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Edited by FREDERICK S.B. KIBENGE Department of Pathology and Microbiology, Atlantic Veterinary College, University of Prince Edward Island, Charlottetown, PE, Canada BERNARDO BALDISSEROTTO Full Professor, Department of Physiology and Pharmacology, Federal University of Santa Maria, Santa Maria, Rio Grande do Sul, Brazil ROGER SIE-MAEN CHONG Registered Veterinary Specialist of Fish Health and Production, Royal College of Veterinary Surgeons, London, United Kingdom Registered Specialist of Veterinary Aquatic Animal Health, Queensland Veterinary Surgeons Board; Senior Veterinarian (Aquatic Pathology) Biosecurity Sciences Laboratory, QLD, Australia Current affiliation: Australian Commonwealth Scientific Industrial Research Organization (CSIRO), Brisbane, QLD, Australia Academic Press is an imprint of Elsevier 125 London Wall, London EC2Y 5AS, United Kingdom 525 B Street, Suite 1650, San Diego, CA 92101, United States 50 Hampshire Street, 5th Floor, Cambridge, MA 02139, United States The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, United Kingdom © 2021 Elsevier Inc. All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, recording, or any information storage and retrieval system, without permission in writing from the publisher. Details on how to seek permission, further information about the Publisher’s permissions policies and our arrangements with organizations such as the Copyright Clearance Center and the Copyright Licensing Agency, can be found at our website: www.elsevier.com/permissions. This book and the individual contributions contained in it are protected under copyright by the Publisher (other than as may be noted herein).

Notices Knowledge and best practice in this field are constantly changing. As new research and experience broaden our understanding, changes in research methods, professional practices, or medical treatment may become necessary. Practitioners and researchers must always rely on their own experience and knowledge in evaluating and using any information, methods, compounds, or experiments described herein. In using such information or methods they should be mindful of their own safety and the safety of others, including parties for whom they have a professional responsibility. To the fullest extent of the law, neither the Publisher nor the authors, contributors, or editors, assume any liability for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions, or ideas contained in the material herein. Library of Congress Cataloging-in-Publication Data A catalog record for this book is available from the Library of Congress

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Typeset by SPi Global, India Contents

Contributors ix About the editors xi Preface xiii

1. Introduction to aquaculture 1 Frederick S.B. Kibenge

1.1 Introduction 1 1.2 Structure of the global aquaculture industry 2 1.3 Mollusk aquaculture 9 1.4 Crustacean aquaculture 10 1.5 Chemicals in aquaculture 10 1.6 Governance of aquaculture 11 References 12

2. General introduction to toxicology of aquatic animals 17 Bernardo Baldisserotto

2.1 Introduction to toxicology 17 2.2 Water quality criteria/guidelines 18 2.3 Intraspecies variation of toxicity 20 2.4 Models to predict toxicity of contaminants 21 References 22

3. Antifoulants and disinfectants 25 Samantha Eslava Martins and Camila de Martinez Gaspar Martins

3.1 Overview 25 3.2 Definitions and uses 26 3.3 Mode of action 28 3.4 Ecotoxicity and biological effects 33 3.5 Ecological risks and regulation 46 3.6 Further considerations 49 References 50

v vi Contents

4. Metals 59 Claudia B.R. Martinez, Juliana D. Simonato Rocha, and Paulo Cesar Meletti

4.1 Introduction 59 4.2 Biochemical effects 62 4.3 Physiological effects 67 4.4 Behavioral effects 69 References 73

5. Agrochemicals: Ecotoxicology and management in aquaculture 79 Vania Lucia Loro and Bárbara Estevão Clasen

5.1 Water and soil contamination by agrochemicals 79 5.2 Environmental contamination by agrochemicals and risk assessment in aquaculture: Effects on aquatic organisms and food for human consumption 82 5.3 Mitigation of agrochemicals 94 5.4 Agrochemicals banned from use in agriculture and aquaculture 98 5.5 Regulatory process for new chemicals and good agricultural practices 98 References 101

6. Pharmaceutical pollutants 107 Helena Cristina Silva de Assis

6.1 Introduction 107 6.2 Pharmaceuticals in the aquatic environment 108 6.3 Pharmaceutical sources and pathway to the environment 109 6.4 Pharmaceutical exposure effects in nontarget species 111 6.5 Final considerations 123 References 124

7. Oil and derivatives 133 Helen Sadauskas-Henrique, Luciana Rodrigues Souza-Bastos, and Grazyelle Sebrenski Silva

7.1 Oil and derivatives and the aquatic contamination 133 7.2 Aquaculture and the problem of oil and derivative contamination 137 Contents vii

7.3 Effects of oil and derivatives on fish species 140 7.4 Effects of oil and derivatives on mollusks and crustaceans 160 7.5 Interaction of oil and derivatives with water characteristics 169 7.6 Future perspectives on oil and derivative contamination and aquaculture 172 References 173

8. Ecotoxicological effects of microplastics and associated pollutants 189 Fábio Vieira de Araújo, Rebeca Oliveira Castro, Melanie Lopes da Silva, and Mariana Muniz Silva

8.1 Introduction 189 8.2 Impacts of microplastic on marine animals 190 8.3 Plastic additives 191 8.4 Microplastic and persistent organic pollutants (POPs) 195 8.5 Microplastics and metals 203 8.6 Microplastics and microorganisms: The plastisphere 205 8.7 Microplastics and other compounds 213 8.8 Final considerations 214 References 216

Index 229 This page intentionally left blank Contributors

Bernardo Baldisserotto Department of Physiology and Pharmacology, Universidade Federal de Santa Maria, Santa Maria, RS, Brazil Ba´rbara Estevaõ Clasen Department of Environmental Sciences, State University of Rio Grande do Sul, Três Passos, RS, Brazil Rebeca Oliveira Castro Programa de Po´s Graduaçaõ em Biologia Marinha e Ambientes Costeiros, Universidade Federal Fluminense, Niteroí, RJ, Brazil Camila de Martinez Gaspar Martins Universidade Federal do Rio Grande—FURG, Instituto de Ciências Biolo´gicas, Rio Grande/RS, Brazil Fa´bio Vieira de Arau´jo Faculdade de Formaçaõ de Professores, Universidade do Estado do Rio de Janeiro, Saõ Gonçalo, RJ, Brazil Grazyelle Sebrenski Silva Departamento de Morfologia, Universidade Federal do Amazonas (UFAM), Manaus, Brazil Luciana Rodrigues Souza-Bastos Instituto de Tecnologia para o Desenvolvimento (LACTEC), Laborato´rio de Toxicologia e Avaliaçaõ Ambiental, Ambiental, Curitiba, Brazil Mariana Muniz Silva Programa de Po´s Graduaçaõ em Biologia Marinha e Ambientes Costeiros, Universidade Federal Fluminense, Niteroí, RJ, Brazil Melanie Lopes da Silva Laborato´rio de Ecologia e Dinâmica Bêntica Marinha, Faculdade de Formaçaõ de Professores, Universidade do Estado do Rio de Janeiro, Saõ Gonçalo, RJ, Brazil Frederick S.B. Kibenge Department of Pathology and Microbiology, Atlantic Veterinary College, University of Prince Edward Island, Charlottetown, PE, Canada Vania Lucia Loro Laboratory of Aquatic Toxicology, Department of Biochemistry and Molecular Biology, Federal University of Santa Maria (UFSM), Santa Maria, RS, Brazil Paulo Cesar Meletti Laboratory of Animal Ecophysiology, Department of Physiological Sciences, State University of Londrina, Londrina, Parana, Brazil Claudia B.R. Martinez Laboratory of Animal Ecophysiology, Department of Physiological Sciences, State University of Londrina, Londrina, Parana, Brazil Samantha Eslava Martins Norwegian Institute for Water Research (NIVA), Ecotoxicology and Risk Assessment Section, Oslo, Norway; Universidade Federal do Rio Grande—FURG, Instituto de Ciências Biolo´gicas, Rio Grande/RS, Brazil Helena Cristina Silva de Assis Department of Pharmacology, Federal University of Parana´, Curitiba, Parana´, Brazil

ix x Contributors

Helen Sadauskas-Henrique Universidade Santa Cecı´lia (Unisanta), Laborato´rio de Biologia de Organismos Marinhos e Costeiros (LABOMAC), Santos, Brazil Juliana D. Simonato Rocha Laboratory of Animal Ecophysiology, Department of Physiological Sciences, State University of Londrina, Londrina, Parana, Brazil About the editors

Dr. Frederick S.B. Kibenge is Professor of Virology at the Atlantic Vet- erinary College, University of Prince Edward Island, Charlottetown, PEI, Canada, where he has been Chairman of the Department of Pathology and Microbiology for several years, and teaches veterinary virology in the second year of the DVM curriculum. He has been working with animal viruses for more than 30years in addition to prior extensive postdoc- toral research experience in virology in the United Kingdom and the United States. He is a Diplomate of the American College of Veterinary Microbiol- ogists, ACVM (subspecialty Immunology). He has published extensively on the detection and virology of fish viruses. He is editor of the accompanying two textbooks on the theme of aquaculture: Aquaculture Pathophysiology and Aquaculture Pharmacology published by Elsevier Inc., Academic Press. Dr. Bernardo Baldisserotto is a full Professor of Physiology of the Departamento de Fisiologia e Farmacologia at the Universidade Federal de Santa Maria, in South Brazil. He supervises graduate students from the programs of Animal Husbandry, Pharmacology, and Animal Biodiversity at this university and has published several articles dealing with toxicology of aquatic animals, mainly fish. He is the coauthor on some of the book chapters and has published a book regarding Pharmacology applied to aquaculture (in Portuguese). He is editor of the accompanying two textbooks on the theme of aquaculture: Aquaculture Pathophysiology and Aquaculture Phar- macology published by Elsevier Inc., Academic Press. Dr. Roger Sie-Maen Chong is a veterinary specialist in Australia and the United Kingdom, with expertise in fish and shellfish pathology as applied to the health and biosecurity of aquacultured species. He is officially registered as a specialist by the Queensland Board of Veterinary Surgeons for Veterinary Aquatic Animal Health (Australia) and by the Royal College of Veterinary Surgeons for Fish Health and Production (United Kingdom). He is also a certified fish pathologist recognized by the Fish Health Section of the American Fisheries Society. He has worked in Hong Kong with the Department of Agriculture, Fisheries & Conservation, in Queensland with the Biosecurity Queensland and is presently a research fish pathologist with the Australian Commonwealth Scientific and Industrial Research Organization (CSIRO). He is editor of the accompanying two textbooks on the theme of aquaculture: Aquaculture Pathophysiology and Aquaculture Pharmacology published by Elsevier Inc., Academic Press.

