic degradation of microcystin-LR in water. Catalysis Today, 144, 19–25. 159. Pelaez, M., Falaras, P., Kontos, A.G. et al. (2012) A comparative study on the removal of cylindrospermopsin and microcystins from water with NF-TiO2-P25 composite films with visible and UV–vis light photocatalytic activity. Applied Catalysis B: Environmental, 121–122, 30–39. 160. Pelaez, M., Falaras, P., Likodimos, V. et al. (2010) Synthesis, structural characterization and evaluation of sol–gel-based NF-TiO2 films with visible light-photoactivation for the removal of microcystin-LR. Applied Catalysis B: Environmental, 99, 378–387. 161. Triantis, T.M., Fotiou, T., Kaloudis, T. et al. (2012) Photocatalytic degradation and mineralization of microcystin-LR under UV-A, solar and visible light using nanostructured nitrogen doped TiO2. Journal of Hazardous Materials, 211–212, 196–202. 162. Liu, I., Lawton, L.A., Bahnemann, D.W. et al. (2005) The photocatalytic destruction of the cyanotoxin, nodularin using TiO2. Applied Catalysis B: Environmental, 60, 245–252. 163. Senogles, P.J., Scott, J.A., Shaw, G. et al. (2001) Photocatalytic degradation of the cyanotoxin cylindrospermopsin, using titanium dioxide and UV irradiation. Water Research, 35, 1245–1255. 164. Antoniou, M.G., Choi, H., De La Cruz, A.A. et al. (2008) Intermediates of cyanobacterial toxins with hydroxyl-radical based advanced oxidation technologies (HR-AOTs), Ann. Conf. Exp. (ACE) Am. Water Works Assoc., Atlanta, GA, November, 2008. 165. Lawton, L., Robertson, P.K.J., and Bruce, D. (2003) The destruction of 2-methylisoborneol and geosmin using titanium dioxide photocatalysis. Applied Catalysis B: Environmental, 44, 9–13. 166. Bellu, E., Lawton, L.A., and Robertson, P.K.J. (2008) Photocatalytic Destruction of Geosmin Using Novel Pelleted Titanium Dioxide. Journal of Advanced Oxidation Technologies, 11, 384–388. 167. Bamuza-Pemu, E.E. and Chirwa, E.M.N. (2011) Photocatalytic degradation of geosmin: intermediates and degradation pathway analysis. Chemical Engineering Transactions, 24, 91–96. 168. Antoniou, M.G., De La Cruz, A.A., and Dionysiou, D.D. (2010) Intermediates and reaction pathways from the degradation of microcystin-LR with sulfate radicals. Environmental Science & Technology, 44, 7238–7244. 169. Song, W., De La Cruz, A.A., Rein, K. et al. (2006) Ultrasonically induced degradation of microcystin-LR and -RR: identification of products, effect of pH, formation and destruction of peroxides. Environmental Science & Technology, 40, 3941–3946. 170. Buxton, G.V., Greenstock, C.L., Helman, W.P. et al. (1988) Critical review of rate constants for  reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals OH/O2( in Aqueous Solution. Journal of Physical and Chemical Reference Data, 17, 513–886. 171. Song, W., De La Cruz, A.A., Rein, K. et al. (2006) Ultrasonically induced degradation of microcystin-LR and -RR: Identification of products, effect of pH, formation and destruction of peroxides. Environmental Science & Technology, 40, 3941–3946. 172. Song, W., Xu, T., Cooper, W.J. et al. (2009) Radiolysis studies on the destruction of microcystin-LR in aqueous solution by hydroxyl radicals. Environmental Science & Technology, 43, 1487–1492. 173. Hudnell, H.K. (2008) Cyanobacterial Harmful Algal Blooms: State of the Science and Research Needs, Springer, New York. 174. Newcombe, G. and Burch, M. (2003) Toxic blue-green algae: coming to a neighborhood near you? American Water Works Association, 29(1), 4–7. 175. Fotiou, T., Triantis, T.M., Kaloudis, T. et al. (2012) Photocatalytic Degradation of Taste and Odour Compounds in the presence of Polyoxomatalates: Degradation pathway and comparison with TiO2, Applied Catalysis B: Environmental, Submitted for publication. 176. Antoniou, M.G., De La Cruz, A.A., and Dionysiou, D.D. (2010) Degradation of microcystin-LR using sulfate radicals generated through photolysis, thermolysis and e transfer mechanisms. Applied Catalysis B: Environmental, 96, 290–298.

708 Transformation Products of Emerging Contaminants in the Environment 177. Neta, P., Huie, R.E., and Ross, A.B. (1988) Rate Constants for Reactions of Inorganic Radicals in Aqueous Solution. Journal of Physical and Chemical Reference Data, 17, 1027–1284. 178. Anipsitakis, G.P. and Dionysiou, D.D. (2004) Transition metal/UV-based advanced oxidation technologies for water decontamination. Applied Catalysis B: Environmental, 54, 155–163. 179. Anipsitakis, G.P., Dionysiou, D.D., and Gonzalez, M.A. (2005) Cobalt-mediated activation of peroxymonosulfate and sulfate radical attack on phenolic compounds. implications of chloride ions. Environmental Science & Technology, 40, 1000–1007. 180. Pope, M.T. (1983) Heteropoly and Isopoly Oxometalates, Springer-Verlag, Berlin. 181. Pope, M.T. and M€uller, A. (1991) Polyoxometalate chemistry: an old field with new dimensions in several disciplines. Angewandte Chemie-International Edition, 30, 34–48. 182. Day, V.W. and Klemperer, W.G. (1985) Metal oxide chemistry in solution: the early transition metal polyoxoanions. Science, 228, 533–541. 183. Papaconstantinou, E. (1989) Photochemistry of polyoxometallates of molybdenum and tungsten and/or vanadium. Chemical Society Reviews, 18, 1–31. 184. Mylonas, A. and Papaconstantinou, E. (1994) Photocatalytic degradation of chlorophenols to CO2 and HCl with polyoxotungstates in aqueous solution. Journal of Molecular Catalysis, 92, 261–267. 185. Hiskia, A., Mylonas, A., and Papaconstantinou, E. (2001) Comparison of the photoredox properties of polyoxometallates and semiconducting particles. Chemical Society Reviews, 30, 62–69. 186. Hiskia, A. and Papaconstantinou, E. (1992) Photocatalytic oxidation of organic compounds by polyoxometalates of molybdenum and tungsten. Catalyst regeneration by dioxygen. Inorganic Chemistry, 31, 163–167.

Part V Global Spaciality and Environmental Status of Transformation Products in the Environment

24 Occurrence of Transformation Products of Emerging Contaminants in Water Resources Carlos GonSc alves1, Maria A.D. Sousa1,2 and Maria de F
atima Alpendurada1,3 1

2

IAREN – Water Institute of the Northern Region, Portugal Department of Bromatology and Hydrology, Faculty of Pharmacy, University of Porto, Portugal 3 Faculty of Pharmacy, University of Porto, Portugal

24.1 Brief Introduction on the Sources of Transformation Products of Emerging Contaminants ECs are defined as contaminants not yet included in monitoring studies enforced by national or international regulations and which are continuously discharged into the environment, due to anthropogenic activities. Although studies on their environmental fate and (eco)toxicity are still very scarce and preliminary, they are suspected to pose threats to aquatic and terrestrial ecosystems [1,2]. Several studies demonstrate that exposure of non-target organisms to real environmental concentrations is unlikely to elicit acute toxicity, but often at these concentrations, chronic, sub-lethal effects were found to exist. Wastewaters are loaded with a huge diversity of organic anthropogenic substances, both natural and synthetic, typically in considerable amounts. Discharge into inland surface waters and sea, and reuse practices, including irrigation of landscape and agricultural areas or groundwater recharge, cause the distribution of these substances into the environment. Either directly or indirectly, fresh surface water bodies are the ultimate receiving media and the most impacted environmental compartment, while also being a source of potable water for human consumption [2,3].

Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

712 Transformation Products of Emerging Contaminants in the Environment

Pharmaceuticals, an important group of ECs, are released into the environment as the unchanged parent substance or as metabolites. When subjected to environmental factors, these substances suffer biotic and abiotic transformations. Along the path of distribution and transformation in the environment, a wide range of related substances is formed. Biochemical transformations of organic toxicants are governed by specific pathways mediated by enzymes, thus they are reasonably predictable and limited, while abiotic processes are extremely diverse and non-specific (e.g., catalyzed by radicals), resulting in myriads of TPs with different predominance [3]. As total mineralization is achieved in a negligible proportion, at the time scale relevant for exposure effects and when compared to the input rate of these ECs to the environmental compartments, all parent substances, metabolites and TPs should be studied from the occurrence, fate and ecotoxicological standpoints. Fatta-Kassinos et al. [3] introduced the topic of TPs resulting from advanced oxidation processes (AOPs) used for water and wastewater treatment, aimed at the destruction of recalcitrant organic substances but which usually fail to achieve total mineralization. The disappearance of the original substance does not guarantee the complete efficacy of the process and formed intermediates may preserve the mode of action of the parent compound or even be biologically more active [4]. In summary, the generation of TPs involves three main processes: (i) most pharmaceuticals are extensively metabolized (and often bioactivated) in human/animal bodies, before entering the wastewaters; (ii) abiotic and biotic transformations do occur in the environment at different rates, depending on the physicochemical properties of the substances; and (iii) advanced water treatment processes are source of TPs by incomplete detoxification of contaminants or conversion of benign substances into toxic TPs (e.g., during disinfection treatment) [2,5,6]. ECs, and pharmaceuticals in particular, are composed of an enormous variety of chemical structures, which are exposed to infinite environmental conditions. Studies on identification of TPs are vital for proper assessment of their environmental impact. However, a question remains as to how these studies can substantially contribute to the understanding of their overall fate under field conditions, where “cocktail” effects in aqueous environments are much more complex, with numerous organic and inorganic chemicals, including nutrients and suspended solids, interacting simultaneously [3]. Since the identification of all TPs is excessively challenging, attempts can be made to assess the final biological activity, though the assays must be relevant for human or environmental toxicology [5]. Following this strategy, another field is entered, since a large battery of assays has to be used to arrive at valid conclusions. Studies that have been conducted on the toxicity of pharmaceutical TPs were carried out on the overall solution, that is, composed of parent substance and by-products. Valid assays should tackle both real targets and real concentration levels, which is not always the case. Abiotic transformations comprise chemical hydrolysis, oxidation/reduction, isomerization, photolysis, and so on, influenced by chemical and physical factors, such as temperature, pH, salinity and chemical sensitizers that are a part of water and soil composition, and sunlight radiation with different intensity and wavelength, depending on the latitude, altitude and season of the year [3,4,7]. To improve our knowledge on this issue, novel strategies involving sophisticated analytical instrumentation, predictive tools, bioassays, and laboratory simulation tests are being used for critical toxicant identification/characterization, based on the concepts of

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 713

mode of action bio-diagnostics, effect based monitoring and toxicity identification evaluation (TIE). Escher and Fenner [5] emphasized the relevance of taking into account the TPs in environmental risk assessment studies, due to their diversity, amount and intrinsic hazard. Kern et al. [8] shared this opinion, indicating TPs of pesticides, biocides and pharmaceuticals as the major classes presenting a burden to the environment. The majority of TPs are as yet unknown. Humans and aquatic ecosystems are exposed to a plethora of chemicals composed of parent pollutants and their respective TPs, which, although in typically low concentrations, may display an additive or even synergistic effect on target organisms, adding a degree of complexity to environmental risk assessment [2,5]. Risk assessment can be regarded as an exposure-driven approach, in which pollutants are identified and quantified by chemical means, followed by effect-assessment; or an effect-driven approach, in which the toxicity of complex mixtures is first characterized, triggering the further identification of the toxicants (if toxicity is not eliminated along with the parent pollutant transformation) [5]. An innovative approach exploiting the potential of exact mass spectrometry was proposed by Garcia-Reyes et al. [9] and Gomez-Ramos et al. [4], for the identification of TPs of pesticides and pharmaceuticals, respectively. It is based on the assumption that organic compounds are often transformed into TPs in the environment in the same fashion that they are fragmented inside the analytical instrument, summarized as “fragmentation-degradation relationship”. This enables the compilation of exact mass databases, not only for target compounds, but especially also for screening unknown TPs in environmental samples, in an automated, simpler and much faster way. Identifying pollutants’ TPs in environmental matrices is a very complex task faced by researchers, which includes dedicated studies performed under laboratory conditions. Nevertheless, this should not hamper the accessing of total toxicity and biological potency of mixtures, as this will give useful insights on whether a treatment or natural process contributes to mitigation or enhancement of their environmental risks [3,4]. Different TPs with diverse environmental behavior and ecotoxicological profile can be formed, depending on the predominant chemical or biochemical process taking place (e.g., biodegradation, photolysis, hydrolysis), as well as the surrounding media where it occurs (surface water, groundwater, soil, sediment, wastewater effluents, etc.) [1,7]. The review published by Farr
e et al. [2] provides a good background for this chapter and gives an account of the levels of several ECs and respective TPs in wastewaters (influents and effluents) and surface waters until 2008. Nevertheless, it is discernible from the present literature review that much progress has been made on this subject in the last four years.

24.2 Transformation Products in Natural Waters: From Contamination Sources to Drinking Water Production Natural waters are the most relevant environmental compartment as regards the impact of pollution by ECs and their TPs, and the first to be endangered. Indeed, natural waters are the habitat of numerous species, give support to important human activities and are the source of environmental services (such as the supply of fish and raw water for the production of drinking water), which ought to be protected and preserved. Horvat et al. [1] have studied the occurrence and fate of anthelmintics and their TPs in the environment, based on their high applicability in animal husbandry and aquaculture.

714 Transformation Products of Emerging Contaminants in the Environment

Benzimidazoles, the most popular group of anthelmintics, may undergo sulfide oxidation in the environment, as well as in human metabolism. As a result, albendazole is converted into albendazole sulfoxide, also known as ricobendazole, the major active metabolite responsible for albendazole efficacy. Some of the sulfoxide is further transformed to albendazole sulfone, which does not appear to have any anthelmintic activity. Similarly, flubendazole is transformed into active fenbendazole sulfoxide, known as oxfendazole. All the benzimidazole drugs are sensitive to light. They can undergo ester-group demethylation, followed by decarboxylation of the carbamic group, giving rise to the amine-derivative, which has been reported to be the main photoTP and also the major metabolite of these drugs [10]. Ricobendazole amine has been identified. Diphenylsulfide bithionol also has a sulfide group that can be oxidized to bithionol sulfoxide and bithionol sulfone. These TPs may be produced as a result of direct photolysis or by intervention of Pseudomonas sp. and Nocardia sp. bacteria isolated from sewage sludge [1]. In some cases, the anthelmintic metabolites were found to have a greater effect than the parent compound, exemplified by some pro-drugs discussed above. Exposure to low concentrations of anti-parasitic drugs in the environment may also promote the development of resistant strains of parasites. Conversely, pyrantel is subject to photoisomerization upon exposure to light, with a loss of potency [1]. The information available on the levels of anthelmintics in the environment is very limited, and even scarcer for metabolites and TPs. Although levels of flubendazole and thiabendazole were measured in wastewaters (in the range 19.9–89.7 mg L
1 for influent and 55.0– 671 ng L
1 for effluent) and surface waters (3.9 to 27.3 ng L
1), no metabolites are reported. On the other hand, 22,23-dihydroavermectin B1a was found in sediments under a fish farm where ivermectin had been administered, at a mean concentration of 5.0 ng g
1 in the top 3 cm [11]. GonSc alves et al. [12] have studied the behavior of oseltamivir carboxylate (OC), the active metabolite of oseltamivir (OE, Tamiflu), in the context of the predicted pandemia of influenza virus H1N1. The levels measured in the Ebro river basin during fall/winter 2009 reached maximum concentrations in S
astago, downstream from the city of Zaragoza (OC 50 and OE 100 ng L
1), Miranda de Ebro (OC 46 and OE 83 ng L
1) and El Bocal (OC 42 and OE 83 ng L
1). Photolysis TPs of OE, identified as TP330 (photoinduced hydration) and TP312, (photoisomerization) were found in the above sites at vestigial levels, which could not be quantitated due to the absence of pure standards [12]. Also some tributaries presented levels of antivirals: river Huerva contained a concentration of OC of 46 and OE of 73 ng L
1, while the concentrations in river Segre amounted to OC 22 and OE 35 ng L
1 [12]. Ghosh et al. [13] found 6.6 to 190 ng L
1 of oseltamivir carboxylate during the peak of the flu season (January/February 2009) in the Katsura river catchment, receiving about 80% of wastewaters from Kyoto city, Japan. Prasse et al. [14] detected oseltamivir carboxylate in almost all surface water samples collected in the Hessian Ried region, south of Frankfurt am Main, Germany, in September 2009, with values ranging from 0.6–24 ng L
1. In the Ruhr watershed, levels did not exceed 2.4 ng L
1. The authors used the high rate of oseltamivir/ oseltamivir carboxylate levels (13.8) observed in Rhine river to infer about the contribution of another source besides the wastewater treatment plant (WWTP) discharges, which was attributed to wastes from a manufacturer [14]. Iodinated X-ray contrast media (ICM) and their TPs are ECs that have lately been given increasing attention, due to both the levels encountered in the environment and the great

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 715

variety of chemical entities formed. Environmental data support this statement. Kormos et al. [15] dedicated their research work to the simulation and elucidation of the biodegradation products of ICM, 19 of which were identified in environmental samples. In the Rhine river, close to Mainz, 15 TPs were detected above the respective limit of quantification, exhibiting a maximum concentration of 110 ng L
1 for iomeprol TP629. The authors explained the high levels found in the water as a result of the significant stability of these TPs, as treated wastewater inflows account for much less than 10% of the water stream [15]. The TPs detected in higher concentrations were generally those formed at the end of the proposed transformation pathways [15]. Environmental levels are given in Table 24.1. For further details, particularly the structure of the TPs, please check the original publication [15]. Jongh et al. [6] studied the fate and levels of pharmaceuticals and TPs along the watercycle, from source to finished drinking water. Concentrations in pre-treated surface waters, by such simple treatments as fast sand filtration, infiltration in dunes or sole storage in large reservoirs, generally reduced the concentrations by one order of magnitude. This was observed for tramadol, venlafaxine, and carbamazepine and its TPs. Other compounds kept similar concentrations in both waters. Notably, the river bank filtrate concentrations of phenazone and 1-acetyl-1-methyl-2-phenylhydrazide (AMPH) significantly exceeded the concentrations in surface waters by almost one order of magnitude. This finding was explained by the authors as possibly historical contamination, due to phenazone and dimethylaminophenazone high usage rates in the river Meuse and Rhine catchments some decades ago, and due to the long residence time of river bank pore water [6]. The concentrations of TPs rarely exceeded 100 ng L
1, with average values of several tens of ng L
1. The concentrations of O-desmethyltramadol in surface water samples ranged between 27 to 73% of its parent compound tramadol, which corresponds to the ratio of the excreted human metabolite via urine. This implies that both compounds undergo similar removal rates in wastewater treatment and behave similarly in the environment. In a different way, the concentrations of carbamazepine-10,11-epoxide were also below those of the parent compound carbamazepine (about 13 to 37%), despite the main human metabolite being carbamazepine-10,11-diol, to which the epoxide is transformed. Nevertheless, a study on wastewater showed that the ratios of carbamazepine-10,11-epoxide to carbamazepine formed are preserved in surface waters (12–13%) [6]. Conversely, the concentrations of O-desmethylvenlafaxine were higher than its parent venlafaxine in surface water samples (between 128 and 208%). This ratio is conserved from WWTP effluents (154 to 211%, in the Netherlands), although this composition does not correspond to the excretion ratios by humans, in which the metabolites far exceed the non-metabolized dose [6]. An exemplary case of toxicity enhancement is that of O-desmethyltramadol, which possesses a two- to four-fold higher pharmacologic activity than its parent tramadol. Therefore, risks for the environment can result from this TP, despite its lower levels. Based on these grounds, Jongh et al. [6] established a risk assessment framework, taking into account the mixture toxicity of parents and TPs, which shared a common toxicophore or pharmacological mechanism of action, and assuming an additive effect for the compounds within each group. Even so, no appreciable concern to human health was deduced from the levels of pharmaceuticals and TPs present in drinking water, which is in agreement with the review of risk assessment studies performed by the World Health Organization [16]. In a study aimed at the assessment of the behavior of pharmaceuticals and illicit drugs during drinking water treatment, Boleda et al. [17] quantified several TPs of these substances at the water intake. Norcodeine, the TP of codeine, attained concentrations of 3.1 to 6.0 ng L
1, compared with 29.9 to 43.9 ng L
1 for the parent compound. Among the synthetic

730 Transformation Products of Emerging Contaminants in the Environment

opiates, both methadone and its metabolite EDDP could be identified at low ng L
1 levels. The metabolites from cocaine, namely benzoylecgonine and norbenzoylecgonine, were detected at similar concentrations as found in European rivers (below 20 ng L
1), although a concentration of benzoylecgonine as high as 520 ng L
1 has been measured in Belgium [17]. Among several other pharmaceuticals, omeprazole and its TP hydroxy-omeprazole were detected, although these substances are sparsely found in the surface, due to their light sensitivity. Sildenafil was detected at a residual concentration (2.5 ng L
1), but its plasma metabolite N-desmethylsildenafil was not identified in any sample, while hydroxysildenafil, the main urinary metabolite, has not been searched for [17]. Among an extensive survey of pesticides, pharmaceuticals and respective TPs in wasteand surface waters, Gomez et al. [18] detected concentrations above 100 ng L
1 of four TPs of the analgesic-antipyretic dipyrone in surface waters (see Table 24.1). The authors observed that dipyrone metabolites are major contributors to the load of urban pollutants and were simultaneously present at higher concentrations in both effluents and river waters. The same authors also identified p-aminophenol, a TP more toxic than the parent compound, acetaminophen, and a TP of azithromycin with m/z 591.4218 [18]. According to these results, TPs should not be neglected in future monitoring studies. Pesticides’ TPs in the environment are a particularly concerning group of substances, because of their potential acute toxicity and legal monitoring requirement. Hernandez et al. [19] carried out the entire job of elucidation, identification and quantitation of several of these substances in surface and groundwaters of Valencia, Spain, an agricultural area much affected by the use of pesticides. TPs were detected in higher number than parent substances and at concentration levels reaching some mg L
1, which exceed those of the respective parent compounds (see Table 24.1). These concentrations exceed the 0.1 mg L
1 limit generally established for natural waters, which would pass unnoticed if TPs were not monitored. Hadlik et al. [20] performed an extensive study of surface waters of the upper Chesapeake Bay, aiming to clarify the relevance of the TPs of triazines and chloroacetamides and the distribution in the depth profile of the bay. The authors detected 19 out of the 20 searched neutral chloroacetamide degradates of interest, along with three ionic oxanilic acid derivatives. The results clearly demonstrated that the concentrations of most neutral chloroacetamide degradates exceeded those of the parent compounds (up to 10-fold), while the total concentration of TPs was 20–30 times that of the parents. The reader should consult this interesting study for further information on the structures of the TPs and their concentration levels. In Table 24.1, a concentration range is given considering the highest level measured in the surface layers [20]. This study is a good example of where the concentration levels of the TPs and their diversity, combined with their often higher toxicity, resulted in a significant threat to the environment. Bearing in mind the reported difficult removal of alachlor (and eventually its TPs) during drinking water production processes, these surface waters would also pose a risk to human health [20]. Facing the limited availability of standards of TPs needed for targeted search for these substances in natural waters, as well as the uncertainty in transferring the findings obtained under laboratory simulation experiments to a real environmental scenario, Kern et al. [8] developed a computer-aided strategy for predicting and searching TPs in real samples, based on known microbiological transformations and data from the literature. Representative surface water samples from various locations in Switzerland were broadly enriched by solidphase extraction (SPE) followed by Orbitrap LC-MS/MS analysis, which provided high resolution and accurate mass spectrometry chromatograms. Potential TPs from 24 pesticides, 7

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 731

biocides and 21 pharmaceuticals were searched. Among the 19 TPs identified (see Table 24.1), a majority originated from pesticides. Pharmaceutical TPs found in samples were mostly human metabolites, while 4 of them might also have been formed by microbial metabolism, for example, during activated sludge treatment [8]. The environmental contamination with polybrominated diphenyl ethers (PBDEs) poses a singular scenario. Although less brominated PBDEs have been phased-out, it has been shown that heavier ones, like deca-BDE, degrade by photolytic cleavage into lighter ones (debromination), including penta-BDE. This last substance is not only more toxic, but also more stable and more easily transported in the atmosphere, crossing long distances to remote regions. Schenker et al. [21] have estimated that about 13% of penta-BDE and about 2% of tetraBDE homologues found in the environment arise from the degradation of deca-BDE, which must, therefore, be combined with emission rates. Although this study focused on the air compartment, deposition in rainfall leads to the contamination of water bodies. Qu et al. [22] carried out an extensive protocol aimed at the identification of neurotoxic brominated flame retardants (BFRs) and their TPs in environmental samples, collected near a BFRmanufacturing plant in China, using a bioassay-directed fractionation and LC-MS/MS strategy. Tetrabromobisphenol A diallyl ether was identified as a key developmental neurotoxicant in the most potent fraction of sediments, where it was also measured in huge concentrations (around 10 mg kg
1). This compound partitions to the water phase in the ng L
1 concentration range and it was detected 3.1 km farther downstream from the BFR manufacturing plant [22]. Alkylphenol ethoxylates (APEOs) constitute one of the main nonionic surfactant groups with particular environmental relevance, since some of their possible metabolites (nonylphenol (NP), A9PEO1 and A9PEO2) have been shown to be weak endocrine disrupting compounds. Nonylphenol polyethoxylates are large volume production chemicals used as detergents, emulsifiers and dispersing agents, during the last four decades, both for domestic and industrial purposes [2,23–25]. They are good examples of ECs whose continuous introduction into the environment keeps their concentration levels and those of their TPs high. As these products suffer extensive transformation, the whole chain of TPs can be found in surface waters, with concentrations ranging from hundreds of ng L
1 to several mg L
1 and with different toxic burden [23]. Since all of these compounds exist together as a mixture in the environment, a risk assessment of the overall mixture is required, as will be discussed in a subsequent section. Despite a reasonable amount of data being currently available concerning the occurrence of APEOs in fresh waters, knowledge of their subsequent fate in saline waters is still limited. Maximum estuarine and sea concentrations reported in the literature are around 1 mg L
1, with some exceptions, such as 10 mg L
1 in Spain and 25 mg L
1 in Israel. These values are found to be roughly one order of magnitude lower than those in fresh waters [26,27]. To help suppress this lacuna, Jonkers et al. [28] carried out a field study with the goal to determine the sources of these nonionic surfactants in the Dutch coastal zone, as well as their fate. They analyzed surface waters, sediments and suspended particulate matter for A9PEOs, NP and the carboxylated metabolites (A9PECs). According to different sampling campaigns, A9PEO1,2 and A9PEO>2 maximum concentrations varied between 0.14–0.73 and 0.27– 35 mg L
1, respectively, while NP and A9PEC maximum levels ranged from 0.031–1.7 and 0.11–0.63 mg L
1, respectively. On the other hand, on the fresh water side of the locks at the mouth of the North Sea canal, slightly higher concentrations of metabolite compounds were obtained: 0.042 mg L
1 for NP and 0.31 mg L
1 in the case of A9PEC. Tubau et al. [29] studied the levels of alkylphenol polyethoxylate TPs and linear alkylbenzene sulfonate (LAS) surfactants in urban groundwater, in the city of Barcelona.

732 Transformation Products of Emerging Contaminants in the Environment

Groundwaters were characterized by analyzing well and piezometer waters in different aquifers. Although LAS are currently more used than APEOs and were quantified at higher levels in recharge waters, the levels of APEOs TPs in groundwater were, on some occasions, higher than the former. The results obtained in zone Z3 pertaining aquifers, near the river Besos, which flows over deltaic aquifers with high transmissivity and water levels close to the soil surface, are given in Table 24.1. According to the wide survey carried out at the European level in search of the occurrence of 35 selected compounds in river water samples from 122 European streams, several APEOs TPs, such as alkylphenol carboxylates (APECs), were often found, at concentrations sometimes surpassing the established environmental quality standards (EQS) [30]. Indeed, NPE1C was found in 97% of the samples at a median concentration around 230 ng L
1, while NP showed a median of 134 ng L
1, but a frequency below 30% [30]. In a similar pan-European study conducted on 164 groundwaters, NPE1C was also detected in high concentration (max. 11.3 mg L
1 and average of 263 ng L
1), while NP achieved a maximum of 3.8 mg L
1 and average of 83 ng L
1 [31]. The frequencies of detection were 41.5 and 11%, respectively, while tert-octylphenol was present in 23.2% of the samples, with an average concentration of just 1 ng L
1. Loos et al. [31] concluded that NPECs are persistent chemicals, widespread in European groundwaters. Chloridazon TPs, chloridazon-desphenyl and chloridazonmethyldesphenyl, were present in groundwaters in very relevant concentrations (max. 13 and 1.2 mg L
1, respectively) and an average concentration two orders of magnitude lower. It is worthwhile noting that chloridazon-desphenyl was the chemical which exceeded the threshold value of 0.1 mg L
1 most frequently (26 times), followed by NPE1C (20 times). Poly- and perfluoroalkyl substances (PFAs) comprise a diverse range of high production volume chemicals that have been used in industrial and consumer products for the last five decades (in metal plating, firefighting foams and oil/water repellents). Chemicals like perfluorooctanosulfonates (PFOS) and perfluoroalkylcarboxylates (PFCA) are themselves better known than their chemical precursors of industrial origin (ammonium perfluorooctanoate (A-PFO), and fluorotelomer alcohols (FTOHs) and perfluoroalkane sulfonamidoethanols (PFASEs), respectively) [32]. These compounds are so persistent and ubiquitous that they could be found in sea waters of remote regions, most surface waters and also in wildlife [30,32]. Maclachlan et al. [33] conducted a survey of perfluorinated carboxylates (PFCA) in 14 major rivers, including the Rhine, Danube, Elbe, Oder, Seine, Loire and Po. The highest concentrations measured were 200 ng L
1 for perfluorooctanoate (PFOA) in the Po river, followed by 32 ng L
1 of perfluorohexanoic acid (PFHxA) in the Thames river. In most other cases, levels of PFAs substances remained below 10 ng L
1. The Po river accounted for two-thirds of the total PFOA discharge in the studied rivers, a fact which the authors attributed to a fluoropolymer manufacturing facility in the Po watershed [33]. In recent years, attention has been expanded from environmental monitoring of acidic PFAs to include also neutral compounds, such as fluorotelomers (olefins, alcohols, acrylates), fluoroalkylsulfonamides and fluoroalkylsulfonamidoethanols [34].

24.3 Wastewaters as a Major Source of Transformation Products Effluents from municipal wastewater treatment plants were identified as the major source of ECs and their TPs (hormones, pharmaceuticals and personal care products) in environmental matrices [35]. The estrogen-mimicking chemicals deserve special attention due to their

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 733

reproductive effects on aquatic organisms, especially natural and synthetic steroid estrogens. Fatta-Kassinos et al. [3] stress that knowledge of the biodegradation products of pharmaceuticals, that is, TPs formed during biological wastewater treatment, is currently very limited, therefore studies focusing on this topic should be undertaken so that the possible hazards related to the biotransformation products released into the environment can also be assessed. Applying the systematic approach proposed above and named “fragmentation-degradation relationship”, Gomez-Ramos et al. [4] identified 8 TPs in wastewaters out of 147 precursor compounds that were established in the database. The most frequent TPs in 8 samples collected from 4 WWTPs in Spain were erythromycin anhydride, cleaved azithromycin and 4-aminophenol. The TP of erythromycin, identified as erythromycin anhydride (m/z 716), was generated by internal dehydration of the original molecule, and was quantified in effluent samples in the range 27–159 ng L
1, using a commercial authentic standard. A hydrolysis product of erythromycin, erythralosamine (m/z 540), was also identified. The authors call particular attention to the presence of 4-aminophenol in wastewaters as a TP of acetaminophen, although other origins could be hypothetically assigned. Levels of this TP in the mg L
1 range were quantified, raising ecotoxicological concerns as 4-aminophenol has been described as very toxic to aquatic organisms, causing long term adverse effects [4]. TPs of carbamazepine, an ubiquitous, persistent and abundant drug in the aquatic environment, were also detected: the monohydroxylated product with m/z 255 was assigned as 10,11-dihydro-10,11-hydroxycarbamazepine, while the dihydroxylated product was also evidenced with m/z 271, corresponding to 10,11-dihydro-10-dihydroxycarbamazepine, corroborating a previous report [36]. Carbamazepine 10,11- epoxide, which is said to be frequently detected in wastewater samples, was also identified. Ghosh et al. [13] have measured oseltamivir carboxylate at a maximum concentration of 293 ng L
1 in effluents from conventional activated sludge-based WWTPs, but the concentration decreased to 37.9 ng L
1 when an advanced WWTP with ozonation as a tertiary treatment was used. The levels of oseltamivir carboxylate measured by Prasse et al. [14] in two German WWTPs were 42.7 and 17.3 ng L
1 in the influent and effluent of WWTP1 and 29.4 and 12.2 ng L
1 in WWTP2 respectively, while the levels of oseltamivir were constantly less than those of the metabolite, in agreement with the human excretion rate of about 80% in the carboxylate form. A TP of amoxicillin, a commonly used antibiotic belonging to the group of penicillins, was recognized as being (5R) amoxicillin diketopiperazine-20 .50 (m/z 366), recently reported in both wastewater and surface waters [37,38]. Jia et al. [39] developed a method for the determination of trace levels of 6 tetracyclines (TC) and 10 of their TPs in environmental waters. Tetracyclines are an important group of antibiotics, used in the treatment of several human and animal diseases. However, they elicit ecotoxicity on fresh water species, such as the cyanobacteria Micocystis aeruginosa, and phytotoxicity on aquatic higher plants such as Lemna gibba [40,41]. Consequently, their environmental presence, along with their TPs, attracts great attention. Tetracycline abiotical degradation is known to occur, depending on pH, redox and light conditions, through epimerization, dehydration and proton transfer reaction pathways [42]. 4-Epi-TCs can be formed in the aquatic medium under mildly acidic conditions (pH 2–6) and reversed back to their active form in specific alkaline conditions, and in the presence of a complexing metal. Such are the cases of 4-epitetracycline (ETC) and 4-epioxytetracycline (EOTC), originated from TC and oxytetracycline (OTC), respectively. On the other hand, anhydro-TCs, such as anhydrotetracycline (ATC) and anhydrochlortetracycline (ACTC), are formed under strongly acidic conditions (pH < 2). These compounds can also epimerize to form epi-analogues [39].