xi This page intentionally left blank Preface

The textbook Aquaculture Toxicology focuses on the practical principles of toxicology of finfish (mainly), crustaceans, and mollusks, the three major groups of aquatic animal species cultured for human consumption. Recent books on this subject have a more general approach on aquatic toxicology, focusing more in chemical aspects of toxicology and pollutants (Introduction to Aquatic Toxicology by Nikinmaa, M., Elsevier, 2014; Microscale Testing in Aquatic Toxicology by Wells et al., CRC Press, 2018) or the specific aspects of the effects of pollutants (Aquatic Toxicology: Molecular, Biochemical, and Cellular Perspectives by Malins D.C. and Ostrander, G.K., 2018). Aquaculture is a rapidly growing agricultural sector and requires updated research information on general aspects of several types of toxicants to control or avoid their negative impacts on the culture of aquatic animals. The book is multiauthored and edited. All contributing authors for the different chapters are highly respected international experts recognized in their specific areas, who have published several articles dealing with the subject of their respective chapters; some have also written book chapters on the subject. Each book chapter has been written from the perspective of the expert, providing the reader with nuances of the topic without compromis- ing on the essential and useful facts or being encyclopedic but reflecting global research. Aquaculture Toxicology presents an overview of the practical information on the effects and sources of most common pollutants: disinfectants and antifoulants, metals, agrochemicals, pharmaceuticals, oils and derivatives, and microplastics. The authors were encouraged to use the same format in all chapters to facilitate ease of reading and studying, and to maintain consistency throughout the book although the ultimate approach chosen by different authors is reflective of the state of knowledge on the respective topics. A list of subtitles is included at the beginning of each chapter to provide the reader with a quick summary of what is covered in each chapter. This book provides the target audience with a reliable and valuable up- to-date, “all-inclusive” reference and guide to the state of the field, serving as a systematic and concise resource for aquaculture specialists/researchers interested in toxicity of various agents to fish, crustaceans, and mollusks, and particularly researchers, clinical veterinarians, and possibly government

xiii xiv Preface personnel interested in aquaculture, the fisheries and comparative biology, and how treatments may affect the aquatic animals directly. It is hoped that having such a comprehensive easily readable book focused on aquaculture will identify areas where more research is needed to generate more knowledge to support a sustainable aquaculture industry. The editors acknowledge the authors for their contributions.

Frederick S.B. Kibenge Bernardo Baldisserotto Roger Sie-Maen Chong EDITORS CHAPTER ONE

Introduction to aquaculture

Frederick S.B. Kibenge Department of Pathology and Microbiology, Atlantic Veterinary College, University of Prince Edward Island, Charlottetown, PE, Canada

1.1 Introduction The Food and Agricultural Organization of the United Nations (FAO) attributes aquatic organisms that are harvested by an individual or corporate body which has owned them throughout their rearing period to aquaculture. In contrast, aquatic organisms that are exploitable by the public as a common property resource, with or without appropriate licenses, are the harvest of fisheries (FAO, 2015). Two critical criteria of human activity, ownership of stock and deliberate intervention in the production cycle, distinguish aquaculture from capture fisheries and also account for the environmental concern centered on water quality/water pollution that consti- tutes Aquaculture Toxicology. It has been noted that the environmental impacts of aquaculture vary with species, system, management, production intensity, location, and environmental carrying capacity to absorb impacts (FAO, 2016). For example, the intensive rearing of fish has led to an increase in water pollution per unit of farmed fish produced (Hall et al., 2011). On the contrary, the filter feeding finfish typically raised in inland multispecies aquaculture systems (mostly silver carp Hypophthalmichthys molitrix and bighead carp H. nobilis, and most recently Mississippi paddlefish Polyodon spathula) and marine bivalve mollusks (oysters, mussels, clams, and scallops) raised in seas, lagoons, and coastal ponds require no artificial feeding (FAO, 2018) and improve the water quality in the production system by removing waste materials, including waste from fed aquatic animal species, and lowering the nutrient load (FAO, 2018; Hishamunda et al., 2014; Parisi et al., 2012). Deep-sea aquaculture is associated with a lower environmental impact on the seafloor (Welch et al., 2019) because of deeper waters and stronger ocean currents that disperse organic matter from farms, in contrast to nearshore farming. This introductory chapter describes the structure of the global aquaculture industry (finfish, mollusks, crustaceans, in cold water and warm water,

Aquaculture Toxicology © 2021 Elsevier Inc. 1 https://doi.org/10.1016/B978-0-12-821337-7.00007-4 All rights reserved. 2 Frederick S.B. Kibenge commercial and noncommercial, and small and big operations) and its importance in supplying animal protein food to the growing human population. A better understanding of the aquaculture methods will allow the reader to recognize the full implications of the toxic effects associated with different aquaculture environments and help improve toxicological risk management for aquaculture. The chapters that follow provide an overview of practical information covering five main aspects of Aquaculture Toxicology in the three main aquaculture groups (fish, crustaceans, and mollusks): (1) water quality criteria/guidelines used internationally and models used to predict the toxicity of contaminants; (2) the primary disinfectants and organic antifouling co-biocides currently used in aquaculture, pharmaceuticals used in human and veterinary medicine and their by-products that enter the aquatic environment; (3) water pollutants such as metals, oils and their derivatives, and microplastics that are toxic to aquatic animals but arising from nonaquaculture sources; (4) mechanisms of toxicity and the responses to toxic agents including aspects of uptake, metabolism, and excretion of toxicants in fish, crustaceans, and mollusks, and toxic effects on nontarget aquatic organisms; and (5) brief overviews of the current regulatory aspects of the use of these products.

1.2 Structure of the global aquaculture industry The contribution of wild-catch fisheries as an animal protein source for human consumption plateaued in the mid-1980s, and since then, most natural stocks in marine waters have been harvested at, or near-maximum rates, with the quantity of fish produced by capture fisheries, projected to increase at only 0.2% per year for the period 2019–2028 (OECD/FAO, 2019). In contrast, the average growth in the aquaculture sector in the same period is projected at 2.0% per year (OECD/FAO, 2019). Since 2015, aquaculture has overtaken capture fisheries as the source of seafood for human consumption, as global fisheries have remained flat and will continue to be flat for the foreseeable future (OECD/FAO, 2019). This share of aquaculture is projected to rise to 62% by 2030 as catches from wild capture fisheries level off (FAO, 2014; World Bank, 2013). Fish, mollusks, and crustaceans represent the most economically important global aquaculture industry subsectors (Fig. 1.1), at 80.04 million tonnes in 2016 with an estimated total farm gate value of USD 154.2 billion (FAO, 2018), and expected to reach USD 202.96 billion by 2020 (Grand View Research Inc., 2014). Introduction to aquaculture 3

1% 3%

11% 21% Mollusks Mollusks

Crustaceans 24% Crustaceans 10% Fish Fish 68% Other species 62% Other species

(A)(B) Fig. 1.1 World aquaculture by farmed species sector. (A) Aquaculture industry subsectors by quantity 2016 and (B) aquaculture industry subsectors by value 2016. Adapted from FAO, 2018. Leading species in global aquaculture production in 2016 (in million metric tons). Statista. Statista Inc. https://www.statista.com/statistics/240268/top-global- aquaculture-producing-countries-2010/ Accessed: July 14, 2019.

Most of the global aquaculture production (89% by volume) is located in countries in Asia (most notable being China, Indonesia, India, Vietnam, South Korea, and the Philippines; production involves mostly freshwater species, i.e., inland aquaculture). In terms of marine aquaculture, Norway’s Atlantic salmon industry makes it the world’s largest marine aquaculture producer, although this accounts for just 3% of total world farmed finfish output (Cooke, 2016). Another top aquaculture producer is Egypt, whose tilapia and shrimp farming have exponentially increased in the last two decades, and by 2017 at approximately 1.3 million metric tons, surpassed total production by Norway (1.19 million metric tons) and Chile (1.11 million metric tons) (Tran, 2019).

1.2.1 Aquatic animal species in aquaculture Among the three main aquaculture groups (fish, crustaceans, and mollusks), there were 369 finfishes (including five hybrids), 109 mollusks, and 64 crustaceans species recorded by FAO as cultured in 2016 (FAO, 2018). Carp dominates production in both China and the rest of Asia; for Europe and South America, it is salmonids; African aquaculture production is almost exclusively tilapias; for Oceania, shrimps and prawns dominate (Hall et al., 2011). In the USA, catfish predominate (Cooke, 2016). The most important farmed species (with a production of 1% or more of the total in 2016) (FAO, 2018) are listed in Table 1.1. However, the analysis of aquaculture production and details about farmed species remain approximations because many indigenous aquatic species are used in aquaculture without 4 Frederick S.B. Kibenge

Table 1.1 Major farmed species in world aquaculture. Aquaculture % of total, sector Species or species groupsa 2016 Finfish Grass carp (Ctenopharyngodon idellus) (FW) 11 Silver carp (Hypophthalmichthys molitrix) (FW) 10 Tilapias (Oreochromis niloticus) & other cichlids (FW) 10 Common carp (Cyprinus carpio) (FW) 8 Bighead carp (Hypophthalmichthys nobilis) (FW) 7 Crucian carp (Carassius carassius) (FW) 6 Major (Indian) carp (Catla catla) (FW) 6 Freshwater fishes nei, Osteichthyes (FW) 4 Atlantic salmon (Salmo salar) (D) 4 Roho labeo (Labeo rohita) (FW) 3 Pangas catfishes (Pangasius spp.) (FW) 3 Milkfish (Chanos chanos) (D) 2 Torpedo-shaped catfishes nei (Clarias spp.) 2 Marine fishes nei (Osteichthyes) (M) 2 Wuchang bream (Megalobrama amblycephala) (FW) 2 Rainbow trout (Oncorhynchus mykiss) (D) 1 Cyprinids (Cyprinidae) (FW) 1 Black carp or black Chinese roach (Mylopharyngodon 1 piceus) (FW) Northern snakehead (Channa argus) (FW) 1 Other finfishes 16 Finfish total 100 Crustaceans White leg shrimp (Litopenaeus vannamei, formerly 53 Penaeus vannamei) Red swamp crawfish (Procambarus clarkii)12 Chinese mitten crab (Eriocheir sinensis)10 Giant tiger prawn or Asian tiger shrimp (Penaeus 9 monodon) Oriental river prawn (Macrobrachium nipponense)4 Giant river prawn (Macrobrachium rosenbergii)3 Other crustaceans 9 Crustaceans 100 total Mollusks Cupped oysters nei (Crassostrea spp.) 28 Japanese carpet shell or Manila clam (Ruditapes 25 philippinarum) Scallops (Pectinidae)11 Introduction to aquaculture 5

Table 1.1 Major farmed species in world aquaculture—cont’d Aquaculture % of total, sector Species or species groups 2016 Marine mollusks (Mollusca)7 Sea mussels (Mytilidae)6 Chinese razor clam or Agemaki clam or constricted 5 tagelus (Sinonovacula constricta) Pacific oyster, Japanese oyster, or Miyagi oyster 3 (Crassostrea gigas) Blood cockle or Blood clam (Anadara granosa)3 Chilean mussel (Mytilus chilensis)2 Other mollusks 10 Mollusks total 100 aMajor farmed species or species groups (with a production of 1% or more of the total in 2016) (FAO, 2018). FW, freshwater; D, diadromous (these fish spawn in freshwater and migrate to the sea for the main growing period); M, marine. being registered individually in national statistics. Numerous single species registered in the official statistics of many countries consist, in reality, of multiple species and sometimes hybrids (i.e., the number of finfish hybrids in commercial production is more than five) (FAO, 2018). In China, more than 200 species are farmed commercially, but the total production is registered under fewer than 90 species and species groups. Similarly, in India and Vietnam, the number of cultured species far exceeds the number included in statistics (FAO, 2014). Moreover, for the most widely farmed species, tilapia, the correct number of producer countries is higher than the FAO record of 135 countries because commercially farmed tilapias are not yet reflected separately in national statistics in Canada and some European countries (FAO, 2014). In terms of the projected increase in aquaculture production, by 2028, carp and mollusks are projected to remain the most significant aquaculture groups, accounting for 35.8% and 19.2%, respectively, of total production (OECD/FAO, 2019).

1.2.2 Aquaculture techniques, systems, and facilities The techniques, systems, and facilities used in aquaculture are as diverse as the species currently being raised and also vary by geographical region (Bostock et al., 2010), but typically encompass three stages: incubation/ hatchery seed production, early-rearing, and on-growing (Roberts and Shepherd, 1974). In salmon and trout aquaculture, on-growing has merged with husbandry and breeding (Aarset and Borgen, 2015). For new farmed 6 Frederick S.B. Kibenge species like cod, mainly two methods are used: one is based on capturing wild cod for on-growing, the other focuses on the production of cod from hatchery seed to market size; currently, on-growing of wild cod (e.g., as in sea ranching (Kitada, 2018)) is more economically efficient than using farmed juveniles (Gunnarsson, 2007). For a few species, such as eels “Anguilla spp.,” farming still relies entirely on the wild seed. Although the Japanese eel “Anguilla japonica” life cycle in captivity has been completed, the techniques for mass production of glass eels have not yet been established because of various technical difficulties (Bird, 2013; Masuda et al., 2012). A comprehensive review of the different aquaculture systems worldwide can be found in Boyd and McNiven (2015). World aquaculture production takes place under four main systems: (1) on land (using freshwater or saline water) in static water ponds (e.g., for growing channel catfish and other warm-water species), concrete raceway systems (e.g., for growing rainbow trout), fresh water gravity tanks, or large recirculation tanks; (2) in floating cages and net pens in a lake or river (e.g., for growing salmonids and other cold-water species that require clean water with high oxygen levels) or on the shoreline/nearshore in sheltered waters (mariculture or marine aquaculture <3km from the shore) in floating net cages moored to the bottom in the sea with long lines (Shepherd, 1993; White and Edwards, 2015); (3) in coastal lagoons or brackish-water or marine ponds (e.g., for shellfish aquaculture) (Parisi et al., 2012; Wever et al., 2015; White and Edwards, 2015); and (4) in onshore-based tanks using pumped sea water (e.g., for flat fish farming of turbot or halibut that normally lie on the bottom and will not shoal within the water column inside a floating cage (Shepherd, 1993)).

1.2.3 Finfish aquaculture Seven of the top ten cultured fish species are carps; others are tilapias, Pangas catfishes, and Atlantic salmon (Table 1.1). China, the largest producer of farmed seafood, grows carp mostly in ponds and rice paddies, with little or no attempt to actively nurture the animals and has been practicing this type of aquaculture for thousands of years. Most carp farming is pond-based, with several species stocked in the same pond (polyculture) (SPC, 2011). Tilapia and sometimes catfish are stocked together with carp. Single-species culture (monoculture) is rare, except in the flow-through system and cage culture of common carp in streams or canals. There are different stocking models for polyculture, depending on the availability of the primary source of feed. If grasses (aquatic or terrestrial) are abundant, grass carp can be Introduction to aquaculture 7 stocked as the dominant species. The leftover feed and grass carp excreta would sufficiently fertilize the pond water for the growth of filter feeders. For example, the “80:20 pond fish culture” where 80% of harvest weight comes from the pellet-fed target species (such as grass carp, crucian carp, or tilapia) and the other 20% comes from the filter feeding “service species” (silver carp) (White and Edwards, 2015). Pond-based carp culture has been traditionally integrated with crop farming (rice, mulberry, fruit, and vege- tables) and animal husbandry (ducks, swine, and chicken) in China (White and Edwards, 2015). The practice has been widely introduced to many other parts of the world, with some modifications to fit into local conditions (SPC, 2011). Polyculture (the mixing of species in fish farms) has taken a different trajectory since 2008 and is widely used in sea cages for a different purpose in the form of using “cleaner fish” (lumpfish Cyclopterus lumpus and wrasse (Labridae)) in marine farmed salmonids as a biological control method for sea lice Lepeophtheirus salmonis in Europe (Powell et al., 2018; Skiftesvik et al., 2014) and Canada; it involves the use of wild-caught cleaner fish directly in the salmon farms or as broodstock for hatchery-raised cleaner fish (EURLFD, 2016). In industrial aquaculture, farmed fish are reared at high population densities in floating open net cages nearshore (for marine aquaculture) or in a lake or river (for inland aquaculture) (White and Edwards, 2015). The cages are either of square/hexagonal steel platform construction that are linked together or are of circular polyethylene rings 10–50m deep with netting to hold in farmed fish. Expansion can simply be by increasing the number of fish in a cage (i.e., cage volume) or by increasing the number of cages. Formulated dry feeds supply all nutrition. Uneaten feed and feces fall through the cages and settle to the bottom in the vicinity. Cages are periodically moved to new locations to allow benthic communities affected by sediment to recover—a process called fallowing (Boyd and McNiven, 2015). Current production systems operate on a single-year-class stocking in a site, usually a bay (“All-in, all-out”) with a fallow period of 6 months in between year classes. Closed containment is a practice that entails enclosing fish in floating containers or land-based farms to lessen their damage to the aquatic environment (Real, 2010). Recent improvements in offshore technology now allow even larger floating fish cages out in the sea that are submersible with automated feeding systems and remote monitoring. This has enabled the expansion of aquaculture farther offshore (i.e., deep-sea aquaculture) into high-energy sites and with a low environmental impact on the seafloor 8 Frederick S.B. Kibenge

(Lovatelli et al., 2013; Welch et al., 2019). China currently has one of the world’s biggest fully submerged net cage farming Atlantic salmon in the Yel- low Sea (Owen, 2018); and SalMar’s Ocean Farm 1 is 40km off the coast of Frohavet, Norway (Bennet, 2019). Moving fish farms out to sea helps to solve several problems. The open ocean offers more space, deeper waters, and stronger currents. Ocean currents are particularly effective for reducing the spread of pathogens and allowing better exchange and dilution of wastes than nearshore farming; greater distance offshore also minimizes interactions with wild fish and reduces the risk of unwanted exposure of nontarget organisms to aquacultural chemicals (Park et al., 2012). On land, increased development of land-based recirculating aquaculture system (RAS) (Espinal and Matulic, 2019) is leading to an ever-increasing move to saltwater land-based aquaculture, particularly in Europe and most recently in North America, despite the significant capital costs compared to those for floating cage farms. RAS offers environmentally sustainable methods for farming marine (Tal et al., 2009) and freshwater fish by having minimal effluents compared to other systems (White and Edwards, 2015). For example, in BC-Canada, a recent analysis for the Fraser Basin Council found that a capital investment of $1.1 billion would be required to establish a land-based industry capable of producing 50,000 tons of Atlantic salmon a year on Vancouver Island (Shore, 2019). There is already extensive knowledge derived from land-based hatcheries. The capital costs are compensated by lower operating costs, for example, land-based aquaculture farms have low pharmaceutical costs (e.g., no sea lice therapeutics) and no costs of maintaining boats that go out to feed the fish (Kramer, 2015). Moreover, with land-based farming, there would be no interactions between farmed fish and wild fish (Shore, 2019). However, while land- based aquaculture requires low antibiotic usage, there appears to be a connection in antimicrobial resistance issues found in land-based aquaculture with those established by terrestrial animal agriculture (Done et al., 2015). Another crucial disadvantage has been the development of “off-flavor” of the harvested fish “caused by geosmin and 2-methylisoborneol metabolites released by microbes that grow within the land-based systems” (Sapin et al., 2020). Another aquaculture development that has gained traction in recent years is the biofloc technology (White and Edwards, 2015). Biofloc technology is a technique of enhancing water quality in aquaculture through the manipulation of microbiota to convert the harmful waste from aquaculture production to consumable body biomass, with many other benefits (Dauda, Introduction to aquaculture 9

2019). The technology has recently gained attention as a sustainable method to maintaining optimum water quality parameters under a zero-water exchange system, thus preventing eutrophication and effluent discharge into the surrounding environment, thereby increasing biosecurity, as there is no water exchange except sludge removal (Ahmad et al., 2017). The technology has the added value of producing proteinaceous feed in situ, which results in higher productivity with less impact on the environment (Bossier and Ekasari, 2017). It has been considered the new “blue revolu- tion” in aquaculture (Emerenciano et al., 2017). It may act as a complete source of nutrition for aquatic organisms, along with some bioactive compounds that will enhance growth, survival, and defense mechanisms and acts as a novel approach for health management in aquaculture by stimulating the innate immune system of animals. Beneficial microbial bacterial floc and its derivative compounds such as organic acids, polyhydroxy acetate, and polyhydroxy butyrate could resist the growth of other pathogens and thus serve as a natural probiotic and immunostimulant. This technology is economically viable, environmentally sustainable, and socially acceptable (Ahmad et al., 2017; Daniel and Nageswari, 2017).

1.3 Mollusk aquaculture In the case of mollusk (oysters, clams, scallops, abalone, mussels, cockles, and related species) aquaculture, there are many varieties of culture and grow-out methods that are particular to species and locations (Boyd et al., 2005; Dı´az et al., 2011; Dumbauld et al., 2009; Elston, 1999; Elston and Ford, 2011; Parisi et al., 2012). For example, Parisi et al. (2012), in discussing the diversification of Italian shellfish culture, reviewed the most promising mollusk species and their culture in general (refer to Table 1.1 for a list of the most notable farmed mollusk species worldwide). Most mollusk culture requires no feed inputs (algae are used to feed the larvae and juveniles in hatcheries to provide spat for mollusk culturing) but may require control of fouling and predators. Seed bivalves are produced either in land-based hatcheries (Robert, 2009) or from natural populations (wild spat) (Parisi et al., 2012). They can be transported over considerable distances to grow-out sites in estuaries and even to different countries (FAO, 2013). Farming mollusks is limited in several countries because of limited capacity producing seed from mollusk hatcheries and nurseries (FAO, 2014), relying on the wild seed. The use of wild seed is impossible if the natural populations of mollusks are not suited to culture. 10 Frederick S.B. Kibenge

Intensification will require a reliable, plentiful, and inexpensive supply of hatchery seed, requiring the selection and management of broodstocks and the controlled production of high-quality seed stocks (Elston and Ford, 2011). This would reduce the environmental biosecurity concern of the translocation of non-native species (Parisi et al., 2012).