734 Transformation Products of Emerging Contaminants in the Environment

Jia et al. [39] then applied the developed method to the analysis of wastewaters and surface waters in China, using commercial standards for the unambiguous identification of every compound. Apart from TC and OTC, the main analyte quantified in an influent sample at 72.5 ng L
1 concentration, several degradation products were also detected at low nanogram per liter levels. Such were the examples of ETC, EOTC, ATC, isochlortetracycline (ICTC) and 4-epianhydrochlortetracycline (EACTC), mostly eliminated through the WWTP treatment; besides TC and OTC, only ICTC and EACTC were quantified in effluent samples, ranging from 1.9 to 7.6 ng L
1. Metformin is an antidiabetic drug, highly consumed worldwide, which partially justifies its release in significant amounts (low mg L
1 range) to environmental recipient waters. On the other hand, this compound is not metabolized in humans and the reabsorbed fraction is excreted unchanged in the urine (about 70%) and the rest in feces [43]. Reported measured concentrations of metformin in Belgian WWTP influents went up to 94 mg L
1 [44], while Scheurer et al. [45] published values for German WWTP influents as high as 129 mg L
1. It is also known that metformin is biologically degraded to guanylurea in WWTPs. Due to high metformin concentrations in influents and its high yet incomplete removal during the treatment process, both compounds reach surface waters in considerable amounts of up to several tens of mg L
1 [46]. Although a study from Trautwein et al. [47] indicated guanylurea as the only recalcitrant, aerobic, bacterial TP of metformin, the detected concentrations of guanylurea in the effluent samples did not completely account for the corresponding removed fraction of metformin. Information is also scarce concerning the ecotoxicological relevance of both these compounds. Therefore, Scheurer et al. [46] recently published a systematic study aiming to clarify the direct transformation of metformin into guanylurea in WWTPs, evaluating the effectiveness of different treatment techniques applied in waterworks, and reporting the respective effluent concentrations. In the five WWTPs in the study, guanylurea ranged from undetectable levels to 3 mg L
1 in influent samples, reaching a concentration of 99 mg L
1 in an effluent sample, corresponding to an influent with 105 mg L
1 of metformin. This observed widespread occurrence of guanylurea reinforces the environmental relevance of metabolites. To better assess the impact of ECs and their TPs in the environment and human health, it is necessary to better understand the chemical and physical transformations occurring at the treatment plants, in order to determine their capacity to remove the contaminants or turn them into more persistent and toxic compounds. Gagnon et al. [48] have reported their results on the influence of different wastewater treatment processes on the fate of some pharmaceutical compounds. During this study, 2-hydroxy-ibuprofen was identified as a metabolite of ibuprofen, originating during the treatment process at the plant. This TP, 2-hydroxyibuprofen, was quantified in effluent samples in concentrations ranging from about 300 to 3500 ng L
1, depending on the WWTP where the sample was collected and the treatment technique employed therein. The relatively high concentrations of 2-hydroxy-ibuprofen helped to explain the low concentration of ibuprofen after treatment [48]. Iodinated contrast media TPs, discussed in the previous section, were also detected in a municipal WWTP effluent with concentrations as high as 660 ng L
1 for iomeprol TP791. While the concentration levels measured were the highest among all environmental matrices analyzed, less TPs (only 10) were detected [15]. In another work, Bueno et al. [49] described the development of an enhanced liquid chromatography-mass spectrometry (LC-MS) strategy for the analysis of a selected group of 56

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 735

organic pollutants in wastewaters: 38 pharmaceuticals and 10 of their most frequent metabolites, 6 pesticides and 2 disinfectants. They focused on the development of a suitable analytical methodology, which could allow an exhaustive characterization of wastewater effluent samples. Since target analysis based on LC-MS provides a great performance for quantitative analysis, but fails in the determination of compounds not initially included in the multiresidue methods (non-target compounds), the latest trends to increase the scope of methods include the combination of two complementary LC-MS techniques or the use of hybrid systems with different analyzer designs. Therefore, Bueno and coworkers [49] adopted a LC-MS methodology based on the use of a hybrid triple-quadrupole linear ion trap mass spectrometer (QTRAP), for accurate quantification, in combination with time-offlight mass spectrometry (TOF-MS), providing unequivocal confirmation. The developed methodology was then successfully applied to a monitoring study of six WWTPs, in Spain. Herein, a group of metabolites of pharmaceutical compounds were detected and quantified: carbamazepine-10,11-epoxide (13–110 ng L
1), 1,7-dimethylxanthine (or paraxanthine) (131–80 875 ng L
1), clofibric acid (67–81 ng L
1), fenofibric acid (14–349 ng L
1), and the metabolites of the antipyretic drug dipyrone, 4-methylaminoantipyrine (4-MAA, 9–9253 ng L
1), N-acetyl-4-aminoantipiryne (4-AAA, 2109–25 030 ng L
1), N-formyl-4-aminoantipiryne (4-FAA, 40–10 114 ng L
1), 4-dimethylaminoantipiryne (4-DAA, 122 ng L
1), 4-amino-antipiryne (4-AA, 131–9286 ng L
1), and antipyrine (17–2760 ng L
1) [49]. The authors also draw particular attention to the fact that paraxanthine and 4-AAA were present at the higher levels in the effluent samples collected in the WWTP of Almeria, possibly due to the proximity of a hospital. Moreover, the group of metabolites of dipyrenone was also detected at quite high concentrations, which further reinforces the relevance of this widely consumed antipyretic [49]. Concerning the group of sulfonamides, widely used synthetic antibiotics, Garc
ıa-Gal
an et al. [50] studied the behavior of sulfapyridine (SPY), typically used in human therapy, and the veterinary sulfonamide sulfamethazine (SMZ), as well as their acetylated metabolites, AcSPY and AcSMZ, in wastewater matrices under artificial irradiation conditions. Compounds as SPY and SMZ are among the sulfonamides most frequently detected in effluent wastewaters and surface waters, respectively. Results showed that the photolysis of SPY produced a total of 10 different TPs, 4 major compounds (M) and 6 other minority products (m), upon irradiation for 30 h: the hydroxylated TP (C11H12N3O3S), m/z 266.0599 (M); the SO2 extrusion product, N-(pyridin-2-yl)benzene-1,4-diamine (C11H12N3), m/z 186.1031 (M); the hydroxylated moiety of this desulfonated TP, N-hydroxy-N-(pyridin-2-yl)benzene1,4-diamine (C11H12N3O), m/z 202.0980 (m); a compound resulting from the reduction of the hydroxylamine in the N4 position of the hydroxylated TP (C11H10N3O3S), m/z 264.0443 (M); pyridine-2-amine (C5H7N2), m/z 95.0609 (M); hydroquinone (m); aniline (m); sulfanilamide (m); sulfanilic acid (m); aminosulfonic acid (pyridin-2-ylsulfamic acid) (C5H7N2O3S), which resulted from the cleavage of the bond between the aniline ring and the sulfonic group, followed by an -OH addition to the SO2 group, m/z 175.0177 (m). On the other hand, the photodegradation of N4-acetylsulfapyridine yielded a new molecule with a predicted elemental composition C13H14N3O (m/z 228.1137), through the loss of the sulfonate group. Regarding SMZ, eight different TPs were identified: similarly to what happened with SPY, a desulfonated product of SMZ was detected, presenting two different signals corresponding to the same mass (m/z 215.1297) and yielding the same mass spectra – stereoisomers with the elemental composition C12H15N4; its hydroxylated moiety C12H15N4O, m/z 231.1246, and the desaminated TP C12H14N3, m/z 200.1188; 4,6-dimethylpyrimidin-2-amine (C6H10N3),

736 Transformation Products of Emerging Contaminants in the Environment

m/z 124.0875; the hydroxylated TP of SMZ (m/z 295.0865) and desaminated moiety (m/z 264.0807) of SMZ; a compound resulting from the reduction of the hydroxylamine in the N4 position of the hydroxylated TP (C12H13N4O3S), m/z 293.0708; (4,6-dimethylpyrimidin-2-yl) sulfamic acid (C6H10N3O3S), m/z of 204.0443. Finally, AcSMZ photodegradation yielded 3 TPs: one resulting from the loss of SO2 and the methyl group from the acetylated moiety (C13H14N4O), m/z 243; a second one as the result of a desulfonation reaction (m/z 257); a third TP attributed to a structural rearrangement of the desulfonated photoproduct (C15H19N4O), m/z 271. Benzodiazepines (BZP) are prescribed in high amounts worldwide and constitute potentially new emerging contaminants. The environmental persistence and fate of these pharmaceuticals, as well as their TP, is of high relevance though it is, as yet, scarcely understood. Hence, a study by Kosjek et al. [51] addressed this gap by monitoring environmental concentrations of several benzodiazepine residues and studying their removal during biological and photochemical water treatment, including the identification of stable TP formed during wastewater treatment. As a result, the elucidation of eight novel diazepam (DZP) and four novel oxazepam (OXA) TPs was performed. DZP biodegradation results in the formation of OXA, nordazepam (NRZ) and temazepam (TMZ), which are also human metabolites of DZP and marketed as individual pharmaceuticals. NRZ is formed by N-demethylation on N-1 of DZP, TMZ by hydroxylation of the C-3 atom and OXA comprises both transformation reactions. The newly discovered DZP TPs included 5 isomers with elemental composition C16H14N2O2Cl and nominal mass of 300 Da, resulting from one single hydroxylation of DZP, and 2 isomers with elemental composition C16H14N2O3Cl and nominal mass of 316 Da, resulting from double hydroxylation of DZP, all obtained by photocatalysis; the last TP identified, resulted from biotransformation of DZP and presented the elemental composition C16H16N2O2Cl and a nominal mass of 302 Da. Regarding OXA, 1 degradation product with elemental composition C15H12N2OCl and nominal mass 270 Da was originated by biotransformation reactions, while the other 3 hydroxylated isomers, presenting the elemental composition C15H12N2O3Cl and a nominal mass of 302 Da, were the result of photocatalysis at pH 2. In another study by Calisto et al. [52], the relevance of photodegradation processes to the environmental persistence of four benzodiazepines – OXA, DZP, lorazepam (LZP) and alprazolam (ALP), was investigated. These benzodiazepines were irradiated under simulated solar irradiation, together with three different fractions of humic substances. A total of 19TPs were identified by electrospray mass spectrometry, 7 for OXA, 4 for DZP, 6 in the case of LZP, and finally 2 for ALP (Table 24.1). Sousa et al. [53] also studied LZP phototransformation pathways under artificial UV and natural solar irradiation, through photolytic and TiO2-assisted photocatalytic processes. Herein, 14 compounds (excluding structural isomers) were identified as presumable LZP TPs (Table 24.1). The proposed photodegradation mechanism included the initial opening of the diazepinone seven-membered ring, followed by a rearrangement into a highly stabilized six-membered aromatic ring, and subsequent cleavage and/or hydroxylation reactions. LZP photocatalytic degradation was further assessed on a contaminated municipal effluent, where the TPs generated proved to be more persistent than LZP itself. Estrogens high potency at the nanogram per liter level makes it important to assess their fate in the urban environment and to evaluate the effectiveness of different processes of the urban sewer infrastructure in removing or lowering their estrogenic potential. Furthermore, although estrogens present the strongest estrogenic activity, other chemicals, such as

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 737

octylphenol, nonylphenol, nonylphenol ethoxylates, bisphenol A, and phthalates also express estrogenicity, even if to a lesser extent [24,54]. Limpiyakorn et al. [55] recently published a review summarizing data collected from a total of 130 full-scale WWTP distributed over 14 countries and focusing on the fate of estrogens and estrogen potentials in the sewerage system, starting from human excretion through to municipal sewage treatment facilities. Humans generally excrete 90–95% of estrogens via urine, in the conjugated forms rather than in the free forms. Conjugated estrogens usually found in female urine include estrone 3-(b-D-glucuronide) (E1–3G), estrone 3-sulfate (E1–3S), 17b-estradiol 3-glucoronide (E2–3G), 17b-estradiol 17-glucoronide (E2–17G), 17b-estradiol 2-sulfate (E2–3S), estriol 3-(b-D-glucuronide) (E3–3G), estriol 16a (b-D-glucuronide) (E3–16G) and estriol 3-sulfate (E3–3S). Glucuronide estrogens are found to be the most predominant form [56,57]. Later, in the septic tanks, the dominant transformation mechanism appears to be conjugation, with sulfate estrogens concentrations in the total conjugated estrogens, increasing from 22% in the influent to 55% in the effluent of the septic tank. In sewers, conjugated estrogens E1–3S, E1–3G, E2–3S, E2–17G, E2–3G, E3–3S, E3–16G, and E3–3G were found to be reduced by about 64, 84, 100, 0, 100, 84, 100 and 0%, respectively [56]. Finally, in municipal WWTPs samples, D’Ascenzo et al. [56] were able to quantify E1–3G, E1–3S and E3–3S, of which 0.7, 9.0, and 2.2 ng L
1, respectively, remained in the effluent. In another study conducted by Isobe et al. [58], E1–3S and E2–3S were found in the concentrations 0.3–2.2 and 1.0 ng L
1, respectively, in the effluent of WWTPs in Japan. On the other hand, in a batch study performed by Ternes et al. [59], the conjugated estrogen E2–17G was found to be totally cleaved under aerobic conditions by activated sludge in a German WWTP. Moving to the TPs of pesticides, Gomez-Ramos et al. [4] have identified unequivocally, with the help of a pure standard, the 2-isopropyl-6-methyl-4-pyrimidinol (IMP, m/z 153) as a TP of diazinon, an organophosphorus pesticide widely used in urban and agricultural applications. This TP was quantified in two samples from two different WWTPs at concentrations 288 and 630 ng L
1. It had already been reported in surface waters and referenced as one of the major TPs of diazinon, showing a higher genotoxic potential than diazinon itself [36,60]. As previously mentioned, nonylphenol polyethoxylates (NPEOs) give rise to a great variety of dicarboxylic TPs (CAPECs), particularly those containing five to eight carbon atoms and a carboxyl group in the alkyl chain (CA5-8PECs). Their demonstrated presence and possible persistence in several environmental compartments have been raising concern worldwide [61]. However, reports on CAPECs environmental occurrence and behavior, including their possible eco- and human toxicological effects, are still very scarce. In this context, Hoai et al. [61] developed and validated a method combining SPE, derivatization and gas-chromatography-chemical ionization mass spectrometry (GC-CI-MS) for the measurement of four model compounds of dicarboxylic metabolites (dm-CA5-8P1EC) and other dicarboxylic metabolites (CA5-8P1ECs) of nonylphenol polyethoxylates in wastewater, river and pure water. The dicarboxylic metabolites referred to as dm-CA5-8P1EC present an a,a-dimethyl configuration (expressed as “dm”), five to eight carbon atoms and a carboxyl group in the alkyl chain, besides an ethoxy acetic acid group. Though dm-CA5-8P1ECs metabolites were not detected in the collected water samples, Hoai et al. [61] could identify 21 isomers of CA5-8P1ECs by CI-MS (bearing in mind the similarity of the fragmentation ions in the respective spectra) in surface and wastewater effluents in Japan: 2 isomers of the CA5P1EC metabolite, 5 isomers of the CA6P1EC metabolite, 6 isomers of the CA7P1EC metabolite and 8 isomers of the CA8P1EC metabolite. The metabolites of CA6P1EC and CA8P1EC were identified as the dominant compounds, at

738 Transformation Products of Emerging Contaminants in the Environment

mg L
1 level, from 2 to 15 times higher than CA5P1EC and CA7P1EC metabolites, leading to the conclusion that nonyl chain degradation occurs mainly through the elimination of two carbon units. Hoai et al. [61] also highlighted that dicarboxylic TPs should play an important role when studying the behavior of NPEOs in WWTPs, due to the high concentrations of CA5-8P1ECs metabolites in the studied river and WWTP effluents. Moreover, the obtained results for surface and wastewaters in Japan were compared to previous ones reported in Italy and Taiwan, were CA5-8P1ECs concentrations were about one and two orders of magnitude higher, respectively.

24.4 Origin and Presence of Transformation Products in Drinking Water In their review, Mompelat et al. [7] highlighted the human health risk associated with the occurrence of pharmaceuticals and TPs in drinking water, even at low concentrations. Furthermore, according to these authors, drinking water is the least studied when it deals with the occurrence, fate and behavior of EC TPs, compared to environmental matrices. Regardless of the matrix, active metabolites of pro-drugs, for example, clofibric acid, fenofibric acid and salicylic acid, are the most researched [7]. A good collection of excretion metabolites gathered from marketing authorization literature is given by these authors, however, only a few were actually detected in environmental matrices. A review of the literature on the occurrence of pharmaceuticals and TPs in drinking water revealed the presence of 17 parent compounds and 5 TPs, derived from dimethylaminophenazone, diclofenac, clofibrate and ketoprofen, as given in Table 24.1. In general, non-steroidal anti-inflammatory drugs, iodinated contrast media and their TPs are the chemicals causing most concern. Lee and coworkers [35] have studied the fate of steroid estrogens upon chlorination. Ortho-substituted derivatives were identified as the main TPs, which displayed considerably weaker estrogenic potency. In the presence of increased levels of bromine in raw water and a low concentration of dissolved organic matter and ammonia (which act as quenchers), secondary products of 17a-ethinylestradiol (EE2) were identified, containing one or two bromine atoms in both positions (2- and 4-) of the phenolic ring. These TPs can be further transformed, either by chlorine or bromine, into a cleaved phenolic moiety [35]. During the chlorination process (at pH 7) and under typical concentrations of 17a-ethinylestradiol at ng L
1 levels and bromine between 10 and 100s mg L
1, the predominant products are the brominated ones (4-Br- and 2,4-diBr-17a-ethinylestradiol), until total depletion of bromine. The formation of chlorinated by-products of 17a-ethinylestradiol is faster at increased pH, from 7 to 9, with higher accumulation of 2,4-dichloroderivatives. In the presence of bromine, the transformation rates are greatly enhanced, due to the higher reactivity of bromine, but this trend diminishes with increase in pH. The authors advocate that the same behavior can be extrapolated for other steroid estrogens, such as 17b-estradiol (E2), estrone (E1) and estriol, and other phenol-containing endocrine disrupting compounds, such as bisphenol A and nonylphenol, in which the presence of bromine accelerates the transformation and gives rise to less potent by-products [35]. Pereira et al. [62] have identified two by-products for each estrone and 17b-estradiol, under ozonation conditions, during drinking water disinfection: 10-hydroxy-1,4-estradieno3,17-dione and 10e-17b-dihydroxy-1,4-estradieno-3-one, respectively, and another two analogues with a cleaved phenolic moiety. Besides these experimental results, Pereira et al. [63]

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 739

also produced a very comprehensive review of all estrogen by-products generated during water disinfection treatments, through conventional and advanced processes, totalling 48 molecules derived from E1, E2 and EE2. Chlorinated by-products are recalcitrant to further transformation. Mono-, di- and tri-substituted derivatives have been identified for estrone, ending up in 2,4,10-trichloro-1,4-estradiene-3,17-dione. Although there are divergent opinions, the generalized consensus is that by-products formed during the disinfection processes have reduced or negligible estrogenicity. Small modifications of the phenolic ring tend to decrease the estrogenic activity, namely halogenation at the 2 and 4 positions [63]. Under ozonation conditions in water treatment works, the primary oxidation end-products of metformin were identified as the hydroperoxide of metformin, methylbiguanide and 2-amino-4-imino-5-methyl-1,3,5-triazine [46]. Under chlorination, both substances, metformin and guanylurea, should undergo similar reactions. Dimethylamine was identified as the prime chlorination product, but others such as N-chlorourea, NCl3, NHCl2 and NH2Cl can be formed by hydrolysis of the amine group and further chlorination [46]. The presence of guanylurea in finished drinking water seems, however, to be unlikely, if an underground passage is part of the raw water treatment train [46]. Following the discussion presented in the previous sections about iodinated contrast media, as many as 15 TPs were identified by Kormos et al. [15] in drinking water, but only seven of them exceeded a concentration of 10 ng L
1. Iomeprol TP629 was consecutively highlighted in the different matrices, with a maximum concentration of 120 ng L
1 in drinking water. According to these results, even advanced treatment processes, such as ozonation and activated carbon filtration, are not capable of completely removing these TPs. Due to their chemical similarity to the respective parent compounds, probably a similar negligible impact to ecosystems and humans is to be expected from these substances. However, the authors highlighted the possibility of formation of low molecular weight and toxic disinfection by-products in contact with strong oxidants in drinking water systems [15]. Jongh et al. [6] measured relatively low concentrations (up to 35 ng L
1) of phenazone and AMPH in three drinking water samples produced from river bank filtrates. The occurrence of phenazone and AMPH in drinking water is most likely a result of their relatively high concentrations in the source water (up to 258 and 172 ng L
1, respectively) and the hydrophilic character of phenazone and AMPH. Although the studied surface waters were contaminated with a much larger variety of pharmaceuticals and TPs, present also at higher levels, the produced drinking water was not affected by these substances, conversely to that observed in the case of the river bank filtrate. The question remains, if the compounds that persist to bank filtration conditions (biodegradation and sorption to the soil) coincide with those that are afterwards more difficult to treat during drinking water production. Illicit drugs have been recognized in recent years as a new group of water contaminants and a sufficient pool of data already exists on their ubiquity around the world [64]. HuertaFontela et al. [64] have demonstrated in previous works that most illicit drugs and respective metabolites can be removed during the steps of drinking water production. Nevertheless, they can give rise to the formation of disinfection by-products. A common chlorinated by-product, (3-chlorobenzo)-1,3-dioxole, was identified for both 4-methylenedioxyamphetamine (MDA) and 3,4-methylenedioxyethamphetamine (MDEA), while for 3,4-methylenedioxymethamphetamine (MDMA) 3-chlorocatechol was found [64]. The chemical (3-chlorobenzo)-1,3-dioxole was found after the first chlorination step, in concentrations ranging from 1.2 to 3 ng L
1, and was eliminated after ozone and graphitic activated carbon filtration. On the other hand, 3-chlorocatechol, generated mainly after the post-chlorination

740 Transformation Products of Emerging Contaminants in the Environment

step, proved to be recalcitrant and it was found in final treated waters, at concentrations between 0.5 and 5.8 ng L
1 [64]. M€ uller et al. [65] studied the fate of the industrial chemicals 4-methyl-1H-benzotriazole and 5-methyl-1H-benzotriazole (anticorrosives, “silver protection” products in dishwasher detergents, additives in surface coatings, cooling lubricants, hydraulic fluids and printing inks, and UV-light stabilizers for plastics) under ozonation of Danube river water, containing 160 and 40 ng L
1, respectively. This process removed about 90% of the parent compounds, but oxidation products with MW139, 147 and 165 were formed. Nevertheless, these substances were not detected after a subsequent granulated activated carbon filtration, therefore not raising water safety concerns [65].

24.5 Ubiquity and Regio-Specificity of Transformation Products After all the descriptive information regarding the occurrence and fate of ECs and their TPs is gathered, one of the most difficult issues is to prioritize them and extend these findings to other environmental scenarios. Indeed, a great variability may be found in terms of geographical consumption, seasonal and environmental climatic conditions, contexts of water treatment, flora and fauna exposed individuals, and water safety regulations. Therefore, studies on ECs and TPs must be completed with recognized reference protocols for field investigations and data processing, in order to make relevant comparisons possible. In the European Union (EU), the Water Framework Directive (WFD) enforces that specific pollutants that may cause negative impacts on the water quality should be monitored. Several ECs, including TPs, are now included in the list of priority substances for which environmental quality standards have been established (alkylphenols, organometallic compounds, phthalates and brominated diphenylethers), while pharmaceuticals and estrogens (diclofenac, 17b-estradiol and 17a-ethinylestradiol) are being considered in the periodic revisions [66,67]. According to the results given in the review by Mompelat et al. [7] and our own literature review, most of the studies on TPs of ECs have been produced in Germany, followed by France and UK, which is not surprising as these countries lead the environmental research and host the most renowned researchers in the field. Spain and Switzerland also devote great effort to environmental research, which was reflected in the present literature survey. A variety of studies on PDBEs and respective TPs have taken place in Asia, namely in China [68]. The fast growth of the electronics manufacturing capacities for the global market and the accumulation of a large fraction of electronic waste (E-waste) from developed countries, which are submitted to uncontrolled recycling, have been endangering the environment with PBDEs and their by-products. High levels have been observed in recycling sites and industrial regions [68]. Besides the gaseous emissions, wastewaters are also a source of these substances to surface waters, which enter an air–water equilibrium distribution. The results presented in the previous section regarding the detection of tetrabromobisphenol A diallyl ether in an industrial area attest to the relevance of brominated flame retardants contamination in China [22]. As far as ubiquity is concerned, one of the main gaps hampering wider/global perception is the lack of data in “younger” European and accession countries. The studies conducted by Loos et al. [30,31] at the European Union Joint Research Centre (JRC) are exceptions and they proved the ubiquity of NPE1C and NP in surface- and groundwaters, which were even detected in remote areas.

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 741

It is clear from the present review that results on ECs and TPs are mainly available from Mediterranean and Central European countries, as well as from Asiatic (China and Japan) and American countries (Brazil, Canada and USA). Eastern countries contribute very little to this awareness, however, it is believed that today’s global economy and globalized habits may give rise to similar (qualitative, if not quantitative) contamination problems. In other words, with due exceptions, the ubiquity or regio-specificity may be more a matter of availability of studies, rather than an environmental reality. Furthermore, Europe is crossed by big water streams, which conduct organic micropollutants from industrialized/populated areas to far distances that suffer the indirect impact of these substances.

24.6 Transformation Products of Emerging Contaminants: Fate and Behavior Transformation reactions mediated by biotic and abiotic factors often lead to smaller, more polar and thus less hydrophobic molecules, which are in turn less toxic and less bioaccumulative. In some instances, however, for example, when polar or charged parts of a molecule are cleaved off, or heteroatoms are included in the molecule, as during chlorination in water treatment works, hydrophobicity is increased, leading to increased toxicity. The toxic mechanism (e.g., genotoxicity) may also be altered, indeed as often verified in disinfection byproducts [5]. In evaluating the toxicity of a TP, two dimensions should be appreciated: the toxicokinetics, meaning the behavior of a substance in terms of uptake, distribution, metabolism and excretion; and the toxicodynamics, relating to the type and potency of interaction with the target site which prompts a biological effect. These two components may differentiate the behavior of parent compounds and TPs, although models are being constructed which attempt to predict plausible ranges of toxicity, based on their physico-chemical properties and read-across the behavior of parent/product substances, based on the toxicophore concept [5]. Degradation and sorption of pesticides have been more extensively studied as a requisite for obtaining a marketing license and are made available by official authorities. Nevertheless, new substances are continuously being introduced; therefore, new studies on the degradation and impact of pesticides and their residues need to be conducted. For registration purposes, “relevant metabolites” need to be evaluated in a similar way to pesticides, according to the provisions of Directive 91/414/EEC (Annex VI, point 2.5.1.2) and specific Guidance documents [69,70]. Using an effect-driven approach, Dodd et al. [71] conducted several toxicity studies with antibiotics during the process of oxidation by ozone, a process frequently employed in water treatment works. Similarly to roxithromycin, other antibiotics (macrolides, b-lactams, fluoroquinolones, tetracyclines, lincosamides, aminoglycosides, etc.) oxidized by ozone and hydroxyl radicals either did not give rise to TPs in significant amounts and/or their toxicity was much lower than that of the parent compounds. Exceptions were noted for penicillin G and cephalexin, which formed sulfoxides as first generation TPs, retaining antibacterial activity. However, further transformation led to loss of antibacterial activity [71]. Several other examples can be cited concerning activity loss during photolytic or photocatalytic processes (ciprofloxacin, triclosan, sulfa antibiotics, diuron) or under oxidation processes employed in water treatment works (ethinylestradiol lost its estrogenic activity upon chlorination, bromination and hydroxylation) [5]. Since, in nearly all cases, toxicity was eliminated

742 Transformation Products of Emerging Contaminants in the Environment

concomitantly with parent compounds degradation, almost in a stoichiometric ratio, no further identification and characterization of the TPs should be needed. Nalecz-Jawecki et al. [72] studied the biodegradation of propranolol and no appreciable reduction in toxicity was observed, even when the parent compound was degraded by 70%, leading to the conclusion that the TPs retain their bioactivity. Escher and Fenner [5] compiled several reports on the toxicity of ECs and TPs, mainly pharmaceuticals, under accelerated oxidation conditions (simulated photodegradation, photocatalysis and ozonation), where Vibrio fischeri and Daphnia magna were the preferred test species. Under photocatalysis with TiO2, sulfacetamide, sulfathiazole, sulfamethoxazole, and sulfadiazine did not give rise to toxicologically relevant TPs. Under hypochlorite oxidation, the mutagenicity of frazolidone disappeared, while the mutagenicity of nitrofurazone was reduced. The isomerization and polymerization of isoamylmethoxycinnamate and ethylhexylmethoxycinnamate under irradiation led to a decrease in toxicity. The initial TP 2-ethylhexyl-4-(dimethylamino) benzoate maintained the same toxicity, but further degradation decreased toxicity. Dipyrone and three TPs (4-methylaminoantipyrine, 4-formylaminoantipyrine, 4-acetylaminoantipyrine) were submitted to simulated solar irradiation, giving rise to increased toxicity of the reaction mixture after photolysis. Similar behavior was observed for sulfamethoxazole, whose TPs displayed higher toxicity for Daphnia magna. The ozonation of a solution containing clofibric acid gave rise to an initial increase in toxicity. The often studied, and ubiquitous in the environment, anti-inflammatory drug diclofenac photodegraded to 2-[2-(chlorophenyl)amino] benzaldehyde, which proved to be more toxic than diclofenac itself, due to a higher bioconcentration potential [5]. Oseltamivir carboxylate, a TP and active metabolite of the pro-drug oseltamivir, undergoes little biotransformation or physical removal in wastewater treatment plants: sorption to sediments is likely to be low, given its high hydrophilicity, whereas mineralization by aerobic microorganisms is also limited (1). The biggest contributions to the overall risk stem from the three most toxic compounds, namely nonylphenol, NP2EO and NP1EO [23]. The acids NP2EC and NP1EC exhibited lower single risks. For the special case of nonylphenol polyetholxylates, whose TPs are more toxic than the parent compound itself, the TPs accounted for 89% of the overall risk, according to Fenner et al.’s [23] estimations. Prevedouros et al. [74] reviewed the behavior and fate of perfluorinated compounds, concluding that these chemicals are primarily emitted to water, which is their major reservoir and transport media in the environment, and that they accumulate in surface waters being carried from rivers to sea, due to their very high persistence.

24.7 Conclusions In this section, the gaps in knowledge, need for further research and future trends will be discussed. A major conclusion is highlighted by the present review of the literature, building on the remarks of Escher and Fenner: the presence of TPs in the aquatic environment is not negligible and clearly contributes to the environmental and human health risk of organic micropollutants. Based on the results of a monitoring study of surface and drinking waters, which revealed similar concentrations of transformation products and parent drugs with equivalent pharmacological activity, Jongh et al. [6] strengthened the relevance of monitoring TPs and including them in risk assessment. Gomez et al. [18] and Hadlik et al. [20] added to this, publishing results of TPs in higher concentrations than the respective parent compounds. Furthermore, there is a need to study the evolution of TPs spatially and temporarily in water streams. Also much deeper research is needed targeting the distribution of TPs in different environmental compartments, namely sediments and biota, which may be regarded as sinks for toxic compounds or an alternative entrance route into the trophic chain, respectively. Monitoring of TPs in environmental matrices so far has mostly been restricted to major known by-products, due to the limited availability of authentic standards for both unequivocal identification and quantification. This finding is expressed in the data collated in Table 24.1, where several authors just present qualitative or semi-quantitative information on the occurrence of TPs. Some studies have circumvented this limitation by purifying degraded solutions rather than using synthetic standards. In view of the overwhelming extension of the work to be done on the risk assessment of TPs, more rational and integrated strategies should be followed. One of the options might be the use of a battery of bioassays, including one targeting the specific mode of action of the parent compound, complemented by a nonspecific bioassay sensitive to TPs which have developed high nonspecific toxicity or a new mode of toxicity. Nevertheless, chromatographic/mass spectrometric techniques remain essential tools in the elucidation of TPs, due to their superior sensitivity and accuracy. In specific cases, the nuclear magnetic resonance (NMR) technique is necessary for ultimate identification/confirmation of TPs. However, this

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 745

technique is not yet widely available, requires purified samples, and its sensitivity is inadequate for environmental levels. Hyphenation of NMR to LC instruments is rarely found in environmental laboratories. To tackle the lack of spectra libraries and of efficient common strategies for the identification of new compounds, NORMAN has established a web-based accurate mass spectra database for environmental contaminants, named the MassBank database, which is open for contributions and freely accessible [75]. The generation of models which can cope with possible interactive effects of low concentration mixtures of compounds, namely TPs, displaying similar pharmacological activity, is greatly required in environmental studies. As has been extensively demonstrated throughout this chapter, TPs of organic micropollutants do occur in relevant concentration in surface waters, which are a source of potable water. Therefore, another subject needing major developments is the treatability of affected surface waters using the conventional processes employed in water treatment works. Eventually, due to their higher water solubility and oxidation state, TPs can be more resistant to physical and chemical removal processes. Jin and Peldszus [76] developed a systematic approach for selecting representative micropollutants for evaluating the capacity of treatment during drinking water production which, however, did not include TPs. Although there is a large consensus, derived from risk assessment studies, indicating that very low concentrations of pharmaceuticals in drinking water are very unlikely to pose significant risks to human health, there remain considerable knowledge gaps in assessing the risks associated with long-term, low-level exposures to chemicals and possible combined effects of pharmaceuticals, TPs and a multitude of other xenobiotics. Also, effects on sensitive subpopulations must be considered. In the future, prioritization of the TPs which merit inclusion in the monitoring lists of pollutants is a paramount need, as is already being attempted for parent ECs [77]. All efforts of environmental science, including analytical tasks (identification of relevant degradation products through advanced techniques and strategies, occurrence levels), fate and behavior studies (distribution and toxicological aspects) and risk assessment should converge to supply the data to make this goal feasible. A strong emphasis on the thorough study of the environmental behavior of new synthetic substances, as required by REACH (Registration, Evaluation, Authorization and Restriction of Chemical substances) policy, will alleviate the work load that is left for characterizing threats perceived only once they are already established. In line with the aim of anticipating the risk posed by synthetic chemicals, opposite to the historical reality that dangerous chemicals were just identified several years or decades after first marketing, the Environmental Specimen Banks (ESB) are being regarded as tools for retrospective evaluation of toxic substances, eventually degradation products [78]. Studies of the Swedish ESB on archived guillemot eggs from the Baltic Sea allowed description of the concentration trends of BFRs and perfluorinated compounds. These ESBs are also useful to gage the success of banning the use of certain chemicals [78]. As noted in the workshop “Mixtures and Metabolites of Chemicals of Emerging Concern,” organized under the scope of NORMAN activities in Amsterdam, 2009, pesticide and pharmaceutical degradation products still remain the most studied chemicals. Other groups should be studied. Fatta-Kassinos and Kalavrouziotis highlighted the presence of TPs of xenobiotics as a current concern related to wastewater reuse [79].

746 Transformation Products of Emerging Contaminants in the Environment

As an overall conclusion, the study of TPs requires more clever and productive strategies in order to overcome the challenge of characterizing a massive number of possible substances with uncertain human health and environmental relevance.

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degradation” relationships. Analytical Chemistry, 79, 307. 10. Ragno, G., Risoli, A., Ioele, G., and Luca, M.D. (2006) Photo- and thermal-stability studies on benzimidazole anthelmintics by HPLC and GC-MS. Chemical and Pharmaceutical Bulletin, 54, 802. 11. Cannavan, A., Coyne, R., Kennedy, D.G., and Smith, P. (2000) Concentration of 22,23-dihydroavermectin B1a detected in the sediments at an Atlantic salmon farm using orally administered ivermectin to control sea-lice infestation. Aquaculture, 182, 229. 12. GonSc alves, C., Perez, S., Osorio, V. et al. (2011) Photofate of Oseltamivir (Tamiflu) and Oseltamivir carboxylate under natural and simulated solar irradiation: Kinetics, identification of the transformation products and environmental occurrence. Environmental Science and Technology, 45, 4307. 13. Ghosh, G.C., Nakada, N., Yamashita, N., and Tanaka, H. (2010) Oseltamivir carboxylate, the active metabolite of oseltamivir phosphate (Tamiflu), detected in sewage discharge and river water in Japan. Environmental Health Perspectives, 118, 103. 14. Prasse, C., Schl€usener, M.P., Schulz, R., and Ternes, T. (2010) Antiviral drugs in wastewater and surface waters: A new pharmaceutical class of environmental relevance? Environmental Science and Technology, 44, 1728. 15. Kormos, J.L., Schulz, M., Kohler, H.-P.E., and Ternes, T.A. (2010) Biotransformation of selected iodinated X-ray contrast media and characterization of microbial transformation pathways. Environmental Science and Technology, 44, 4998.