1.4 Crustacean aquaculture The most recent dramatic change in aquaculture has been the explo- sive growth in shrimp farming in Southeast Asia (Cressey, 2009). The production has increased almost exponentially since the mid-1970s (FAO, 2010). This rapid increase largely reflects the dramatic increase in white leg shrimp Litopenaeus vannamei (formerly Penaeus vannamei) culture in Asia (China, Thailand and Indonesia) and Latin America (Ecuador and Mexico) (Bondad-Reantaso et al., 2012), where the penaeid shrimps have tended to dominate due to high-value, short-production cycles and accessible technologies (Bostock et al., 2010). Crustacean farming, similarly to marine fish aquaculture, requires a high-quality diet, usually containing fish meal and often fish oil (Bostock et al., 2010).

1.5 Chemicals in aquaculture There is a long list of chemicals that are used in aquaculture, including agricultural limestone, lime, fertilizers, oxidants, coagulants, osmoregulators, algicides, herbicides, fish toxicants, antifoulants, therapeutics, disinfectants, anesthetics, agricultural pesticides, and hormones (Arthur et al., 2000; Boyd and Massaut, 1999; Boyd and Tucker, 1998; Noga, 2010; Schnick, 2001). To mitigate the environmental impact associated with the use of aquacultural chemicals, various regulations have been advocated (FAO, 1997). In many countries, governmental regulations on chemical use in aquaculture have been established (Cabello, 2006; Schnick, 1988, 2001) (refer also to Martins and Martins (2020) in Chapter 3 (Section 3.4), Loro (2020) in Chapter 5 (Section 5.4), and de Assis (2020) in Chapter 6 in this book. Most importing countries also have food safety regulations that involve the inspection of incoming shipments of seafood products for specific chemical residues. A companion textbook, Aquaculture Pharmacology (Kibenge et al., 2020), has a comprehensive review of most aquaculture pharmaceuticals. The pharmaceuticals used in human and veterinary medicine (including aquaculture) and their by-products enter the aquatic Introduction to aquaculture 11 environment as pollutants in a variety of ways, causing adverse effects on nontarget aquatic organisms. The aquatic environment may also be contaminated by chemicals such as agrochemicals, metals, oils and their derivatives, and microplastics as a result of erosion of surface deposits of metal minerals, as well as from human activities, such as agriculture, mining, smelting, fossil fuel combustion, and industrial and commercial uses of metals (Nordberg et al., 2007; Pandey et al., 2019). Martinez et al. discuss metal pollutants that are toxic to aquatic animals in Chapter 4, Loro discusses agrochemical pollutants in Chapter 5, de Assis discusses pharmaceutical pollutants in Chapter 6, Sadauskas-Henrique et al. discuss oil and derivatives pollutants in Chapter 7, and de Arau´jo et al. discuss microplastics and associated pollutants in Chapter 8 in this book. Additional information on chemicals in aquaculture and the contribution of aquaculture to chemical pollution and other adverse impacts of chemical use by humans can be found in Boyd and McNiven (2015).

1.6 Governance of aquaculture The FAO (2017) technical guidelines on “aquaculture governance and sector development” define aquaculture governance as “the set of processes by which a jurisdiction manages its resources concerning aquaculture, how its stakeholders participate in making and implementing decisions affecting the sector, how government personnel are accountable to the aquaculture community and other stakeholders, and how the respectful rule of law is applied and enforced.” Most countries have some form of legislation governing aquaculture operations (with regard to animal welfare, animal diseases, the environment, water use, land use, and waste disposal), with the competent authorities residing in various government ministries and agencies. Given the diversity of countries and aquaculture systems, it is beyond the scope of this chapter to consider legislation in each country. Instead, the reader is referred to a series of comparative national overviews of aquaculture laws and regulations, “National Aquaculture Legislation Overview (NALO),” from aquaculture producing countries that are published by the FAO as NALO Fact sheets (FAO, 2019). These fact sheets are updated every 2 to 3 years (FAO, 2019). At the national level, aquaculture operations require various licenses and permits to operate, ensuring environmental protection relating to the waste generated, prevention, and control of aquatic animal diseases, and the well-being of aquaculture animals (FAO, 2019). There are also Policy Frameworks mandated 12 Frederick S.B. Kibenge nationally, regionally, or internationally on how aquaculture activities are developed and managed. These guidelines and codes of conduct provide good practices for a variety of aquaculture operations and how the activities are regulated in both freshwater and marine environments as well as for land-based aquaculture. The ultimate purpose of governance of the aquaculture sector should be to support sustainable aquaculture by emphasizing environmental sustainability and social responsibility. Enhancement of aquaculture’s growth needs the right policy decisions regarding the use of natural resources (water, land, seed, and feed), as well as sound environmental management. Good governance is fundamental to the successful formulation and implementation of aquaculture development policies, strategies, and plans (FAO, 2016; Hishamunda et al., 2014). Three main environmental actions arising from good governance are: (1) management of effluent and nutrient loading, (2) improvement of aquaculture legislation, and (3) mandatory environmental impact assessment (EIA) (FAO, 2016; Hishamunda et al., 2014).

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Cabello, F.C., 2006. Heavy use of prophylactic antibiotics in aquaculture: a growing problem for human and animal health and for the environment. Environ. Microbiol. 8, 1137– 1144. Cooke, M. 2016. BBFAW Investor Briefing – Animal Welfare in Farmed Fish. Available from https://www.bbfaw.com/media/1432/investor-briefing-no-23-animal-welfare- in-farmed-fish.pdf (Accessed October 03, 2019). Cressey, D., 2009. Future fish. Nature 458, 398–400. Daniel, N., Nageswari, P., 2017. Exogenous probiotics on biofloc based aquaculture: a review. Curr. Agric. Res. J. 5, 88–107. Dauda, A.B., 2019. Biofloc technology: a review on the microbial interactions, operational parameters and implications to disease and health management of cultured aquatic animals. Rev. Aquacult., 1–18. https://doi.org/10.1111/raq.12379. de Assis, H.C.S. 2020. Chapter 6 Pharmaceutical pollutants. In Aquaculture Toxicology. Edited by Kibenge, F.S.B., Baldisserotto, B., Chong, R.S.M. Elsevier Inc, Academic Press. (in preparation). Dıáz, G.A., Seijo, L.C., Gaspar, M.B., da Costa Gonza´lez, F. 2011. Razor Clams: Biology, Aquaculture and Fisheries. Consellerıá do Mar, Xunta de Galicia ed., Santiago de Comp- ostela, Spain. Available from: ftp://ftp.cimacoron.org/razor_clams_2011.pdf (Accessed September 16, 2019). Done, H.Y., Venkatesan, A.K., Halden, R.U., 2015. Does the recent growth of aquaculture create antibiotic resistance threats different from those associated with land animal production in agriculture? Am. Assoc. Pharm. Sci. J. 17, 513–524. https://doi.org/10.1208/ s12248-015-9722-z. Dumbauld, B.R., Ruesink, J.L., Rumrill, S.S., 2009. The ecological role of bivalve shellfish aquaculture in the estuarine environment: a review with application to oyster and clam culture in west coast (USA) estuaries. Aquaculture 290, 196–223. Elston, R.A. 1999. Health Management, Development and Histology of Seed Oysters. World Aquaculture Society, Baton Rouge, LA, USA. 110 pp. Elston, R.A., Ford, S.E., 2011. Chapter 13 Shellfish diseases and health management. In: Shumway, S.E. (Ed.), Shellfish Aquaculture and the Environment, first ed. John Wiley & Sons Inc, pp. 359–394. Emerenciano, M.G.C., Martıńez-Co´rdova, L.F., Martıńez-Porchas, M., Miranda-Baeza, A., 2017. Chapter 5 Biofloc Technology (BFT): A Tool for Water Quality Management in Aquaculture. Intech, https://doi.org/10.5772/66416. Espinal, C.A., Matulic, D., 2019. Chapter 3 Recirculating aquaculture technologies. In: Goddek, S., Joyce, A., Kotzen, B., Burnell, G.M. (Eds.), Aquaponics Food Production Systems. Springer Nature, Switzerland AG, pp. 35–76, https://doi.org/10.1007/978-3- 030-15943-6_3. EURLFD (European Union Reference Laboratory for Fish Diseases). 2016. Scientific meeting report on cleaner fish in aquaculture 2016. https://www.eurl-fish-crustacean.eu/ fish/scientific-reports (Accessed August 01, 2018). FAO, 2010. The State of World Fisheries and Aquaculture 2010. Rome, FAO. 2010. 197 pp. FAO, 2013. Hatchery Culture of Bivalves: A Practical Manual. FAO Fisheries Technical Paper 471. Accessible online ftp://ftp.fao.org/docrep/fao/007/y5720e/y5720e00.pdf. FAO, 2014. The State of World Fisheries and Aquaculture 2014. FAO, Rome. 223 pp. FAO, 2015. Fishery statistical collections: global aquaculture production. Accessible online http://www.fao.org/fishery/statistics/global-aquaculture-production/en Accessed: July 14, 2019. FAO, 2016. The State of World Fisheries and Aquaculture 2016. Contributing to Food Security and Nutrition for all. FAO, Rome. 200 pp. FAO, 2017. Aquaculture Development. 7. Aquaculture governance and sector development. FAO Technical Guidelines for Responsible Fisheries. No. 5. Suppl. 7. Rome, Italy. 50 pp. 14 Frederick S.B. Kibenge

FAO, 2018.Leading Species in Global Aquaculture Production in 2016 (in million metric tons). Statista. Statista Inc. https://www.statista.com/statistics/240268/top-global- aquaculture-producing-countries-2010/ Accessed: July 14, 2019. FAO, 2019. National Aquaculture Legislation Overview (NALO) Fact Sheets for different countries. Available at http://www.fao.org/fishery/collection/nalo/en (web archive link, 01 September 2019) and http://www.fao.org/fishery/nalo/search/en (Accessed September 1, 2019). FAO (Food and Agriculture Organization of the United Nations), 1997. Aquaculture Devel- opment. FAO Technical Guidelines for Responsible Fisheries 5. FAO, Rome. Grand View Research, Inc., 2014. Aquaculture Market Analysis By Culture Environment (Fresh Water, Marine Water, Brackish Water), By Product (Carps, Crustaceans, Mack- erel, Milkfish, Mollusks, Salmon, Sea Bass, Sea Bream, Trout) and Segment Forecasts To 2020. Available at: http://www.grandviewresearch.com/industry-analysis/aquaculture- market and at: http://www.thefishsite.com/fishnews/23604/world-aquaculture- market-to-reach-20296-billion-by-2020#sthash.STzI17Ql.dpuf (Accessed: April 02, 2016). Gunnarsson, V.I., 2007. http://www.fisheries.is/aquaculture/species/cod/. (Accessed 2 April 2016). Hall, S.J., Delaporte, A., Phillips, M.J., Beveridge, M., O’Keefe, M., 2011. Blue Frontiers: Managing the Environmental Costs of Aquaculture. The WorldFish Center, Penang, Malaysia. http://www.conservation.org/publications/documents/BlueFrontiers_ aquaculture_report.pdf. (Accessed 2 April 2016). Hishamunda, N., Ridler, N., Martone, E. 2014. Policy and Governance in Aquaculture: Les- sons Learned and Way Forward. FAO Fisheries and Aquaculture Technical Paper No. 577. Rome, FAO. 48pp. Kibenge, F.S.B., Baldisserotto, B., Chong, R.S.M. (Eds.), 2020. Aquaculture Pharmacology. Elsevier Inc, Academic Press. in preparation. Kitada, S., 2018. Economic, ecological and genetic impacts of marine stock enhancement and sea ranching: a systematic review. Fish Fish. 19, 511–532. Kramer, L. 2015. Land-based salmon aquaculture: a future with potential. Available at https://www.seafoodsource.com/news/aquaculture/land-based-salmon-aquaculture- a-future-with-potential (Accessed September 16, 2019). Loro, V.L., 2020. Chapter 5 Agrochemicals: Contamination of environment, effects on wildlife, strategies for mitigating impacts and new perspectives to legislation. In: Kibenge, F.S. B., Baldisserotto, B., Chong, R.S.M. (Eds.), Aquaculture Toxicology. Elsevier Inc, Aca- demic Press. in preparation. Lovatelli, A., Aguilar-Manjarrez, J., Soto, D. (Eds.) 2013. Expanding mariculture farther offshore—Technical, environmental, spatial and governance challenges. FAO Technical Workshop. 22–25 March. 2010. Orbetello, Italy. FAO Fisheries and Aquaculture Pro- ceedings No. 24. Rome, FAO. 73 pp. Includes a CD–ROM containing the full document (314 pp). (also available at http://www.fao.org/docrep/018/i3092e/i3092e00. htm). Martins, S.E., Martins, C.M.G., 2020. Chapter 3 Antifoulants and disinfectants. In: Kibenge, F.S.B., Baldisserotto, B., Chong, R.S.M. (Eds.), Aquaculture Toxicology. Elsevier Inc, Academic Press. in preparation. Masuda, Y., Imaizumi, H., Oda, K., Hashimoto, H., Usuki, H., Teruya, K., 2012. Artificial completion of the Japanese Eel, Anguilla japonica, life cycle: challenge to mass production. Bull. Fisheries Res. Agency 35, 111-117. Accessible online https://www.fra.affrc.go.jp/ bulletin/bull/bull35/35-13.pdf. Noga E.J. 2010. Fish Disease, Diagnosis and Treatment, second ed. Iowa State Press, Ames, Iowa. 536 pp. Introduction to aquaculture 15

Nordberg, G.F., Fowler. B.A., Nordberg, M., Friberg, L.T., 2007. Handbook on the Tox- icology of Metals, third ed. Elsevier, Amsterdam. 1024 pp. OECD/FAO, 2019. OECD-FAO Agricultural Outlook 2019-2028. OECD Publishing, Paris/Food and Agriculture Organization of the United Nations, Rome, https://doi. org/10.1787/agr_outlook-2019-en. Owen, E., 2018. China gets ready to harvest first batch of farmed salmon from huge, deep sea fully-submersible fish cage. Available at https://salmonbusiness.com/chinas-gets-ready- to-harvest-first-batch-of-farmed-salmon-from-huge-deep-sea-fully-submersible-fish- cage/ (Accessed November 21, 2018). Pandey, L.K., Park, J., Son, D.H., Kim, W., Islam, M.S., Choi, S., Lee, H., Han, T., 2019. Assessment of metal contamination in water and sediments from major rivers in South Korea from 2008 to 2015. Sci. Total Environ. 651, 323–333. Parisi, G., Centoducati, G., Gasco, L., Gatta, P.P., Moretti, V.M., Piccolo, G., Roncarati, A., Terova, G., Pais, A., 2012. Molluscs and echinoderms aquaculture: biological aspects, current status, technical progress and future perspectives for the most promising species in Italy. Ital. J. Anim. Sci. 11, 4. https://doi.org/10.4081/ijas.2012.e72. Park, Y.H., Hwang, S.Y., Hong, M.K., Kwon, K.H., 2012. Use of antimicrobial agents in aquaculture. Revue Scientifique et Technique (International Office of Epizootics) 31, 189–197. Powell, A., Treasurer, J.W., Pooley, C.L., Keay, A.J., Lloyd, R., Imsland, A.K., de Leaniz, C.G., 2018. Use of lumpfish for sea-lice control in salmon farming: challenges and opportunities. Rev. Aquac. 10, 683–702. Real, N. 2010. DFO recognises potential of closed containment methods. Available at http://fis.com/fis/worldnews/worldnews.asp?l¼e&ndb¼1&id¼39422. Robert, R., 2009. Review on molluscs hatchery nursery development in Europe. pp 1–12. In: Proc. Symp. EAS, EUROSHELL Session, Trondheim, Norway. Roberts, R.J., Shepherd, C.J., 1974. Handbook of Trout and Salmon Diseases. Fishing News, Surrey, England, pp. 40–42. Sapin, R., Korban, D., Unlay, N., 2020. What are land-based salmon companies doing to tackle off flavors? Available at https://www.intrafish.com/aquaculture/we-asked- every-major-land-based-salmon-farmer-how-theyre-tackling-off-flavor-problems- heres-what-they-said-/2-1-752023?utm_campaign¼IFCO%3A%20Monthly% 20Land-Based%20Salmon%20-%20Mar%202020&utm_source¼hs_email&utm_ medium¼email&utm_content¼85265239&_hsenc¼p2ANqtz-_ 9KtqcaHfzksx8cItMBR6noN- IZC4x3a4Rwm7yjAelXkIOT7Kz288k2I8MmMU6zG-JbLN8bSlWeT4UQi4KtcC_ ALW1EA&_hsmi¼85266785 (Accessed March 26, 2020). Schnick, R.A., 1988. The impetus to register new therapeutants for aquaculture. Progressive Fish-Culturist 50, 190–196. Schnick, R.A., 2001. Aquaculture chemicals. In: Kirk-Othmer Encyclopedia of Chemical Technology, fourth ed. vol. 3. John Wiley & Sons, New York, pp. 209–225. Shepherd, J., 1993. Aquaculture systems. In: Brown, L. (Ed.), Aquaculture for Veterinarians: Fish Husbandry and Medicine. Pergamon Press Ltd., Oxford, pp. 43–55. Shore, R. 2019. U.S. firm seeks stake in B.C.’s only land-based Atlantic salmon farm. https:// vancouversun.com/news/local-news/u-s-firm-seeks-majority-stake-in-b-cs-only- land-based-salmon-farm. Skiftesvik, A.B., Blom, G., Agnalt, A.-L., Durif, C.M.F., Browman, H.I., Bjelland, R.M., Harkestad, L.S., Farestveit, E., Paulsen, O.I., Fauske, M., Havelin, H., Johnsen, K., Mortensen, S., 2014. Wrasse (Labridae) as cleaner fish in salmonid aquaculture—the Hardangerfjord as a case study. Mar. Biol. Res. 10, 289–300. https://doi.org/ 10.1080/17451000.2013.810760. 16 Frederick S.B. Kibenge

SPC, 2011. SPC Aquaculture Portal, Commodities Carp. Accessible here https://www.spc. int/aquaculture/images/commodities/pdf/Carp_page.pdf. Tal, Y., Schreier, H.J., Sowers, K.R., Stubblefield, J.D., Place, A.R., Zohar, Y., 2009. Envi- ronmentally sustainable land-based marine aquaculture. Aquaculture 286, 28–35. Tran, K., 2019. Egypt tops Norway, Chile in Aquaculture Production. https://www. intrafish.com/aquaculture/egypt-tops-norway-chile-in-aquaculture-production/2-1- 707157. (Accessed 22 April 2020). Welch, A.W., Knapp, A.N., El Tourky, S., Daugherty, Z., Hitchcock, G., Benetti, D., 2019. The nutrient footprint of a submerged-cage offshore aquaculture facility located in the tropical Caribbean. J. World Aquacult. Soc. 50, 299–316. https://onlinelibrary.wiley. com/doi/full/10.1111/jwas.12593. Wever, L., Krause, G., Buck, B.H., 2015. Lessons from stakeholder dialogues on marine aquaculture in offshore windfarms: perceived potentials, constraints and research gaps. Mar. Policy 51, 251–259. White, P., Edwards, P., 2015. Types of culture systems and aqua-ecosystems. Available at http://aquaculture.management/2015/types-of-culture-systems-and-aqua- ecosystems/ (Accessed March 26, 2020). World Bank, 2013. Fish to 2030: Prospects for Fisheries and Aquaculture (English). Agricul- ture and Environmental Services Discussion Paper; no. 3. World Bank Group, Washing- ton, DC. http://documents.worldbank.org/curated/en/458631468152376668/Fish- to-2030-prospects-for-fisheries-and-aquaculture. (Accessed 3 October 2019). CHAPTER TWO

General introduction to toxicology of aquatic animals

Bernardo Baldisserotto Department of Physiology and Pharmacology, Universidade Federal de Santa Maria, Santa Maria, RS, Brazil

2.1 Introduction to toxicology Laboratory studies dealing with exposure to pollutants can last for hours to a few days (acute experiments) and are usually used to calculate the concentration that affects (effective concentration—EC50) or kills (lethal concentration—LC50) 50% of test organisms. Chronic studies are used to determine threshold concentrations (no observed effect concentration, or NOEC; lowest observed effect concentration, or LOEC). These values can be used to compare the effect of a given pollutant in different species or different pollutants between themselves. These values are used worldwide to derive water quality guidelines for the protection of aquatic life, i.e., concentrations that should result in a negligible effect on aquatic biota. Two basic guideline derivation methodolo- gies are in use throughout the world to calculate these concentrations (Nugegoda and Kibria, 2013): one is the assessment factor method, which involves multiplying the lowest value of a set of toxicity data by a factor to arrive at a criterion value. For example, the LC50 value is multiplied by 0.01 and the NOEC value by 0.1. The assessment factor can vary according to the country and the amount of toxicological studies available. The other method is the species sensitivity distribution, a statistical extrap- olation involving the use of individual toxicity data for a range of species to determine the criterion value. This method estimates the concentration of a substance that is predicted to protect 95% of all aquatic species extrapolating from an equation or curve that considers the relationship between probabil- ity x log LC50 and NOEC. Some countries use either one of the methods, and others use a combination of both methods.