Occurrence of Transformation Products of Emerging Contaminants in Water Resources 747 16. WHO (2011) Pharmaceuticals in drinking-water. WHO/HSE/WSH/11.05 Geneva: World Health Organization. 17. Boleda, M.R., Galceran, M.T., and Ventura, F. (2011) Behavior of pharmaceuticals and drugs of abuse in a drinking water treatment plant (DWTP) using combined conventional and ultrafiltration and reverse osmosis (UF/RO) treatments. Environmental Pollution, 159, 1584. 18. G
omez, M.J., G
omez-Ramos, M.M., Malato, O. et al. (2010) Rapid automated screening, identification and quantification of organic micro-contaminants and their main transformation products in wastewater and river waters using liquid chromatography–quadrupole-time-of-flight mass spectrometry with an accurate-mass database. Journal of Chromatography A, 1217, 7038.
19. Hern
andez, F., Ib
a~nez, M., Pozo, O.J., and Sancho, J.V. (2008) Investigating the presence of pesticide transformation products in water by using liquid chromatography-mass spectrometry with different mass analyzers. Journal of Mass Spectrometry, 43, 173. 20. Hladik, M.L., Hsiao, J.J., and Roberts, L. (2005) Are neutral chloroacetamide herbicide degradates of potential environmental concern? Analysis and occurrence in the upper Chesapeake Bay. Environmental Science and Technology, 39, 6561. 21. Schenker, U., Soltermann, F., Scheringer, M., and Hungerb€ uhler, K. (2008) Modeling the environmental fate of polybrominated diphenyl ethers (PBDEs): the importance of photolysis for the formation of lighter PBDEs. Environmental Science and Technology, 42, 9244. 22. Qu, G., Shi, J., Wang, T. et al. (2011) Identification of tetrabromobisphenol A diallyl ether as an emerging neurotoxicant in environmental samples by bioassay-directed fractionation and HPLCAPCI-MS/MS. Environmental Science and Technology, 45, 5009. 23. Fenner, K., Kooijman, C., Scheringer, M., and Hungerb€ uhler, K. (2002) Including transformation products into the risk assessment for chemicals: the case of nonylphenol ethoxylate usage in Switzerland. Environmental Science and Technology, 36, 1147. 24. Jobling, S., Williams, R., Johnson, A. et al. (2006) Predicted exposures to steroid estrogens in U.K. rivers correlate with widespread sexual disruption in wild fish populations. Environmental Health Perspectives, 114, 32. 25. Legler, J., Dennekamp, M., Vethaak, A.D. et al. (2002) Detection of estrogenic activity in sediment-associated compounds using in vitro reporter gene assays. Science of the Total Environment, 293, 69. 26. Jonkers, N., Laane, R.W.P.M., and De Voogt, P. (2003) Fate of nonylphenol ethoxylates and their metabolites in two Dutch estuaries: evidence of biodegradation in the field. Environmental Science and Technology, 37, 321. 27. Bester, K., Theobald, N., and Schroder, H.F. (2001) Nonylphenols, nonylphenol-ethoxylates, linear alkylbenzene sulfonates (LAS) and bis(4-chlorophenyl)-sulfone in the German Bight of the North Sea. Chemosphere, 45, 817. 28. Jonkers, N., Laane, R.W.P.M., and Voogt, P.d. (2005) Sources and fate of nonylphenol ethoxylates and their metabolites in the Dutch coastal zone of the North Sea. Marine Chemistry, 96, 115. 29. Tubau, I., V
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748 Transformation Products of Emerging Contaminants in the Environment 34. Berger, U., Kaiser, M.A., and Barber, J.L. (2009) Analytical developments and monitoring of polyand perfluoroalkyl substances. NORMAN Bulletin, 1, 11. 35. Lee, Y. and Gunten, U.v. (2009) Transformation of 17a-ethinylestradiol during water chlorination: Effects of bromide on kinetics, products, and transformation pathways. Environmental Science and Technology, 43, 480. 36. Kern, S., Fenner, K., Singer, H.P. et al. (2009) Identification of transformation products of organic contaminants in natural waters by computer-aided prediction and high-resolution mass spectrometry. Environmental Science and Technology, 43, 7039. 37. P
erez-Parada, A., Ag€uera, A., G
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750 Transformation Products of Emerging Contaminants in the Environment 72. Nalecz-Jawecki, G., Wojcik, T., and Sawicki, J. (2008) Evaluation of in vitro biotransformation of propranolol with HPLC, MS/MS, and two bioassays. Environmental Toxicology, 23, 52. 73. Singer, A.C., Nunn, M.A., Gould, E.A., and Johnson, A.C. (2007) Potential risks associated with the proposed widespread use of Tamiflu. Environmental Health Perspectives, 115, 102. 74. Prevedouros, K., Cousins, I.T., Buck, R.C., and Korzeniowski, S.H. (2006) Sources, fate and transport of perfluorocarboxylates. Environmental Science and Technology, 40, 32. 75. Schulze, T., Schymanski, E., Stravs, M. et al. (2012) NORMAN MassBank Towards a communitydriven, open-access accurate mass spectral database for the identification of emerging pollutants. NORMAN Bulletin, 3, 9. 76. Jin, X. and Peldszus, S. (2012) Selection of representative emerging micropollutants for drinking water treatment studies: A systematic approach. Science of the Total Environment, 414, 653. 77. Brack, W. (2012) Emerging substances of toxicological concern in a world full of chemicals - the NORMAN way to find the needles in the haystack. NORMAN Bulletin, 3, 1. 78. Ruedel, H. (2009) Environmental specimen banks as tools for the retrospective monitoring of emerging pollutants. NORMAN Bulletin, 1, 2. 79. Fatta-Kassinos, D. and Kalavrouziotis, I.K. (2011) Current concerns related to wastewater reuse and xenobiotics. NORMAN Bulletin, 2, 7. 80. Miao, X.S. and Metcalfe, C.D. (2003) Current concerns related to wastewater reuse and xenobiotics. Analytical Chemistry, 75, 3731. 81. Corcia, A.D., Costantino, A., Crescenzi, C. et al. (1998) Characterization of recalcitrant intermediates from biotransformation of the branched alkyl side chain of nonylphenol ethoxylate surfactants. Environmental Science and Technology, 32, 2401. 82. Ding, W.H. and Tzing, S.H. (1998) Analysis of nonylphenol polyethoxylates and their degradation products in river water and sewage effluent by gas chromatography–ion trap (tandem) mass spectrometry with electron impact and chemical ionization. Journal of Chromatography A, 824, 79. 83. Schmidt, C.K. and Brauch, H.-J. (2008) N, N-Dimethylsulfamide as precursor for N-Nitrosodimethylamine (NDMA), formation upon ozonation and its fate during drinking water treatment. Environmental Science and Technology, 42, 6340.

25 Occurrence of Transformation Products of Emerging Contaminants in Water Resources of the United States Imma Ferrer and E. Michael Thurman Center for Environmental Mass Spectrometry, University of Colorado, USA

25.1 Introduction: Emerging Contaminants The identification of emerging contaminants (ECs) in water samples has been the focus of many water agencies and water treatment facilities around the world. Specifically, in the United States, the Environmental Protection Agency (EPA) has guided and released new regulations [1] in order to narrow the contaminant candidate list (CCL3) to possible toxic emerging compounds of interest. Most recently, a new candidate list called “The Third Unregulated Contaminant Monitoring Rule (UCMR 3)” from EPA was launched in May 2012 [2]. The Unregulated Contaminant Monitoring Rule (UCMR) provides EPA and other interested parties with scientifically valid data on the occurrence of contaminants in drinking water. These data serve as a primary source of occurrence and exposure information that the agency uses to develop regulatory decisions. UCMR 3 monitoring will take place from 2013 to 2015, and includes monitoring for 28 chemicals and 2 viruses. Regulatory water agencies will be required to report concentrations for these contaminants in the near future. No pharmaceuticals are included in this recent list, only hormones. But in the meantime, a trend to try and detect as many compounds as possible in environmental water sources has become the challenge. Pharmaceuticals found in water samples, due to human discharge (via direct or indirect sources), are by far the most extensive range of ECs reported to date. In the last 10 years, pharmaceuticals have been extensively detected in surface water in Europe [3–5] and in the United States [6]. The results of the reconnaissance by the U.S. Geological Survey showed

Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

752 Transformation Products of Emerging Contaminants in the Environment

that 80% of all surface water had detectable concentrations of pharmaceutical compounds. Approximately 82 compounds were detected, including steroids, antibiotics, analgesics, heart medications, and other compounds [6]. Typically the concentrations are in the submicrogram-per-liter range. It has been over a decade since the first U.S. survey of water samples [6] was carried out. This paper is still the most cited paper in the history of pharmaceuticals in water samples. Since then, many other papers and reviews have reported identifications of several of the ECs of concern. In general, there is a trend in the literature to only report and measure already known and published ECs. Only a few studies have reported newly identified and discovered pharmaceutical compounds and their transformation products (TPs) [7–9]. It is important to mention that sometimes TPs or metabolites exceed the concentrations of the parent compounds, becoming then more environmentally relevant than the starting active ingredients. The majority of the pharmaceuticals identified in environmental samples have been detected using liquid chromatography-mass spectrometry (LC/MS). Most specifically, the advent of time-of-flight techniques (TOF) applied to environmental analyses has just begun in the last few years [10]. Applications range from routine analytical methods that analyze a few target compounds to more extensive methods that include a variety of analytes, including also non-target and unknown identification. Due to the high complexity of some environmental samples (i.e., wastewater, sludge samples, soil samples), high-resolution techniques with additional structural information on fragment ions are needed. These techniques provide a high degree of confidence for identification of target analytes and aid the structural elucidation of TPs and unknown compounds, which are usually present in environmental samples. The possibility of creating universal accurate mass databases with time-of-flight analyses for sets of compounds has broadened the range of applications as well, going from target to non-target identification. This chapter gives an overview of the different tools used in LC/MS, with a specific focus on time-of-flight techniques, and the applications that have recently generated in the environmental field. This manuscript gives several examples of EC analysis that exemplify the unique features of time-of-flight for the identification of non-target and unknown compounds. LC/MS employing accurate mass measurement has been proved as a successful technique for both quantitative analysis of target compounds and rapid qualitative analysis of “unknown” environmental mixtures.

25.2 State-of-the-Art Techniques for the Identification of Emerging Contaminants and Their Transformation Products There is no doubt that LC-MS has been the universal method of choice when analyzing ECs and TPs in environmental water samples for at least 20 years now. The most popular technique for identification and confirmation of pharmaceuticals is tandem mass spectrometry (LC/MS-MS) using either collision cells or linear traps to obtain information on fragment ions. This technique is more focused toward target analysis, where the analyst is looking at a specific group of analytes that may vary from a few analytes within a family (3–4) to large multiresidue methods (>100). However, sensitivity usually becomes an issue when targeting a large number of compounds. This is the reason why time-of-flight mass spectrometry techniques have become popular in the last few years, since they give full-spectra data at all times. A large number of compounds (virtually no limit) can be analyzed in a single run, while obtaining valuable accurate mass information for each compound that ionizes. Furthermore,

Occurrence of Transformation Products of Emerging Contaminants 753

extra information on metabolites or TPs can be achieved by exploring the accurate mass spectra of unknown peaks in the chromatogram. In this section we will discuss both techniques of detection, a targeted approach using LC/MS-MS techniques and a non-targeted tactic for the discovery and identification of relevant compounds using LC/TOF-MS. 25.2.1 Liquid Chromatography/Tandem mass Spectrometry (LC/MS-MS) for the Analysis of Target Compounds. EPA Method 1694 LC/MS-MS using linear traps and triple quadrupoles seems to be the preferred method for routine analysis of pharmaceutical compounds in environmental samples. Overall, hundreds of papers have been published reporting findings of pharmaceuticals in nontreated and treated waters [11]. However, in spite of the numerous papers reporting analysis of pharmaceuticals, no analytical methodology seems to be the preferred one as a standardized methodology for these types of compounds until recently. EPA Method 1694 [12] was published in December 2007 as a guiding and screening method for those scientists analyzing pharmaceuticals in environmental samples. The standard EPA protocol uses solid-phase extraction (SPE) for water samples followed by the analysis of extracts by tandem mass spectrometry using a single transition for each compound, with retention time guidelines for identification. We applied EPA Method 1694 in our lab for the analysis of pharmaceuticals in wastewater, surface water and drinking water samples [13]. The implementation for this method consisted of the analysis of 70 analytes (of 75 total analytes in the original method) and 18 labeled internal standards (of 20 total), which are a mixture of pharmaceuticals and personal care products that are analyzed by LC/MS-MS. In our work we addressed some of the analytical issues that were not covered in the original method, such as degradation of some compounds in solvent mixtures and assignment of a second transition for multiple reaction monitoring transitions (MRM) for additional mass spectrometry quality assurance. The main goal of this work was to show the usefulness of the EPA method for generic screening and monitoring of pharmaceuticals in water and wastewater. The method was applied to the analysis of several drinking water, surface water and wastewater samples from several locations in Colorado, USA. Surprisingly, only 8 out of the 70 compounds were consistently found in environmental water samples: caffeine, carbamazepine, clarithromycin, diltiazem, diphenhydramine, erythromycin, sulfamethoxazole and trimethoprim, which were confirmed with two MRM transitions. The results for the concentrations found are shown in Table 25.1. These samples are representative of several inputs of wastewater contamination. One drinking water sample was also analyzed, and gave a positive hit for carbamazepine, a common antiepileptic and anti-depressant prescribed drug. Since then, we have refined our target methods using triple quadrupole mass spectrometry for a subset of 25–30 compounds that are regularly found in surface and wastewater samples. From these 30 compounds, usually 18–20 analytes are always found in surface and groundwater impacted by wastewater sources. It is the view of the authors that confirmation of positive identifications in real samples requires the additional second MRM transition and the evaluation of ion ratios between the two monitored transitions as compared to a reference standard [14]. Confirmation of the identity of target analytes in real samples is usually based on ion ratio statistics for the transitions monitored. Thus, the confirmation criteria using tandem mass spectrometry cover a range of maximum permitted tolerances according to relative ion intensity, expressed as a percentage of the intensity of the most intense transition [13,14].

Occurrence of Transformation Products of Emerging Contaminants 755

25.2.2 Liquid Chromatography/Time-of-Flight/Mass Spectrometry (LC/TOF-MS) for the Analysis of Non-target Compounds Recently, LC/TOF-MS has been used for the unequivocal confirmation of contaminants (including pharmaceuticals, pesticides and surfactants) in a variety of samples, such as water and sediments [10] by accurate mass measurement of protonated molecules. In this sense, several authors have reported accurate mass confirmation of pharmaceuticals in surface and wastewater samples [15–19] as well as sediment and sludge [20] using time-of-flight techniques. Detection of drugs in urine is also one of the topics that have been recently covered by LC/TOF-MS techniques [21–24]. In many of these studies time-of-flight techniques were successfully used for the unequivocal identification of TPs of known contaminants, as well as unknown compounds [25–28]. It is worth mentioning also several applications of time-offlight mass analysis for the identification and confirmation of metabolites or TPs of pesticides and pharmaceuticals in environmental samples [29–39]. One of the main reasons that TOF has become so popular in the last few years is because accurate mass measurements are specific and universal for any kind of analyte and do not depend on the type, brand, or specific instrumentation used. The degree of fragmentation may vary depending on the instrument used, but the specific accurate mass value and/or accurate isotope information will be consistent for a given analyte, no matter what type of ionization, collision induced dissociation and MS-MS fragmentation is used. Accurate mass determination allows one to obtain unique information for a given molecule, plus additional information from isotopic patterns, mass defect and specific fragment ions [10]. Sometimes, a single stage time-of-flight mass analyzer (TOF/MS) generates valuable information by imparting enough energy into the [M þ H]þ ions in the source region to cause fragmentation [39]. Time-of-flight mass analysis generates increased resolving power of signals on the m/z axis in comparison to quadrupole mass spectrometers. Furthermore, this enhanced resolving power benefits analyses involving complex environmental matrices by separating isobaric interferences from the contaminant signals of interest. The improved resolution also facilitates the measurement of accurate masses within 3 ppm, which are accepted for the verification of elemental compositions. Elemental compositions of contaminants and their fragment ions clearly constitute higher order identifications than those obtained by nominal mass measurements. 25.2.3 Liquid Chromatography/Quadrupole-Time-of-Flight/Mass Spectrometry (LC/Q-TOF-MS) for Structural Elucidation of Unknown Compounds and Transformation Products Most published methods only include information on the exact mass of the protonated or deprotonated molecule, a few report just one fragment ion per compound. To our knowledge, no studies also include accurate mass information of more than one fragment ion obtained by MS-MS for a large number of compounds (>80). Only recently, an extensive accurate mass library was developed and commercialized by Broecker et al. [40] for more than 2500 compounds. Another study by our group compiled information on a 100 pharmaceutical compounds, including detailed data on fragment ions obtained by a Q-TOF-MS instrument [41]. We also included a total of 16 different metabolites for the most environmentally relevant pharmaceuticals. Accurate mass information for each compound was obtained and compiled in an extensive table. Accurate mass measurements of fragment ions become particularly important in the structure elucidation of non-targets and unknowns. In this sense, the Q-TOF MS/MS is unique

756 Transformation Products of Emerging Contaminants in the Environment

among TOF instruments in its ability to give accurate mass measurements (1 to 2 millimass units) of the fragment ions that are ejected from the collision chamber. This is very useful when trying to elucidate the identity of unknown or non-target compounds, the more fragment accurate mass information one can get from time-of-flight mass techniques the better understanding for the structural elucidation of a certain compound. The same reasoning applies to the elucidation of possible TPs. When knowing what the starting compound is, the information about fragment ions and their accurate masses will play an important role in deciphering the chemical structure of the metabolite or TP. Another important tool that has made TOF one of the key methodologies for identification of compounds is the existence of accurate mass databases, as published extensively. An individual scientist can apply these universal databases to each particular problem and then often get an identification of the analyte of interest [42–44]. Other tools, that are available with TOF instrumentation, and will be discussed in this chapter, include the use of molecular features, accurate mass filters and the isotopic mass defect, and the use of mass profiling to distinguish between control samples and positive samples. Examples will be given for each one of these accurate mass tools.