2.2 Water quality criteria/guidelines The United States Environmental Protection Agency (USEPA) has water quality criteria/guidelines for water quality, and its website contains tables with the highest concentration of the most common metals, pesticides (mainly insecticides), and other pollutants or contaminants that are not expected to pose a significant risk to the majority of species in a given environment (USEPA, 2019). The same agency maintains the ECOTOX Knowledgebase Explore function (https://cfpub.epa.gov/ecotox/explore. cfm), which is an interactive way to examine search paths by chemical, species, and effects. The guidelines of the Canadian Council of Ministers of the Environment (CCME, 2003) explain that some water quality parameters can affect the toxicity of contaminants. Some are described later. – pH: At alkaline pH, many metals form insoluble hydroxides or carbon- ates that precipitate, while at acidic pH, a higher release of metals from the sediments can occur (CCME, 2003). Water pH may also influence toxicity in pharmaceuticals because the toxicity of sertraline and fluoxetine was higher and diclofenac was lower at pH8.2 than at pH5.8, but the toxicity of ethinyl estradiol was not affected by pH. Probably these differences are due to the higher uptake of sertraline at pH8.2 and no change in ethinyl estradiol uptake by pH (Alsop and Wilson, 2019). The piscicide 3-trifluoromethyl-4-nitrophenol (TFM), which is applied in the Great Lakes to control streams infested with larval sea lamprey (Petromyzon marinus), has 5.5-fold higher uptake rates at low pH (6.86) compared to alkaline pH (8.78). This higher uptake is most likely because the unionized, lipophilic form of TFM exists in greater amounts at a lower pH. On the contrary, the elimination rates are 1.7–1.8-fold higher at pH8.78 than at pH6.86, indicating that TFM is more effective at low pH (Hlina et al., 2017). – Hardness and alkalinity: The increase of water hardness usually reduces the toxicity of many metals, largely due to the formation of metal- carbonate complexes and Ca2+ and/or Mg2+ antagonism (Blewett and Leonard, 2017). CCME (2003) presents equations in which the values of water hardness are used to predict the concentrations of zinc and copper to protect aquatic life. The toxicity of the insecticides endosulfan (Thiodan; 1,4,5,6,7-hexachloro-8,9,10-trinorborn-5-en-2,3-yelen) General introduction to toxicology of aquatic animals 19

(dimethyl) sulfite and methiocarb [Mesurol; 3,5-dimethyl-4-(methylthio) phenyl methylcarbamate] to rainbow trout, Oncorhynchus mykiss,islower at higher alkalinity (40–121mg CaCO3/L) than at lower alkalinity (19–20mg CaCO3/L), but hardness ranging from 50 to 147mg CaCO3/L had no effect (Altinok et al., 2006; Capkin et al., 2006). Hardness (6–309mg CaCO3/L) also did not affect the toxicity of another insecticide, fenvalerate, to bluegill (Lepomis macrochirus)(Dyer et al., 1989). The increase of hardness increases surfactant toxicity to aquatic species in some cases, but the outcome is compound and species-specific (Lewis, 1992). – Salinity: The toxicity of metals in general decreases with the increase of salinity due to complexation with seawater ions (Blewett and Leonard, À2 2017; Tan et al., 2019). For example, nickel may complex with SO4 À and Cl and free nickel reduces (Blewett and Leonard, 2017). The same À complexation with Cl can be observed with cadmium (Tan et al., 2019). The uptake of cadmium and zinc by the teleost Acanthopagrus schlegelii reduced with the increase of salinity, probably due to the higher waterborne Ca2+ levels at higher salinities, because both metals use the Ca2+ uptake pathway (Zhang and Wang, 2007). Salinity does not affect LC50 of fenvalerate to bluegill, but the accumulation of this insecticide decreased with the increase of salinity (Dyer et al., 1989). – Total dissolved solids: It can change the complexing and precipitation processes of metals (CCME, 2003). The presence of suspended solids and naturally occurring dissolved substances decreases the bioavailability of cationic surfactants but not that of anionic and nonionic surfactants (Lewis, 1992). – Temperature: Besides the direct effect of temperature on the metabolism of ectothermic animals, it can also affect the toxicity of several pollutants. The effect depends on the pollutant (CCME, 2003). – Dissolved oxygen: The toxicity of pollutants can be magnified by low dissolved oxygen levels (CCME, 2003). – Dissolved organic matter (DOM): This can also affect the toxicity of metals. The major chemical components of DOM are humic substances (humic and fulvic acids). Its protective effect varies according to its origin and composition. Overall, darker organic matter is expected to be more protective against metal toxicity (Al-Reasi et al., 2013). DOC also decreased uptake and toxicity of sertraline, but not of ethinyl estradiol (Alsop and Wilson, 2019). 20 Bernardo Baldisserotto

Water quality guidelines in China are in the “Surface Water Environmental Quality Standards” GB3838-2002 (http://www.codeofchina.com/ standard/GB3838-2002.html). This guide considers that class II water is for “precious fish reserves” and up to class III water is adequate for “winter- ing grounds of fish and shrimp, migration channels, aquaculture areas of fishing waters.” The fourth amendment of this guideline used the data from native species and the assessment factor method and/or species sensitivity distribution to calculate concentrations that should result in a negligible effect on aquatic biota (Zhao et al., 2018). However, the eutrophication control management is incomplete compared to developed countries because it is based only on total nitrogen and total phosphorus values (Su et al., 2017). South Africa has a very recent and complete guideline of water quality for coastal marine waters for use to mariculture with maximum values for several pollutants, based on the values from USEPA, CCME, and EU (2013), as well as from other African countries (RSA-DEA, 2018). From Latin America, Brazil follows the resolution from the Brazilian National Environment Council (CONAMA, 2005), which has specific tables with maximum values for contaminants for fisheries and aquaculture for fresh-, brackish-, and saline waters. Mexico, on the contrary, uses only four water quality parameters (biochemical and chemical demand of oxygen, fecal coli- forms, total suspended solids) to indicate water quality (Comisio´n Nacional del Agua, 2018). This country has also a law that states maximum values of fats, oils, and a few metals (NOM, 1998).

2.3 Intraspecies variation of toxicity The effects of metals and pharmaceuticals in fish can be affected by sex, size, “personality” (defined as bold or shy, according to their risk-taking behaviors), and social status. For example, different responses were observed in males and females of zebrafish, Danio rerio, to cocaine, and subordinate rainbow trout are more vulnerable than dominant ones to copper and silver exposure (Demin et al., 2019). Survival of rainbow trout exposed to endosulfan was significantly increased with increasing fish size (Capkin et al., 2006), but the opposite was observed in those exposed to methiocarb (Altinok et al., 2006). General introduction to toxicology of aquatic animals 21

2.4 Models to predict toxicity of contaminants 2.4.1 Biotic ligand model (BLM) The USEPA uses the BLM to predict dissolved metal concentrations that correspond to lethal accumulations of a metal on biotic ligands for unique water compositions, i.e., depends on the concentrations of some cations (e.g., K+,Na+,Ca2+,Mg2+, and H+). BLMs are computational models that determine metal speciation and predict metal toxicity to biota in aqueous systems and combine an equilibrium geochemical speciation model, a metal–organic binding model, and a toxicological model (Smith et al., 2015). This methodology was applied mainly for copper (USEPA, 2007), but there are data from the USA and Canadian species for other metals and there are also proposals to analyze the mixture of metals (Balistrieri and Mebane, 2014). The binary metal interactions tested (silver, cadmium, copper, nickel, lead, and zinc) demonstrated interactions, mostly inhibitory, nonreciprocal, and caused by silver or copper (Cremazy et al., 2019). Nev- ertheless, BLMs must be constructed for different water qualities and species. For example, LC50 for copper in cardinal tetra Paracheirodon axelrodi from the ion-poor blackwater Rio Negro, Amazon (high level of DOC) was not predicted properly using BLM models developed using temperate DOC and temperate species (Cremazy et al., 2016).

2.4.2 Toxicokinetic-toxicodynamic (TK-TD) models These models consider the processes that lead to toxicity at the organisms over time instead of at a certain endpoint (96h, for example), as the BLM. Nevertheless, these models use BLM-estimated binding constants calculated with toxicity data. The TK-TD models simulate the temporal aspects of toxicity and provide a conceptual framework to better understand the causes for variability in different species’ sensitivity to the same compound as well as the causes for different toxicity of different compounds to the same species (Giesy et al., 2010). Recent models can predict metal toxicity as a function of the waterborne concentrations of some cations (Feng et al., 2018a) and the mixture of metals (Gao et al., 2016). The interactions between metals may be time-dependent (Feng et al., 2018b). These models are also useful to extrapolate toxicity of pesticides from relatively 22 Bernardo Baldisserotto constant to relatively variable exposure profiles, because in the environment, they generally occur in fluctuating and highly variable patterns (Ashauer et al., 2013).

2.4.3 Principal component analysis (PCA) The PCA is a multivariate statistical procedure that uses an orthogonal transformation to convert a set of correlated variables into a set of values of lin- early uncorrelated variables called principal components, where the first principal component has the largest possible variance and each succeeding component in turn has the highest variance possible. Plotting the principal components can elucidate potential correlations between the set of variables (Marques et al., 2019).

2.4.4 Generalized additive modeling (GAM) The GAM approach is an extension of the generalized linear model (GLM). It is a generalized linear model with a linear predictor involving a sum of smooth functions of covariates that allows a flexible description of complex responses to environmental changes. Therefore, it can allow statistical infer- ence to nonlinear correlations between the contaminant levels and biomarker responses. Consequently, it provides a model to predict the bioaccumulation of chemical contaminants based on biomarker responses and vice-versa, besides other influencing factors, such as seasonality and gra- dients in water chemistry parameters (Marques et al., 2019).