25.3 Use of Accurate Mass Tools for the Identification of Emerging Contaminants 25.3.1 Molecular Features For many years, the use of reverse-search methods for gas chromatography/mass spectrometry (GC/MS) has made it possible to search large National Institute for Standards and Testing (NIST) pesticide libraries in minutes [45] and has made screening quite simple for pesticides amenable to GC/MS. Unfortunately similar reverse-search methods have not been available for LC/MS for two reasons. First, the single quadrupole and triple quadrupole mass spectrometers do not operate in full scan mode for pesticide screening because of a lack of sensitivity [46]. Secondly, although libraries for LC/MS three-dimensional ion trap have been made, they have not been popular due to difficulties in reproducibility of fragmentation and the need for authentic standard analysis for each instrument [47–49]. So, the only approach that uses full spectrum information is LC/TOF-MS, which is both sensitive and accurate [50], but uses only the accurate mass of the [M þ H]þ ion. The combination of accurate mass and sensitivity is needed for screening of compounds by their empirical formula. The MFE software compiles accurate mass ions, excludes background noise, and plots extracted ion chromatograms of the most intense peaks found in a chromatogram. So a molecular feature is defined as a discrete molecular entity defined by combination of retention time, mass and response in an LC/MS analysis. In general, MFE operates on raw mass spectral data generating lists of chemically qualified molecular features (background is removed, interferences are resolved, isotopic cluster and molecular adducts are recognized). The screening criteria usually consist of  5 ppm accurate mass window,  0.2 min retention time window, and a minimum 1000 counts (signal to noise of 10:1). The ions are grouped by entities that include common adducts (sodium, ammonia, etc.) and isotope clusters. As an example, Figure 25.1 shows the total ion current chromatogram (a) and the molecular feature extraction (b) for a surface water sample taken in the South Platte River (Colorado, USA). As can be seen in this figure, a total of 1498 ion features were found in the chromatogram. One can generate as many empirical formulae as wanted and from there one

758 Transformation Products of Emerging Contaminants in the Environment

can try to elucidate the chemical structure. But, the most common approach is to compare the data obtained to a known database to try to match as many compounds as possible. This approach is explained in the next section. Strengths of the MFE include rapid screening of 100 compounds at sensitive levels compared to a manual approach and the ease of use of the database for any accurate mass spectrometer instrumentation capable of routine sub 5-ppm mass accuracy. 25.3.2 Accurate Mass Databases The pioneer efforts to search data using an accurate mass database were made by several authors, such as Thurman et al. [28], Bodeldijk et al. [51], Ojanpera et al. [24]. For example, Thurman et al. [28] used an approach of TOF, ion trap, and the Merck Index database to identify pesticides in food and also TPs, without the initial use of primary standards [29]. Bobeldijk et al. also used the Merck Index, the NIST library, and their own database to screen water pollutants [51]. The methods in these examples rely on manually searching the databases, compound by compound. Recently, several papers [24,52] have extended this approach, using mass accuracy of 30 ppm and database analysis to identify 600 drugs in blood and urine without the use of primary standards, using only the protonated molecule. In spite of the progress that has been made, the ability to do true library analysis is still a problem to be solved for LC/MS and for rapid analysis of environmental samples. The problems to be overcome include reproducible spectra and ion ratios, routine programs for rapid screening of samples rather than manual checking of data, and some estimate of the probability of the correct identification. Variation in fragmentation intensity is not critical with the use of accurate mass, since the accurate mass of the fragment ion gives its molecular formula. In fact, accurate mass measurements are specific and universal for every target analyte, regardless of the instrumentation used. Usually, unambiguous identification is accomplished by means of accurate mass measurements from (de)protonated molecules, fragment ions, and isotope intensity/signature matching. Thus, the accurate mass database approach is a screening tool, and it is powerful and fast because only the molecular formula is needed. Here we will describe two different approaches to the use of databases. The first uses an automated molecular feature database, as described in the last section, and then a commercial or homemade database based on a csv file. The second approach is called “reversed database search” in which a total ion chromatogram is searched for ions included in such a database. Databases usually contain information of the monoisotopic exact mass of the MHþ, at least one product ion, and retention time of the compound. The advantages and limitations of both approaches are discussed, as well as the reliability (match probability) of a database search using accurate mass. Following the first approach (as described in the last section) a compilation of ions was gathered through a molecular feature extraction (see Figure 25.1) for one of the surface water samples collected in the South Platte River. The next step was to match these accurate masses with any compounds included in a commercial database (called “forensics database” from Agilent Inc.). As shown in Figure 25.2, one of the hits obtained was for gabapentin, a common anti-epileptic pharmaceutical. In the insert, a detailed mass spectrum is plotted. Each one of the fragments was elucidated from the parent chemical structure (see Figure 25.3). The fact that this molecule gave such a rich CID spectrum allowed the detailed fragmentation elucidation by using accurate mass, as seen in Figure 25.3. The protonated molecule loses water, followed by ammonia, and then an additional water loss occurs, followed by a loss of C4H4 to give the smallest fragment. This example shows how the use of

760 Transformation Products of Emerging Contaminants in the Environment Gabapentin O

O

O

HO

–H2O Exact Mass: 18.0106

–NH3 Exact Mass: 17.0265 NH2

NH2

+

C9H13O

C9H18NO2+ C9H18NO2+

Exact Mass: 137.0961

Exact Mass: 154.1226

Exact Mass: 172.1332

–H2O Exact Mass: 18.0106

–C4H4 Exact Mass: 52.0313

+

C5H7 Exact Mass: 67.0542

C9H11

+

Exact Mass: 119.0855

Figure 25.3 Fragmentation pathway for gabapentin.

accurate fragmentation information can be used as a complementary set of data to provide unambiguous confirmation of the new finding. This is an example of a successful hit using this approach as, when this experiment was carried out, no commercial standard was available at our lab. Later, we purchased the standard and confirmed, both by retention time and accurate mass, as well as fragmentation pattern, the accuracy of the finding. The second approach, also known as “reversed database search,” consists of running the database to see if any positive hits are found. This approach is usually slower than the last one as each one of the database entries has to be verified against each of the extracted ion chromatograms for a given accurate mass. This automatic screening method requires a thorough full optimization of the accurate-mass window used and retention time (always optional) tolerances, which play an important role in the selectivity, accuracy, and successfulness of the whole procedure. Using this approach, and by running the same database as mentioned before, we verified the presence of one of the metabolites of dextromethorphan, also known as dextrorphan, in a surface water sample impacted by a wastewater source. Figure 25.4 depicts an excerpt of the automated generated report of a database search. It is important to note the high score obtained for this particular hit. This score is a combination of mass accuracy, isotope intensity and isotope matching. Also, as shown in the figure a good mass accuracy (with an error below 2 ppm) was obtained for this identification, thus confirming the presence of this compound in the sample. Again, in this case, no standard had been analyzed by this instrument when this finding was made, so a pure standard was purchased, analyzed, and we verified this positive identification in a water sample.

762 Transformation Products of Emerging Contaminants in the Environment

25.3.3 Accurate Mass Filters and Isotopic Mass Defect Chlorine appears in many pesticides and pharmaceutical products that are important to environmental analysis. Because chlorine contains two isotopes, Cl35 and Cl37, there is a distinctive A þ 2 isotope pattern that is generated by a single chlorine atom in a molecule. Furthermore, there is an isotopic mass defect that occurs with chlorine-37 that makes the identification of chlorine in a molecule relatively easy [53]. More than one chlorine atom in a molecule generates an A þ 2 and A þ 4 isotopic pattern, which is characteristic and commonly shown in all mass spectrometry books as a key to compound identification of chlorinated compounds [54]. Using this rule, a chlorine mass filter was developed by our group [55]. The chlorine mass-filter is used to screen both LC/TOF-MS and LC/QTOF-MS data files in order to discover compounds that contain chlorine. The chlorine filter uses MassHunter software to generate formulae for chlorine-containing compounds. An example is given for a wastewater sample. The initial identification of lamotrigine, an anti-depressant pharmaceutical not previously reported in water samples, was accomplished using the mass-defect filter that looked for chlorinated analytes in the extract of a wastewater sample after LC/TOF-MS analysis in MS-only mode. The mass defect filter essentially looks at the accurate mass of the monoisotopic mass of an analyte and the A þ 2 isotopic mass. Both the intensity and the accurate mass are used to detect chlorinated compounds using the mass defect filter. In the case of lamotrigine, the mass defect filter detected a peak at 13.7 min with a mass of m/z 256.0153 and an A þ 2 isotope with a mass of m/z 258.0122 and an intensity of 66% (see Figure 25.5). The mass defect filter showed that the A þ 2 peak had a relative isotopic mass defect of 0.0030 u, indicating a chlorinated compound with two chlorine atoms [29,56]. The second step after the mass defect filter was to determine the molecular formula of the unknown chlorinated compound. The best fit for the ion formula was C9H8Cl2N5 with a match of 99 out of 100 based on MassHunter Software, which evaluates the accurate mass of the A ion, the isotope intensity matching, and isotope spacing (also called the isotopic mass defect) or accurate mass of the isotopes. The neutral formula, C9H7Cl2N5, was then run through the Merck Index database for a formula match and gave lamotrigine as its only formula. When the formula was put through a much larger database, ChemSpider, the match was for 65 compounds; however, there were only 13 patented structures and only 1 compound was listed in a Wikipedia-available article and that was lamotrigine. A quick read showed that this compound is the number three most used bipolar medication in the US at this time; thus, it was given the most likelihood of a correct identification. Later, a standard was purchased and the identification was verified [7]. Also, a metabolite of lamotrigine was discovered using the same procedure as described above. A second peak of much less intensity and earlier retention time (9.9 min) had been detected in the 256 m/z extracted ion chromatograms of several wastewater samples containing lamotrigine. The spectrum of this peak revealed a much larger ion at 432.0472 m/z, thus an MS-MS experiment was carried out to confirm that the 256 ion formed indeed from the 432 ion. A literature search for the empirical formula C15H15Cl2N5O6 (at 432.0472 m/z) revealed that this was a potential glucuronide metabolite of lamotrigine [57,58]. The finding was verified by analyzing a pure standard of 2-N-glucuronide lamotrigine. A second MS-MS experiment was performed to fragment the ion at m/z 256, hence simulating a pseudo MS3 experiment, and a spectrum that matched that of lamotrigine was obtained, thus totally confirming the identification of the 2-N-glucuronide metabolite (see Figure 25.6).

Occurrence of Transformation Products of Emerging Contaminants 765

The combination of mass accuracy, database matching, and identifying a fragment ion shows the power of using the chlorine mass-filter to find and identify trace chlorinated substituents in water samples impacted by wastewater. This approach works really well for complex water matrices by identifying specific chlorinated compounds, which in turn could be potential metabolites from known target analytes. 25.3.4 Accurate Mass Profiling Urine metabolic profiling combined with LC/QTOF-MS was used to find and identify the metabolites of dextromethorphan, a common over-the-counter (OTC) cough suppressant [8]. Chromatograms of both blank urine and urine taken 4 h after ingestion of dextromethorphan were compared using Mass Profiler software. The software first analyzes all groups of ions (known as features) in the chromatogram of both samples and compiles this into a database. Three replicates of each sample are taken and averaged. Next the software compares the two samples, looking for features (plotted as gray dots) that are unique to the dextromethorphan urine (Figure 25.7a). The

Figure 25.7 (a) Mass profiler plot of a urine sample 4 h after taking a 10 mg dose of dextromethorphan. (b) Extracted ion chromatograms of seven major glucuronide metabolites.

766 Transformation Products of Emerging Contaminants in the Environment Table 25.2 Compounds, formulae, exact mass, measured mass, error, and structures of dextromethorphan and its metabolites. Reproduced with permission from [8] Copyright (2012) Elsevier Ltd. Compound Name

Formulae (MH þ )

Exact mass MHþ (m/z)

Measured mass MHþ (m/z)

Error (ppm)

Dextromethorphan

C18H26NO

272.2009

272.2010

0.4

Structure of the compound

O

H N

Dextrorphan

C17H24NO

258.1852

258.1854

0.8

HO

H N

N-demethyldextrorphan C16H22NO

244.1696

244.1699

1.2

HO

H NH

Dextrorphan Glucuronide

C23H32NO7

434.2173

434.2175

0.5

GluO

H N

N-demethyldextrorphan C22H30NO7 Glucuronide

420.2017

420.2016

0.2

GluO

H NH

comparison resulted in 27 features that were unique to this sample, and 136 individual ions. Ions at the same retention time, for example, 15.1 min, were the same fragment ions of a feature, based on MS/MS analysis discussed in [8]. Figure 25.7b shows a chromatogram with 7 major metabolites identified. Table 25.2 shows the structure and accurate masses of dextromethorphan and its four major metabolites, which are reported in the pharmaceutical literature [59–62]. The metabolites are dextrorphan and N-demethyldextrorphan and glucuronides of each of these two compounds. The calculated exact masses for each of these compounds were extracted from

Occurrence of Transformation Products of Emerging Contaminants 767

the total ion chromatogram of the positive urine sample and compared to the measured masses (data shown in Table 25.2). The measured masses for the protonated molecule of each compound varied from 0.1 to 0.3 mmu, which is 1 ppm mass accuracy or less for all targeted compounds.

25.4 Occurrence of Transformation Products in Environmental Waters in the US Our lab has analyzed several hundred samples for pharmaceuticals in the last 2–3 years from different locations in the US and comprising different types of water samples (drinking, groundwater, surface water, lake water and wastewater) by LC/TOF-MS and LC/Q-TOF-MS. The majority of drinking and groundwater samples contain few or no pharmaceuticals, with the exceptions of those cases where groundwater comes from production wells along main rivers or reservoirs. The main detections of pharmaceuticals usually occur in surface water samples and wastewater samples. Interestingly, from all the pharmaceuticals analyzed, the same set of compounds occurs in the majority of water samples [41]. Table 25.3 shows the most representative data from surface water samples collected during 2011 in the U.S. After analyzing a large number of samples we have come up with some findings (new compounds detected and new metabolites) that are worth mentioning here and this is why these compounds were included in previous data sets [41]. Identities of compounds were based on retention time and accurate mass of the protonated/deprotonated molecules and their fragment ions. MS-MS acquisition was performed on those cases where a new compound or metabolite was discovered. For example, a new finding was the anti-convulsant (also used as anti-depressant) lamotrigine and its N2-glucuronide found in wastewater, surface water and even groundwater samples [20]. To date no other environmental reports of this pharmaceutical and/or metabolite have been reported in the literature. This compound is frequently detected in water samples (see Table 25.3) and at high concentrations, suggesting that it is replacing the “older” anti-convulsant drugs (carbamazepine, citalopram, fluoxetine, etc.) prescribed for human intake. Other findings include metabolites of already well-known drugs, such as bupropion, carbamazepine and venlafaxine, to mention a few. These are important findings as the metabolite concentrations often exceed the parent compound concentration. Figure 25.8 shows an example of a common detected drug (metoprolol) and its newly identified acid metabolite in a surface water sample. The MS-MS experiments at 30 V revealed the most important fragments of this metabolite (as shown in the inset spectrum). Finally, we have summarized the most important findings in Table 25.3, which represents the study of a sub-selected and representative set of approximately 100 surface water samples (downstream of effluent discharge) analyzed by LC/Q-TOF-MS. The percent of detections for each compound found in the water samples is depicted by a percentile number. From this table we can conclude that about 36 different pharmaceuticals are commonly detected in waters impacted by wastewater sources. Some of the pharmaceuticals not included in this table have never been detected, or detections were lower than 10%. Compounds such as carbamazepine, bupropion, lamotrigine, diphenhydramine, gemfibrozil, metoprolol, propanolol, sulfamethoxazole, thiabendazole, trimethoprim, venlafaxine, and their respective metabolites, are the most common pharmaceuticals detected in water

768 Transformation Products of Emerging Contaminants in the Environment Table 25.3 Detections (%) of the most commonly found pharmaceuticals in water samples (surface water downstream from effluent discharge). Reproduced with permission from [41] Copyright (2012) Elsevier Ltd. Compound

Detections in water samples (%)

Average concentration (ng/L)

1,7-Dimethylxanthine 10,11-Dihydroxy-carbamazepine 10-Hydroxy-carbamazepine Atenolol Bupropion Caffeine Carbamazepine Cetirizine Citalopram Clarithromycin Cotinine Demethyl-dextrorphan Des-venlafaxine Dextrorphan Diltiazem Diphenhydramine Erythrohydrobupropion Erythromycin Erythromycin Anhydrate Fluoxetine Gabapentin Gemfibrozil Hydroxy-bupropion Ibuprofen Lamotrigine Metoprolol Metoprolol acid 2N-glucuronide lamotrigine Naproxen Nor-citalopram Propanolol Sulfamethoxazole Thiabendazole Triclocarban Trimethoprim Venlafaxine

10 45 85 74 68 70 95 82 79 75 22 65 78 75 69 80 78 55 35 25 44 74 75 20 97 91 85 68 64 66 88 95 75 64 76 78

110 80 255 166 140 220 350 70 85 46 40 10 84 50 47 57 180 137 62 65 54 95 150 21 455 237 74 95 22 74 53 320 188 96 264 310

samples. Also shown in this table is the average concentration found for each one of the compounds analyzed. The highest concentrations were those corresponding to the detections of anti-convulsants, anti-depressants, psychiatric drugs and beta-blockers in water samples. More environmental studies would be needed to understand the fate and transport of these type of pharmaceuticals in the environment, especially those of new appearance.