References Al-Reasi, H.A., Wood, C.M., Smith, D.S., 2013. Characterization of freshwater natural dissolved organic matter (DOM): mechanistic explanations for protective effects against metal toxicity and direct effects on organisms. Environ. Int. 59, 201–207. Alsop, D., Wilson, J.Y., 2019. Waterborne pharmaceutical uptake and toxicity is modified by pH and dissolved organic carbon in zebrafish. Aquat. Toxicol. 210, 11–18. Altinok, I., Capkin, E., Karahan, S., Boran, M., 2006. Effects of water quality and fish size on toxicity of methiocarb, a carbamate pesticide, to rainbow trout. Environ. Toxicol. Pharmacol. 22 (1), 20–26. Ashauer, R., Thorbek, P., Warinton, J.S., Wheeler, J.R., Maund, S., 2013. A method to predict and understand fish survival under dynamic chemical stress using standard ecotoxicity data. Environ. Toxicol. Chem. 32 (4), 954–965. Balistrieri, L.S., Mebane, C.A., 2014. Predicting the toxicity of metal mixtures. Sci. Total Environ. 466, 788–799. Blewett, T.A., Leonard, E.M., 2017. Mechanisms of nickel toxicity to fish and invertebrates in marine and estuarine waters. Environ. Pollut. 223, 311–322. Capkin, E., Altinok, I., Karahan, S., 2006. Water quality and fish size affect toxicity of endosulfan, an organochlorine pesticide, to rainbow trout. Chemosphere 64 (10), 1793–1800. General introduction to toxicology of aquatic animals 23

CCME (Canadian Council of Ministers of the Environment), 2003. Canadian Water Quality Guidelines for the Protection of Aquatic Life: Guidance on the Site-Specific Application of Water Quality Guidelines in Canada: Procedures for Deriving Numerical Water Quality Objectives. Comisioń Nacional del Agua, 2018. Estadisticas del Agua en Mexico. Secretarıá de Medio Ambiente y Recursos Naturales, Ciudad de Mexico. CONAMA (Conselho Nacional do Meio Ambiente), 2005. Resoluçaõ CONAMA Nº 357, from March 17th, 2005. Cremazy, A., Wood, C.M., Smith, D.S., Ferreira, M.S., Johannsson, O.E., Giacomin, M., Val, A.L., 2016. Investigating copper toxicity in the tropical fish cardinal tetra (Paracheirodon axelrodi) in natural Amazonian waters: measurements, modeling, and reality. Aquat. Toxicol. 180, 353–363. Cremazy, A., Brix, K.V., Wood, C.M., 2019. Using the biotic ligand model framework to investigate binary metal interactions on the uptake of Ag, Cd, Cu, Ni, Pb and Zn in the freshwater snail Lymnaea stagnalis. Sci. Total Environ. 647, 1611–1625. Demin, K.A., Lakstygal, A.M., Alekseeva, P.A., Sysoev, M., Abreu, M.S., Alpyshov, E.T., Serikuly, N., Wang, D.M., Wang, M.Y., Tang, Z.C., Yan, D.N., Strekalova, T.V., Vol- gin, A.D., Arnstislayskaya, T.G., Wang, J.J., Song, C., Kalueff, A.V., 2019. The role of intraspecies variation in fish neurobehavioral and neuropharmacological phenotypes in aquatic models. Aquat. Toxicol. 210, 44–55. Dyer, S.D., Coats, J.R., Bradbury, S.P., Atchison, G.J., Clark, J.M., 1989. Effects of water hardness and salinity on the acute toxicity and uptake of fenvalerate by bluegill (Lepomis macrochirus). Bull. Environ. Contam. Toxicol. 42 (3), 359–366. EU (European Union), 2013. Directive 2013/39/EU of the European Parliament and of the Council of 12 August 2013. Off. J. Eur. Union 2008, L348/84-97. Feng, J.F., Gao, Y.F., Chen, M., Xu, X., Huang, M.D., Yang, T., Chen, N., Zhu, L., 2018a. Predicting cadmium and lead toxicities in zebrafish (Danio rerio) larvae by using a toxicokinetic-toxicodynamic model that considers the effects of cations. Sci. Total Envi- ron. 625, 1584–1595. Feng, J.F., Gao, Y.F., Ji, Y.J., Zhu, L., 2018b. Quantifying the interactions among metal mixtures in toxicodynamic process with generalized linear model. J. Hazard. Mater. 345, 97–106. Gao, Y.F., Feng, J.F., Han, F., Zhu, L., 2016. Application of biotic ligand and toxicokinetic- toxicodynamic modeling to predict the accumulation and toxicity of metal mixtures to zebrafish larvae. Environ. Pollut. 213, 16–29. Giesy, J.P., Naile, J.E., Khim, J.S., Jones, P.D., Newsted, J.L., 2010. Aquatic toxicology of perfluorinated chemicals. Rev. Environ. Contam. Toxicol. 202, 1–52. Hlina, B.L., Tessier, L.R., Wilkie, M.P., 2017. Effects of water pH on the uptake and elimination of the piscicide, 3-trifluoromethyl-4-nitrophenol (TFM), by larval sea lamprey. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol. 200, 9–16. Lewis, M.A., 1992. The effects of mixtures and other environmental modifying factors on the toxicities of surfactants to fresh-water and marine life. Water Res. 26 (8), 1013–1023. Marques, D.D., Costa, P.G., Souza, G.M., Cardozo, J.G., Barcarolli, I.F., Bianchini, A., 2019. Selection of biochemical and physiological parameters in the croaker Micropogonias furnieri as biomarkers of chemical contamination in estuaries using a generalized additive model (GAM). Sci. Total Environ. 647, 1456–1467. NOM, 1998. Norma Oficial Mexicana NOM-002-ECOL-1996. In: Secretarıá de Medio Ambiente, Recursos Naturales y. Pesca, Mexico. Nugegoda, D., Kibria, G., 2013. Water quality guidelines for the protection of aquatic ecosystems. In: Ferard, J.-F., Blaise, C. (Eds.), Encyclopedia of Aquatic Ecotoxicology. Springer Science + Business Media, Dordrecht. https://doi.org/10.1007/978-94- 007-5704-2. 24 Bernardo Baldisserotto

RSA-DEA (Republic of South Africa, Department of Environmental Affairs), 2018. South African Water Quality Guidelines for Coastal Marine Waters - Natural Environment and Mariculture Use. Cape Town. Smith, K.S., Balistrieri, L.S., Todd, A.S., 2015. Using biotic ligand models to predict metal toxicity in mineralized systems. Appl. Geochem. 57, 55–72. Su, J., Ji, D.F., Lin, M., Chen, Y.Q., Sun, Y.Y., Huo, S.L., Zhu, J.C., Xi, B.D., 2017. Devel- oping surface water quality standards in China. Resour. Conserv. Recycl. 117, 294–303. Tan, Q.G., Lu, S.H., Chen, R., Peng, J.H., 2019. Making acute tests more ecologically relevant: cadmium bioaccumulation and toxicity in an estuarine clam under various salinities modeled in a toxicokinetic-toxicodynamic framework. Environ. Sci. Technol. 53 (5), 2873–2880. USEPA (United States Environmental Protection Agency), 2007. Aquatic Life Ambient Freshwater Quality Criteria—Copper: 2007 Revision. U.S. Environmental Protection Agency; 2007204. USEPA (United States Environmental Protection Agency), 2019. National Recommended Water Quality Criteria—Aquatic Life Criteria Table. https://www.epa.gov/wqc/ national-recommended-water-quality-criteria-aquatic-life-criteria-table#table (Accessed April 26th, 2019). Zhang, L., Wang, W.X., 2007. Waterborne cadmium and zinc uptake in a euryhaline teleost Acanthopagrus schlegeli acclimated to different salinities. Aquat. Toxicol. 84 (2), 173–181. Zhao, X.L., Wang, H., Tang, Z., Zhao, T.H., Qin, N., Li, H.X., Wu, F.C., Giesy, J.P., 2018. Amendment of water quality standards in China: viewpoint on strategic considerations. Environ. Sci. Pollut. Res. 25 (4), 3078–3092. CHAPTER THREE

Antifoulants and disinfectants

Samantha Eslava Martinsa,b and Camila de Martinez Gaspar Martinsb aNorwegian Institute for Water Research (NIVA), Ecotoxicology and Risk Assessment Section, Oslo, Norway bUniversidade Federal do Rio Grande—FURG, Instituto de Ci^encias Biolo´gicas, Rio Grande/RS, Brazil

3.1 Overview In the past 50years, the primary aquatic food supply for humans changed from wild-caught to production in farms. Aquaculture increased not only in the levels of production, but in the types of species cultivated, driven by high demand in a globalized environment. In 2014, the contribution of aquaculture to the production of fish for human consumption outweighed the acquisition of fish through wild fishing. Currently, the aquaculture sector plays an important role in food security and livelihood, being the source of income and social development. Fish and fishery goods represent one of the most-traded segments of the world food sector, with a high international trade competition. For several countries and for numerous coastal and riverine regions, production of fish and fishery products is essential to their economies, accounting for more than 40% of the total value of traded commodities in some countries, mainly islands. This represents a value >9% of global agricultural exports and >1% of world commodity trade (FAO, 2016). This impressive growth of the aquaculture sector was possible through the development of technologies that confer improved productivity, allowing the production of aquatic animals in line with the increasing demand for seafood as a source of animal protein. However, major constraints impinging aquaculture include the biofouling on aquaculture equipment and infrastructure, and the losses from diseases. Biofouling is of major concern in both shipping and aquaculture industries. In the shipping industry, biofouling brings economic constraints because fouled organisms result in the deformation of the shape of hulls and hence the hydrodynamics of the boat, leading to an increase in the fric- tional drag and thereby fuel consumption (Abbott et al., 2000). Biofouling is a problem especially for sea cage-based finfish aquaculture where impacts are well documented. In particular, a major concern with the fouling of cages is the occlusion of netting mesh, which restricts water flow resulting in

Aquaculture Toxicology © 2021 Elsevier Inc. 25 https://doi.org/10.1016/B978-0-12-821337-7.00005-0 All rights reserved. 26 Samantha Eslava Martins and Camila de Martinez Gaspar Martins reduced water quality (De Nys and Guenther, 2009). Another major challenge is the development of effective disease treatment and prevention. Studies have shown that caged salmon may come in contact with nemato- cysts of fouling cnidarians when nets are cleaned and hydroids are released into the water (Bloecher et al., 2018; Floerl et al., 2016). Hydroid’s nema- tocysts lead to irritation and pathological damage to salmon gills and may facilitate the spread of the amoebic gill disease (AGD), an increasing threat to Atlantic salmon farming (Bloecher et al., 2018). In this sense, in order to maintain optimum culture conditions and ensure aquaculture as a profitable industry, proper husbandry techniques are mandatory, requiring that antifouling biocides are used to minimize biofouling. About 90% of the aquaculture occurs in developing countries, and the production losses due to disease reach up to 6 billion dollars per annum (Assefa and Abunna, 2018; Leung and Bates, 2013). The proliferation of opportunistic pathogens (bacteria, virus, fungi, or protozoa) is driven by the high density of fish stocks and the feed wastes; thus, the use of quick and effective antipathogens is necessary (Castillo-Rodal et al., 2012; Twiddy et al., 1995). Therefore, disinfection is commonly used in aquaculture facilities.