770 Transformation Products of Emerging Contaminants in the Environment

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Occurrence of Transformation Products of Emerging Contaminants 773 55. Agilent Technologies Application Note: Using a chlorine filter for accurate-mass data analysis of environmental samples. 56. Thurman, E.M., Ferrer, I., and Zweigenbaum, J.A. (2006) High resolution and accurate mass analysis of xenobiotics in food. Analytical Chemistry, 78, 6702–6708. 57. Ohman, I., Beck, O., Vitols, S., and Tomson, T. (2008) Plasma concentrations of lamotrigine and its 2-N-glucuronide metabolite during pregnancy in women with epilepsy. Epilepsia, 49, 1075–1080. 58. Saracino, M.A., Bugamelli, F., Conti, M. et al. (2007) Rapid HPLC analysis of the antiepileptic lamotrigine and its metabolites in human plasma. Journal of Separation Science, 30, 2249–2255. 59. Eichhold, T.H., Greenfield, L.J., Hoke, S.H.II, and Wehmeyer, K.R. (1997) Determination of dextromethorphan and dextrorphan in human plasma by liquid chromatography tandem mass spectrometry. Journal of Mass Spectrometry, 32, 1205–1211. 60. Lutz, U., Bittner, N., Lutz, R.W., and Lutz, W.K. (2008) Metabolite profiling in human urine by LC-MS/MS: Method optimization and application for glucuronides from dextromethorphan metabolism. Journal of Chromatography B, 871, 349–356. 61. Lin, S., Chen, C.H., Ho, H.O. et al. (2007) Simultaneous analysis of dextromethorphan and its three metabolites in human plasma using an improved HPLC method with fluorometric detection. Journal of Chromatography B, 859, 141–146. 62. Kikura-Hanajiri, R., Kawamura, M., Miyajima, A. et al. (2011) Chiral analyses of dextromethorphan/levomethorphan and their metabolites in rat and human samples using LC-MS/MS. Analytical and Bioanalytical Chemistry, 400, 165–174.

26 Spatial Modeling for Elucidation of Perfluorinated Compound Sources and Fate in a Watershed Yasuyuki Zushi1 and Shigeki Masunaga2 1

Center for Environmental Measurement, Organochemical Measurement Laboratory, National Institute for Environmental Studies, Japan 2 Graduate School of Environment and Information Sciences, Yokohama National University, Japan

26.1 Introduction 26.1.1 Transformation Products of PFCs Perfluorooctane sulfonate (PFOS) has come to be known as a new persistent organic pollutant (POP) owing to its persistence and bio-accumulative and toxic properties. In addition, PFOS has been listed in the Stockholm Convention on POPs since 2009 with its precursor, perfluorooctane sulfonyl fluoride (PFOSF) [1]. Several chemicals are similar to PFOS, including perfluorooctanoate (PFOA), its precursors (perfluorooctane sulfonamides (FOSAs), perfluorooctane sulfonamidoethanols (FOSEs), fluorotelomer alcohols (FTOHs)), and its synthetic starting material (PFOSF) (Table 26.1). Moreover, homologs of PFOS and PFOA with different chain lengths, including branched chain isomers, and their precursors exist. These chemicals are known as perfluorinated compounds (PFCs) or perfluoroalkyl substances, and some have been used in industrial and commercial products since the 1950s. FOSAs, FOSEs and FOSAs are known transformation products (TPs) of FOSA acrylate polymer. Recently, perfluorooctane sulfonamidoethanol-based phosphate (SAmPAP) was reported to be a precursor of those compounds together with FOSA acrylate [2,3]. These compounds ultimately degrade to PFOS and/or PFOA via photodegradation in the atmosphere [4] and aquatic environments [5], as well as through biodegradation [6]. Conversely, Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 779

FTOHs, fluorotelomer carboxylates and fluorotelomer unsaturated carboxylates are known TPs of FT acrylate polymer [7] or polyfluoroalkyl phosphate esters [8]. These compounds degrade to PFOA through photodegradation in the atmosphere [9] and aquatic phase [10], and via biodegradation [8,11]. Moreover, x:3 polyfluorocarboxylates (x:3 PFCA) only degrade to PFOA through biodegradation and are expected to be markers of biodegradation pathways in the environmental process [12]. Most of the compounds described above continue to be detected in the environment and have been reported since the first report of global pollution by PFOS [13]. Owing to the complex relationship between precursors and transformed compounds and the limited information pertaining to these compounds, PFCs are still receiving great attention as important emerging contaminants (ECs). Studies of these compounds include identification of their sources and fates in the environment. 26.1.2 History, Regulation and Pollution of PFCs Since the invention of polytetrafluoroethylene (PTFE) in 1937, various PFCs have been developed and used since the 1950s as water repellents, upholstery, textiles, fire-fighting materials, hydraulic fluids, photo-resistant materials, emulsifying agents for PTFE production, and insecticides. Because polyfluorinated alkyl chains exhibit several advantages, including physical and chemical stability, oil and water repellency and solvent surface tension of materials, PFCs containing polyfluoroalkyl chains have been used in many of the aforementioned products. Although most PFCs have been used in ester forms or as building blocks of polymers, ionic or free hydrophilic groups, such as sulfonyl/carboxyl groups in PFOS/PFOA, also play a key role as surfactants in some applications, such as photo-resistant materials and emulsifying agents. After consultation between the USEPA and 3M, the phase-out of PFOS and its related chemicals production by 2003 was declared in 2000 [14]. At the same time, global PFOS pollution was reported in 2001 [13], and several regulations for the trade, production, use and/or emission of PFOS and PFOA were established. In addition, guideline values or regulations were established, such as in TSCA in 2003, provisional tolerable daily intake (pTDI) by the Committee on Toxicity in the UK in 2006, the EU Directive in 2008 and the Stockholm Convention on POPs in 2009 [15]. For the PFOA, additionally, a stewardship program designed to reduce the emission of PFOA and its homologs by the major producers of PFOA was developed and has been continued by the USEPA [16]. Most PFCs are ionic chemicals (e.g., PFOS, PFOA, and their different chain length compounds) that are water-soluble; thus, they are likely to be present in aquatic systems. Although the use, production and trade of PFOS and PFOSF are regulated under international convention and alternatives to these compounds are being investigated, several PFOSrelated chemicals are still used and emitted into the environment and detected in aquatic systems [17,18], beverages [19], wildlife [20], and humans [21]. Even PFOS and PFOA, which are strictly regulated, are still detected in running river water [22–24]. This is likely because potential social stocks and temporal environmental sinks of PFCs exist (e.g., stocks of fire-fighting foams in car parking, industrial stocks, commercial products, and compounds present in water pond, fields, roadside dust, ice, and snow). In addition to the stocks, their precursors may become sources of PFOS and PFOA in the environment. However, the mechanisms for continuing PFOS and PFOA emission and detection in the environment after

780 Transformation Products of Emerging Contaminants in the Environment

introduction of regulations remain a matter of debate and require further study. We consider the existence of nonpoint sources to be one of the most important causes of the detection of PFCs in rivers after implementation of the aforementioned regulations. Therefore, investigations of PFC pollution in aquatic environments with consideration of spatial factors are important in enabling a better understanding of the current situation and successful management of these compounds. In this chapter, PFC pollution in the water environment is discussed while focusing on spatially distributed sources. The developed method to accomplish this has been termed GIS-based source identification. This technique uses multiple geographic indices, such as population density and land-use type, for the identification and characterization of sources. Furthermore, a methodology for spatial prediction of PFC concentrations and their distribution in the watershed is developed. Advanced analysis for source, fate and risk identification and characterization with spatial information requires spatial modeling with finer resolution; therefore, a method for modeling the spatial distribution of PFCs with a 500-m grid using combined techniques of land-use regression (LUR) and the digital elevation model (DEM) was developed. Implementation of this spatial modeling technique based on inverse modeling (receptor modeling) enabled successful elucidation of the spatial distribution of PFC sources and their concentrations in the watershed with high resolution. Further analysis will be useful for detecting regions of concern to humans and wildlife, and for management of tangled PFC pollution in watersheds.

26.2 Source Identification of PFCs Using GIS 26.2.1 Study Area and Dataset The Tokyo Bay Basin was selected as a model area for development and application of the GIS-based source identification method. This basin is one of the most industrialized, urbanized, and populated areas in the world (Figure 26.1), and is therefore predicted to have higher levels of PFCs than other areas. In fact, PFOS, PFOA and perfluorononanoate (PFNA) have been detected in most of the 50 sampling sites located throughout the basin at median concentrations of 5.8, 6.7 and 20.1 ng/L, respectively [24]. This dataset of PFC concentrations in river water samples was used for the GIS-based analysis. The samples used for the analysis were collected from the downstream end of a river in each watershed (Figure 26.2). Because several of the 35 types of PFC measured were below the limit of quantification for most samples and several species were not easily applied to the analysis, the dataset consisted of perfluorohexanoate, perfluoroheptanoate (PFHpA), perfluorooctanoate (PFOA), PFNA, perfluorodecanoate, perfluoroundecanoate, perfluorohexane sulfonate, PFOS, the first fraction in chromatographic separation of the branched isomer of PFOS (PFOSisomer1), the second fraction of the branched isomer of PFOS (PFOSisomer2), FOSA and N-ethyl perfluorooctane sulfonamideacetate (NEtFOSAA). Details pertaining to the samples, including the criteria implemented to prevent bias from influencing the source identification, have been described elsewhere [25]. The quality of the water samples is considered to reflect the characteristics of their watershed, such as levels of industrialization and population levels. To identify the relationship between the quality of water and the characteristics of the watershed, a GIS dataset of the basin of Tokyo Bay was constructed. The prepared candidates of geographic information for PFC source identification were the airport area, number

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 781

Figure 26.1 Location of the Tokyo Bay Basin and spatial distribution of population density.

of waste disposal sites, road area, area of each land-use type (rice paddies, other agriculture, forests, wasteland, buildings, arterial traffic, rivers and lakes, seaside, coastal water, golf courses, and other areas), buffer areas around train stations, predicted boundaries of tidallyaffected areas, population, number of business establishments/employees, sewage treatment areas, and number of public and other facilities. The area or numbers of these geographic indices were calculated for each watershed in the basin. Geographic indices that met the following two criteria were used for source identification and apportionment. (1) The indices should be interpretable as a PFC source. (2) The correlation coefficient between indices should be below 0.7. This criterion was set to avoid multicollinearity in the source identification model. Based on these criteria, 11 geographic indices were selected as the explanatory variables for use in the source identification. These 11 candidates were extracted from the GIS database as density in the watershed. The candidates included the area excluding forest and wasteland, percentages of agricultural area, arterial traffic area, golf courses, river and lake areas, other land-use areas, buffer areas around train stations (50 m in radius), sewage treatment areas, sewage treatment plant catchment areas, number of waste disposal sites, and population in each watershed. In several previous reports, population density showed a positive relationship with PFC pollution; therefore, these factors were used as indicators of diffuse pollution of PFC [22,26,27]. In this study, we attempted to include this indicator to test whether the indicator of population alone or multi-GIS parameters could better explain the diffused PFC sources.

782 Transformation Products of Emerging Contaminants in the Environment

Figure 26.2 Sampling sites and their watersheds in the Tokyo Bay Basin.

26.2.2 Method of GIS-based Source Identification The GIS-based source identification was constructed as shown in Figure 26.3 using a method that has been described in detail elsewhere [25]. Briefly, the dataset of PFC concentration was prepared as objective variables, while geographical indices were prepared as explanatory variables in a multiple linear regression analysis. Generally, environmental variables, such as concentration of chemical species, have a log-normal distribution; thus, the dataset of PFC concentrations was converted to logarithmic values. Owing to the better relationship between log-transformed PFC concentrations and log-transformed geographic indices, multiple linear regression analyses were conducted following conversion of the indices to logarithmic values. The equation of regression analysis was as follows: PFC conc ¼

Yn j¼0

b

fj j

(14.1)

where fj is the score of source (factor) j, and bj is an estimated parameter in multiple linear regression analysis. In the case of j ¼ 0, fj ¼ e is accepted. The PFC loadings from the Tokyo Bay Basin into the bay were estimated by summing the amount of PFC in each basin. The source apportionment between point and nonpoint sources

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 783

Figure 26.3 Workflow of GIS-based source identification and source apportionment.

was conducted by subtraction of the above calculated concentration (loadings from nonpoint sources) from the measured PFC concentration (loadings from both point and nonpoint sources). 26.2.3 Results and Discussion 26.2.3.1 Spatial Trend of PFC Sources GIS-based source identifications by multiple regression analyses were conducted and the results for 12 PFCs are shown here. The selected variables, which were statistically significant as sources (factors), are shown for PFOA, PFNA and PFOS in Table 26.2. Areas excluding forest and wasteland were selected as the source for PFOA. In other words, PFOA pollution showed uniform contributions from artificial land-use. In addition, the catchment area of the sewage wastewater was also selected for PFNA. Thus, sewage treatment plants were indicated as contributors to PFNA. For PFOS, contributions from traffic-related contributions were implicated because the arterial traffic area was among the significant variables. Generally, PFOA is spatially widely detected in environmental samples when compared with PFOS. Taken together, our results directly indicated a wide distribution of perfluoroalkyl carboxylates (PFCAs), including PFOA. For PFOS, higher concentrations of PFOS were detected from road dust collected from heavily traveled areas than from residential areas, suggesting emanation of PFOS from road surfaces [28]. Although PFCAs, including PFOA, were also detected from the road dust at high concentrations, a high correlation between traffic activity and PFOS was still indicated based on their results. The variable “population” was selected as a secondary source of PFOS; thus, this variable might be a good indicator of nonpoint sources of PFOS, but it is not the best indicator. Overall, these results indicate the importance of using the GIS-based source identification and its implementation in the detailed evaluation of nonpoint sources of PFCs.

784 Transformation Products of Emerging Contaminants in the Environment Table 26.2 Regression equation obtained during GIS-based source identification. b1

b2

b3

b0

R2-ad

> LOQ (%) within the collected samples

Selected geographic indices Partial regression coefficient Standardized partial regression coefficient PFOA

PFNA

PFOS

Area excluding forest and wasteland 1.3 0.91 Area excluding forest and wasteland 1.4 0.77 Arterial traffic area 0.83 0.51

—

—

Constant

0.84

98

— — Catchment area of sewage wastewater 0.15 0.22 Population

— — —

2.2 — Constant

0.68

94

— — —

4.4 — Constant

0.84

94

0.42 0.46

— —

1.6 —

R2-ad: determination coefficient adjusted for the degrees of freedom.

The spatial distributions of analyzed PFC sources are depicted in Figure 26.4. The spatial trend of PFC sources (factors) is obvious in this visual expression. The sources (factors) of perfluoroalkyl sulfonate (PFSA) and its precursors were aggregated in Tokyo, while those of PFOA and PFHpA were more spatially uniform. The other PFCAs also tended to be spatially uniform when compared with PFSA and its precursors. An advantage of the GIS-based source identification is that the finding and scoring abilities of the spatially distributed sources (factors) are provided with visual expression. 26.2.3.2 Calculation of the Contribution of Point Sources and Nonpoint Sources Source apportionments between point and nonpoint sources were conducted for individual rivers flowing into Tokyo Bay (the Tama, Sumida, Ara, Naka, Edo and Tsurumi rivers). These six rivers are first-class rivers flowing into Tokyo Bay that comprise the bulk of the water flow into the bay. The loadings from nonpoint sources in each river were calculated according to the method described above, and the loadings from point sources were estimated as the difference between observed loads and loading from nonpoint sources. The results obtained for PFOS are shown in Figure 26.5. The PFOS contributions from nonpoint sources were comparable to those of the point sources in the Naka, Edo and Tsurumi rivers. However, the contributions from point sources were five to seven times higher than those from nonpoint sources in the Tama, Sumida and Ara rivers). During estimation of the loadings from point sources, only one grab sample from each river was used. Thus, the error bar for the point source loading (Figure 26.5) does not include the fluctuation of concentration in the rivers. Moreover, the loadings of nonpoint sources do not include the effects of rain

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 785

Figure 26.4 Spatial distribution of nonpoint sources of PFCs in the Tokyo Bay Basin.

Figure 26.5 Source apportionments of PFOS between point and nonpoint sources in each river flowing into Tokyo Bay.

786 Transformation Products of Emerging Contaminants in the Environment

runoff, which increase the PFC loadings of nonpoint sources by flushing out contaminants present on road surfaces [29]. Although accurate calculation of the impact of nonpoint sources is difficult using GIS-based source identification alone, this method provides important information pertaining to PFC loadings from nonpoint sources in the steady state while assigning sources contributing to nonpoint pollution.

26.3 Spatial Distribution of PFOS and PFOA Contributed by Nonpoint Sources 26.3.1 Method for Spatial Prediction with Fine Scale; Use of the Digital Elevation Model (DEM) and Land-Use Regression (LUR) Model In the previous section, nonpoint source maps were created for individual PFCs after estimating the spatially distributed sources (factors) based on multiple linear regression analysis. Through this analysis, the model equation and its parameters were obtained. The spatial distribution of sources (factors) can be predicted using this obtained model, which is called the LUR model. In this section, a method for increasing the resolution of the source map is introduced, and a pollution map is then constructed using the LUR model and DEM while maintaining its high resolution. The entire workflow for generating the pollution map is shown in Figure 26.6, which also includes the workflow of the previous section (Section 26.2). After deducing the model equation, the high resolution (finer than 500-m grid) GIS data for the indices selected for the regression equations were obtained. The source maps were then reconstructed using the

Figure 26.6 Entire workflow for generation of the pollution map.

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 787

Figure 26.7 Algorithm for water flow simulation based on DEM.

obtained high resolution data. In the following step, the amount of PFCs in each river was computed from the river water flow simulation based on DEM (Figure 26.7). In the first stage of the water flow simulation, the direction of water flow was determined based on the differences in height calculated from the DEM data, after which a code number corresponding to the direction was assigned for each grid cell. The water flow was then simulated based on the direction data for the grid cells and subsequently used to model the total river flow. This simulation was conducted using the ArcGIS 10.1 software (Esri Inc., NY, USA), which included the function library to compensate for the local concavity of the DEM data. The algorithm used to compute a map of PFC pollution after simulating a river is shown in Figure 26.8. At first, the amount of PFC emission from the source map was applied according to the water flow. This amount, which is referred to as PFC loading, was then divided by the amount of water. This procedure was used to determine the concentration of PFC in each cell, and the map of PFC pollution was produced. These computations were conducted using programing software R-2.10.1 [30], and the pollution maps were prepared only for PFOA and PFOS because corresponding high resolution GIS data for the LUR model could only be obtained for those PFCs. 26.3.2 Results and Discussion 26.3.2.1 Simulation Results of River Water Flow in the Tokyo Bay Basin Figure 26.9 shows the simulated river water flow in the Tokyo Bay Basin obtained using DEM data with a 250-m grid. The gray line in the map shows the actual river, and the bold

788 Transformation Products of Emerging Contaminants in the Environment

Figure 26.8 Computation of a map of PFC pollution in combination with DEM and the LUR model.

Figure 26.9 Simulated water flow based on DEM in the Tokyo Bay Basin. The gray line in the map shows the actual river, and the bold black line shows the simulated results based on DEM.

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 789

black line shows the simulated results based on DEM. Except for the upper-mid right portion of the map and lakes, the simulation based on DEM successfully reconstructed the water flow in Tokyo Bay Basin. After this simulation, every four adjacent cells of the 250-m grid cell were combined and converted to 500-m grid cells. This procedure was conducted to adjust the data resolution of several GIS datasets. 26.3.2.2 Results of Spatial prediction of PFOS and PFOA Initially, enhancement of the resolution of PFC source maps was conducted from the watershed-base to a 500-m grid. Generally, the areas of the watershed in this study were approximately 10–500 km2; therefore, enhancement of the map resolution to a 500-m grid (0.25 km2) is expected to improve the expression of the results and make the analysis easier and more precise. Figure 26.10 shows the improvement in the expression of the PFOS source map in response to enhancement of the map resolution. In the source map with the watershed-base, the extent of the value differences for location was small because the potential PFOS sources were averaged within each watershed. Conversely, the source map with the 500-m grid clearly and precisely showed the distribution of PFOS sources with higher variance among values in each location. For example, the map with the 500-m grid indicates a high potential for PFOS pollution near Tokyo. This change in results between the maps with different resolutions is known as the modifiable areal unit problem, in which higher resolution maps provide more precise results for spatial prediction of pollution. In the second step, maps of PFC pollution were prepared from the source map created by the LUR model amended with the water flow algorithm based on DEM. The source map and the pollution map of PFOA are provided for comparison in Figure 26.11. There were values missing from the source maps owing to the absence of values in the original GIS data. These were compensated for during the water flow simulation, but calculated as 0 in this analysis.

Figure 26.10 Comparison of PFOS source map with watershed-base (a) and 500-m grid (b).

790 Transformation Products of Emerging Contaminants in the Environment

Figure 26.11 PFOA source map (a) and pollution map generated from the LUR and DEM (b).

The number of these points was limited in this study and, therefore, they have a negligible impact on the overall results. Although several slight changes are difficult to see in this figure, the lines of the major rivers are represented in the center of the pollution map. This reflects the dilution of PFOA in locations downstream of high concentration areas in response to input of upstream river water with low concentrations of PFOA. The results of predicted and measured values of PFOS and PFOA are compared in Figure 26.12. The predicted values were obtained from points that correspond to the downstream end of the watershed. For detailed evaluation of the model, measured samples not used in the model construction should be prepared and compared with the predicted values, or n-fold cross validation should be undertaken. In the current evaluation, predicted and measured values were within one order of magnitude of each other for most samples. These findings indicate that the LUR model in combination with DEM can predict the PFC pollution in local levels with sufficient accuracy. This methodology has been shown to be useful for identification of pollutant sources and prediction of their spatial influence. Naturally, the environmental risk can be estimated from spatial information after successfully predicting the environmental fate of the compounds. Figure 26.13 shows the risk quotients of PFOS as the predicted no-effect environmental concentration value to be 50 ng/ L [31]. The areas of high concern could be identified from this risk mapping. In this case, areas of concern were near Tokyo and several metropolitan areas. 26.3.2.3 Further Development for Spatial Prediction and Evaluation of ECs Several challenges remain prior to widespread implementation of the spatial methodology for the management of ECs, including determining how to deal with null values and evaluate models. Although highly water-soluble persistent compounds, such as PFOS and PFOA, were predicted well by the model, several functions must be added to enable the developed

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 791

Figure 26.12 Predicted concentrations of (a) PFOS and (b) PFOA and comparison with measured concentrations.

model to deal with degradable compounds. Addition of the function of degradation is not difficult because the degradation rate will not vary with location; however, the transportation mechanism differs greatly between water-soluble compounds and hydrophobic compounds. The amount of particle solid and extent of organic carbon in the particle will affect the results. These types of parameters may vary spatially, which would necessitate use of an advanced spatial model or additional GIS data. Considering the runoff coefficient for each land-use type may compensate for the different transportation mechanisms of hydrophobic compounds. For example, organic-rich particles are abundant in urbanized areas; thus, the runoff coefficients of the target hydrophobic compounds in such areas must be adjusted to correct for differences in transportation mechanisms. A calibration method for detecting optimized runoff coefficients using PFOS data was demonstrated. First, the areas on the grid were categorized as urbanized or natural, based on the ratio of land-use type. Runoff

792 Transformation Products of Emerging Contaminants in the Environment

Figure 26.13 Risk quotients of PFOS in Tokyo Bay Basin (predicted no-effect environmental concentration ¼ 50 ng/L).

coefficients (rate of runoff) were then assigned for each grid based on the grid type, after which the PFOS concentration was predicted under these conditions. The root mean square error between the predicted values and the measured values was used as an indicator for the optimization. The rate of runoff was altered by 0.1 for both urbanized area and natural area, after which a total of 100 patterns were calculated, as shown in Figure 26.14. In this case, model evaluation was conducted by the measured samples used in construction of the model; thus, no change of setting (rate of runoff ¼ 1) was optimal. Further development of the spatial prediction in watersheds can be addressed by these types of approaches.

26.4 Conclusion A GIS-based methodology for elucidation of the sources and fate of PFCs was introduced. This methodology utilized a type of inverse model (receptor model) that employs environmental data and is, therefore, especially effective for modeling new ECs with limited information. Applying this GIS-based approach to the Tokyo Bay Basin exhibited its effectiveness for elucidation of nonpoint sources of PFCs and their fate. PFOS and its

Spatial Modeling for Elucidation of Perfluorinated Compound Sources 793

Figure 26.14 Calibration of rate of runoff for the identification of optimized settings. RMSE: Root mean square error.

relatives were found to have densely concentrated sources in traffic areas, while uniformly distributed sources of PFCAs, especially PFOA and PFHpA, were projected. These results support the findings and data of several other studies. Prediction of the spatial aspect of pollutant sources is one of the advantages of using GIS-based source identification. In this study, elucidation of PFOS and PFOA fate in a watershed were addressed using the LUR model constructed from the GIS-based source identification and DEM. This approach successfully predicted the fate of PFOS and PFOA with a high resolution spatial map. Further development of this approach should lead to estimation of environmental risks associated with pollutants and risk identification of specific areas of concern via precise risk mapping. Although several challenges remain to development and implementation of this method for actual chemical management, especially for ECs, application of this GIS-based approach to watersheds will likely become one of the key approaches to mitigation of pollution by ECs and other contaminants in such areas.

Acknowledgments This study was supported by JSPS Research Fellowships for Young Scientists (Grant No. 249089) and in part by the River Fund (No. 231211012) in charge of the Foundation of River and Watershed Environment Management (FOREM), Japan.

References 1. UNEP (2009) The 9 new POPs under the Stockholm Convention, http://chm.pops.int/ Programmes/NewPOPs/The9newPOPs/tabid/672/language/en-US/Default.aspx. 2. Benskin, J.P., Ikonomou, M.G., Gobas, F.A.P.C. et al. (2012) Observation of a novel PFOSprecursor, the perfluorooctane sulfonamido ethanol-based phosphate (SAmPAP) diester, in marine sediments. Environmental Science & Technology, 46(12), 6505–6514.

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27 Global Distribution of Polyfluoroalkyl and Perfluoroalkyl Substances and their Transformation Products in Environmental Solids Holly Lee and Scott A. Mabury Department of Chemistry, University of Toronto, Canada

27.1 Introduction Perfluoroalkyl and polyfluoroalkyl substances (PFASs) are anthropogenic chemicals that have a fluoroalkyl backbone and a polar headgroup, both of which impart high surface activity and the ability to repel water, oil, and stain to these chemicals [1]. As such, PFASs are crucial components in non-stick, greaseproofing, and surface treatment applications. Commercial fluorochemical production has largely proceeded by two manufacturing processes, electrochemical fluorination (ECF) and telomerization [1], with the bulk of the production centered on high molecular weight (MW) fluorinated polymers and surfactants [2,3] and a minor proportion directed towards the synthesis of specific perfluoroalkane sulfonate (PFSA) and perfluoroalkyl carboxylate (PFCA) congeners. Perfluorooctane sulfonate (PFOS, C8) was the only PFSA deliberately produced to be used in AFFFs and various performance applications [4,5] until it was phased out of production in 2000–2002 [6]. Among the PFCAs, perfluorooctanoate (PFOA, C8) and perfluorononanoate (PFNA, C9) are primarily used as processing aids in the manufacture of fluoropolymers [7], although other PFCAs of varying chain length have been detected as residual impurities in commercial products [8]. Since the first discovery of PFOA and PFOS in human blood [9] and wildlife [10], these perfluoroalkyl acids (PFAAs) and other PFASs (Table 27.1) have emerged as common contaminants in surface water [11], sediments [12], WWTP sludge [12], and soil [13]. Detection Transformation Products of Emerging Contaminants in the Environment: Analysis, Processes, Occurrence, Effects and Risks, First Edition. Edited by Dimitra A. Lambropoulou and Leo M. L. Nollet. # 2014 John Wiley & Sons, Ltd. Published 2014 by John Wiley & Sons, Ltd.

798 Transformation Products of Emerging Contaminants in the Environment Table 27.1 Names, acronyms, and structures of various PFASs of interest. Name

Acronym

Fluorotelomer-based Substances x:2 Fluorotelomer alcohol x:2 FTOH x:2 Fluorotelomer acrylate x:2 FTAC x:2 Polyfluoroalkyl phosphate monoester x:2 Polyfluoroalkyl phosphate diester x:2 Polyfluoroalkyl phosphate triester x:2 Fluorotelomer sulfonate Semifluorinated x-alkane Semifluorinated alkene x:2 Fluorotelomer carboxylate x:2 Fluorotelomer unsaturated carboxylate

x:2 monoPAP x:2 diPAP x:2 triPAP x:2 FTSA SFA or FxHy SFAene or FxHyene x:2 FTCA x:2 FTUCA

Perfluoroalkane Sulfonamido-based Substances Perfluorooctane sulfonamide FOSA N-Methyl perfluorooctane MeFOSA sulfonamide N-Ethyl perfluorooctane EtFOSA sulfonamide Perfluorooctane FOSAA sulfonamidoacetate N-Methyl perfluorooctane MeFOSE sulfonamidoethanol N-Ethyl perfluorooctane EtFOSE sulfonamidoethanol N-Methyl perfluorooctane MeFOSAA sulfonamidoacetate N-Ethyl perfluorooctane EtFOSAA sulfonamidoacetate N-Ethyl perfluorooctane SAmPAP sulfonamidoethyl phosphate diester Perfluoroalkyl Acids (PFAAs) Perfluoroalkyl carboxylate Perfluoroalkane sulfonate Perfluoroalkyl phosphonate Perfluoroalkyl phosphinate

PFCA PFSA Cx PFPA Cx/Cy PFPiA

Structure F(CF2)xCH2CH2OH, x ¼ 4, 6, 8, 10, . . . F(CF2)xCH2CH2OC(O)CH ¼ CH2, x ¼ 4, 6, 8, 10, . . . F(CF2)xCH2CH2OP(O)O2, x ¼ 4, 6, 8, 10, ... [F(CF2)xCH2CH2O]2P(O)O, x ¼ 4, 6, 8, 10, ... [F(CF2)xCH2CH2O]3P(O), x ¼ 4, 6, 8, 10, . . . F(CF2)xCH2CH2SO3, x ¼ 4, 6, 8, 10, . . . F(CF2)x(CH2)yH, x ¼ 3–20, y ¼ 3–20 F(CF2)xCH ¼ CH(CH2)y, x ¼ 3–20, y ¼ 3–20 F(CF2)xCH2CO2, x ¼ 4, 6, 8, 10, . . . F(CF2)x-1CF ¼ CHCO2, x ¼ 4, 6, 8, 10, . . .

F(CF2)8SO2NH2 F(CF2)8SO2NH(CH3) F(CF2)8SO2NH(CH2CH2) F(CF2)8SO2NH(CH2C(O)O) F(CF2)4SO2N(CH3)CH2CH2OH F(CF2)4SO2N(CH2CH3)CH2CH2OH F(CF2)8SO2N(CH3)(CH2C(O)O) F(CF2)8SO2N(CH2CH3)(CH2C(O)O) [F(CF2)8SO2N(CH2CH3)(CH2CH2O)]2P(O)O-

F(CF2)xCO2, x ¼ 1–13 F(CF2)xSO3, x ¼ 4, 6, 8, 10 F(CF2)xP(O)O2, x ¼ 6, 8, 10 F(CF2)xP(O)(O)((CF2)yF), x ¼ 6, 8; y ¼ 6, 8, 10, 12; x þ y  18

Global Distribution of Polyfluoroalkyl and Perfluoroalkyl Substances 799

of PFASs has been reported worldwide and even in remote environments, like the Arctic and Antarctica. PFASs may be directly released into the environment via emissions of contaminated discharges from fluorochemical manufacturers and the disposal of commercial products in which these chemicals are either present as active ingredients or as residual impurities. Environmental degradation of commercial fluorinated polymers and surfactants present in disposed products to the PFSAs and PFCAs also represents an indirect source of these PFAAs to the environment. In addition, commercial manufacture of fluorinated chemicals is typically a crude process during which unreacted starting materials or byproducts may be incorporated into the consumer products [4,7]. In fact, analysis of various commercial fluorinated products revealed the presence of fluorotelomer alcohols (FTOHs) and N-methyl perfluorooctane sulfonamidoethanol (MeFOSE) as residual impurities in amounts up to 4% [14]. As FTOHs and MeFOSE are volatile, the release of these and potentially other volatile materials via offgassing from commercial products may represent a significant source to the atmospheric fluorochemical burden. Atmospheric transport and degradation of volatile fluorinated precursors to PFCAs [15,16] and PFSAs [17,18], and the subsequent deposition of these transformation products (TPs) may in part contribute to the background levels of PFAAs observed in the environment (Figure 27.1). Results from modeling the formation of PFOA from the atmospheric oxidation of 8 : 2 FTOH predicted ubiquitous PFOA pollution in the Northern hemisphere atmosphere, with higher concentrations occurring in remote regions (e.g., Arctic) than in source regions [19]. This distribution is consistent with the high nitrogen oxide (NOx) environment of urban locations in which NOx may interfere with the atmospheric formation of PFCAs and thus reduce their yields. As such, point sources may be more important contributors to local PFAS contamination in near-source regions. Upon exiting a WWTP, domestic and industrial effluents may further contaminate receiving water bodies, as have been documented by the detection of PFASs in WWTP samples, surface water, and sediments collected downstream from these facilities (Figure 27.1). However, these data do not account for the potential release of PFASs from the treated sludge or biosolids co-generated at these WWTPs. The disposal of these solid waste materials may have implications for human and wildlife exposure, especially if they are applied as a soil fertilizer onto agricultural fields (Figure 27.1). PFASs have a demonstrated capacity to sorb to environmental solid matrices, such as clay minerals [20–22], sediments [23–27], soils [28,29], and sludge [30]. The fact that PFASs may sorb to these environmental solids has implications for the long-term retention and release of these chemicals to the aqueous environment. A number of studies have attributed land application of contaminated WWTP sludge [31,32], AFFF use at fire-training facilities [33,34], and leaching of urban wastewater and runoffs [35,36] as potential sources of groundwater and surface water PFAS contamination. The primary concern of this contamination, particularly in the groundwater, centers over its potential as a route of human exposure to PFASs through drinking water. Human and wildlife exposure may also occur through ingestion of contaminated field crops, as evidenced by recent experimental and field data that demonstrated the transfer of PFCAs and PFSAs from contaminated soils to assorted plants [37–39]. This chapter summarizes the monitoring data collected from sediment, WWTP sludge, and soil samples collected around the world in the context of these environmental pathways. Source elucidation of PFASs will also be evaluated by identifying diffuse sources (i.e., atmospheric transport, urban/industrial discharges) as contributors to ambient levels and

Global Distribution of Polyfluoroalkyl and Perfluoroalkyl Substances 801

distinguishing them from known fluorochemical point sources. A discussion of various processes known to control the distribution of PFASs in the environment is also presented.

27.2 Global Contamination of PFASs in Environmental Solid Matrices 27.2.1 Sediments The detection of PFCAs and PFSAs of varying chain lengths has been widely reported in freshwater, coastal, and marine sediments collected around the world (Table 27.2). The data presented are considered to be background levels in this review and representative of the diffuse sources in the environment. Total PFCA (SPFCA) and PFSA (SPFSA) concentrations range from low (