3.2 Definitions and uses 3.2.1 Antifoulants When an object is placed into water, microorganisms followed by algae and animals start to adhere to it. This process is termed biofouling, defined as the settlement and attachment of organisms on the external surfaces of submerged or semisubmerged objects (Lewis, 1998). Antifouling technologies have developed to protect structures against biofouling. Antifouling coatings have been used for many centuries. The Phoeni- cians and Carthaginians were the first to use pitch and likely copper sheeting to prevent the settlement of organisms on the bottom of wooden ships (Hellio and Yebra, 2009). From the late 18th century, the use of metals in coatings increased successfully and these compounds, particularly copper, are still incorporated into modern coatings (Dafforn et al., 2011). The efficacy of metals as antifoulants can be enhanced by co-biocides (or booster biocides) in the painting formulations. Such co-biocides found their market from the early 1960s when organotin-based paints were introduced as marine antifoulants and believed to be the solution to preventing biofouling Antifoulants and disinfectants 27

(Yebra et al., 2004). However, severe impacts on the marine environment occurred following the introduction of tributyltin (TBT), including the well-known phenomenon of imposex, which decimated the marine gastropod Nucella lapillus in coastal areas of Southwest England (Gibbs and Bryan, 1996). The severity of adverse ecological effects of organotin-based antifoulants culminated in the gradual restriction until the global ban of TBT as an active ingredient in antifouling paints, in 2008 (AkzoNobel, 2000; Dafforn et al., 2011). Subsequently, organotin-free technologies flooded the paint market, turning paint formulations into a mixture of inorganic biocides (typically cuprous oxide) and one or more organic/organometallic co- biocides (Hellio and Yebra, 2009), such as Irgarol 1051, Diuron, DCOIT (Seanine), chlorothalonil, dichlofluanid, TCMTB, thiram, zinc pyrithione (ZnPT), and copper pyrithione (CuPT) (Castro et al., 2011; Hellio and Yebra, 2009). Such co-biocides are not expected to leach into the water bodies to trigger toxic concentrations to nontarget species (Hellio and Yebra, 2009). However, it is already known that several organisms may be very sensitive to such antifoulants, particularly during the early life stages, including farmed aquatic animals (Martins et al., 2018).

3.2.2 Disinfectants Disinfectants are used in intensive cultures, particularly in finfish and shrimp hatcheries and grow-out ponds in routine practices to disinfect culture facilities and equipment, to maintain hygiene throughout the production cycle, and often to treat bacterial disease outbreaks. A disinfectant is an agent that destroys infection-producing organisms. In this sense, disinfectants can act on microorganisms in two distinct ways: (1) through growth inhibition (bacteriostasis and fungistasis) or (2) through lethal action (bactericidal, fungicidal, or virucidal effects). The particular reason for disinfection will determine the disinfection strategy used and how it should be applied. General practices of the disinfection of aquaculture systems involve the application of chemical compounds in sufficient concentrations, and for sufficient periods, to kill or reduce the pathogenic organisms, mainly virus and fungus, avoiding the loss of production. Essential disinfection protocols include: (1) removal of all aquatic animals (both dead and alive) from the facility; (2) a deep cleaning and washing to eliminate remaining organic matter adhered to the surfaces; (3) the application of disinfectants on equipment and installations; (4) a final wash; and (5) the neutralization using chemical products, if needed. The disposal of diseased animals should not be done in 28 Samantha Eslava Martins and Camila de Martinez Gaspar Martins

Table 3.1 Efficacy and characteristics of commonly used disinfectants. Target organism Bacterial Spore-forming Disinfectant Efficacy Virus Bacteria spore Fungi protozoan Formaldehyde High + + + + + H2O2 High + + VA + + PAA High + + + + + Chlorine High + + VA + LA Iodophors Intermediate + + +LA QACs Low VA VA + Ozone High + + VA + LA

QACs, quaternary ammonium compounds; PAA, peracetic acid; +, effective; , nonrecommended; VA, variable activity; LA, limited activity. Information on specific efficacies followed Scarfe et al. (2006) and AQUAVETPLAN (2008). receiving waters due to the risk of contamination of wild populations or neighboring farms that use the same water supply. Therefore, the animals must be euthanized and collected by a company designed to treat or store hazardous waste. The disinfectants must be stored in such a way that they will not pose direct or indirect dangers to animal or human health and the environment (OIE - World Organisation for Animal Health, 2009). Among the most used disinfectants are formaldehyde, hydrogen peroxide, chlorine, potassium permanganate, isopropyl alcohol, iodophors, peracetic acid and quaternary ammonium compounds (Bowker et al., 2014; Scarfe et al., 2006). In general, they are characterized by high solubility and low persistence in the aquatic environment. The efficacy and characteristics of the commonly used disinfectants are summarized in Table 3.1 (Kohn et al., 2017; Rico et al., 2012).

3.3 Mode of action 3.3.1 Antifoulants The mode of action of antifouling biocides depends on the target organisms for which the formulation was designed for. Antifouling herbicides such as Irgarol 1051 and Diuron act by inhibiting the transport of electrons during photosystem II (Hall et al., 1999), affecting photosynthetic organisms. Anti- fouling fungicides may act through different mechanisms. Chlorothalonil acts through the inhibition of glycolysis or depleting glutathione (Caux et al., 1996); despite being designed to act as a fungicide, the presence of multiple reactive electrophilic centers makes chlorothalonil extremely toxic Antifoulants and disinfectants 29 to nontarget aquatic organisms (Castro et al., 2011), causing effects in animals and plants. Dichlofluanid is a potent inhibitor of fungal spore germination (PPDB, 2007–2017), but its degradation products may play a major role in toxicity since dichlofluanid quickly undergoes hydrolysis in water (Schouten et al., 2005). Thiram inhibits spore germination and mycelial growth (PPDB, 2007–2017) as well as nontarget species for being a multisite inhibitor (KEMI, 2015). The organometallic antifouling biocides such as zinc pyrithione (ZnPT) and copper pyrithione (CuPT) show microbiocidal activity. There is a lack of information on the mode of action of pyrithione salts, but it has been reported that ZnPT and CuPT catalyze the electroneutral exchange of H+ and other ions with K+ across cell membranes, disrupting the proton motive force in target organisms. As a consequence, the transport of nutrients across membranes is impaired, leading organisms to starvation and even- tual death (KEMI, 2014). Some antifoulants act as broad-spectrum biocides with efficacy against either weeds or fungi (Ferna´ndez-Alba et al., 2002). For instance, the antifouling co-biocide DCOIT (4,5-dichloro-2-octyl-1,2-thiazol-3-one) reacts with the proteins of fouling specimens when they find the painting surface, breaking metabolic processes, and hence preventing the attachment of the organism to solid surfaces. Another important broad-spectrum co- biocide is TCMTB (1,3-benzothiazol-2-ylsulfanylmethyl thiocyanate) that inhibits the mitochondrial electron transport chain (Ferna´ndez-Alba et al., 2002) in a wide range of nontarget organisms. In addition to herbicides, fungicides, microbiocides, and broad-spectrum biocides, some antifoulants are categorized as “emerging compounds” because they have a relatively new use as an antifouling biocide encompassing mainly regional markets. Among them, tralopyril, medetomidine, and TPBP (pyridine-triphenylborane) should be highlighted as synthetic biocides. Tralopyril is used in coatings to enhance the antifouling efficacy of copper-free antifouling paintings by uncoupling mitochondrial oxidative phosphorylation (EU, 2014; International, 2014). Medetomidine acts through the activation of analogous octopamine leading to an anti-settling effect (EU, 2015) and is designed to protect against shell-building marine organisms. TPBP has been largely used in Japan (Mochida et al., 2012), but its mode of action is still unknown (Wendt et al., 2016). Capsaicin is a natural co-biocide extracted from chili peppers, which acts on the nervous system and also disrupts metabolism and damages membranes (Gervais et al., 2008). The capsaicin derivative nonivamide acts the same way. 30 Samantha Eslava Martins and Camila de Martinez Gaspar Martins

3.3.2 Disinfectants Formaldehyde is one of the most applied disinfectants in intensive aquaculture, being frequently used to control parasitic infections on fish and crustacean surfaces, and it is also used for treatment against water mold (fungus) on fish eggs (Boyd and Massaut, 1999). Although its use is encouraged in hatcheries to control the presence of “fungus,” it is not recommended for ponds because it kills algae present in pond water, reducing the production of oxygen by phytoplankton and augmenting the organic matter and decomposition (Francis-Floyd and Pouder, 1996). Formaldehyde when diluted in water form is known as formalin. Formalin may be applied as a prophylactic measure or for therapeutic purposes and is extremely effective against most protozoan parasites (Ichthyophthirius spp., Costia spp., Epistylis spp., Chilodonella spp., Scyphidia sp., and Trichodina spp.) and monogenetic trematodes (Cleidodiscus spp., Gyrodactylus spp., and Dactylogyrus spp.) (FDA, 1995; Francis-Floyd and Pouder, 1996; Shao, 2001). Due to its electrophilic character, formalin can react with functional groups of several biological macromolecules, such as proteins, DNA and RNA, polysaccharides, and glycoproteins (Leal et al., 2018). Hydrogen peroxide (H2O2) and chlorine are potent oxidizing agents not only used as broad-spectrum disinfectants in finfish and shellfish produc- tions, but can also be used to treat fungal infections and as pesticides during the pond preparation between production cycles (Rico et al., 2012). Chlo- rine gas and powdered forms such as calcium hypochlorite and sodium hypochlorite are used to disinfect water supplies in fish and shrimp hatcheries. They can be used in ponds after physical sediment removal, also between production cycles (AQUAVETPLAN, 2008). The chlorine and H2O2 oxidize electrons from susceptible chemical groups and become themselves reduced in the process. Oxidizing agents are usually low-molecular-weight compounds ready to easily pass through cell walls/membranes and then react with internal cellular components disrupting them, and leading to apoptotic and necrotic cell death (Finnegan et al., 2010). At cellular level, low levels of oxidation can be reversible and prokaryotic organisms have developed many defenses against these effects. Recently, Pedersen et al. (2019) showed a relation between H2O2 decomposition and particle-associated bacterial activity, suggesting that the quantification of H2O2 in water samples can be used as a rapid and feasible indicator of microbial activity in fresh and saltwater aquaculture systems, ranging from pond farming to intensive recirculating systems. Antifoulants and disinfectants 31

Iodophors are polyvinylpyrrolidone-iodide-iodine complexes in aqueous solution, used as disinfectants for nonhardened fish eggs to prevent spreading pathogens from the broodstock fish to the offspring (Lahnsteiner and Kletzl, 2016). Iodine acts by decreasing the oxygen requirements of aerobic microorganisms. It acts on microorganisms’ respiratory chain by blocking the transport of electrons through electrophilic reactions with the enzymes of the respiratory chain (Maris, 1995). The optimum concentration is >200mgL 1 free iodine with a contact time of 2min and 100mgL 1 free iodine for cleaned and dried equipment. In a nonfood- contact application, the concentration may rise to 500–800mgL 1. Iodo- phors should be used with caution because they are highly toxic to fish (Boyd and Massaut, 1999). Ammonia treatment in sanitation is a relatively new approach. In contrast to oxidants, ammonia is not consumed during the treatment. The ammonia effect on sanitation is sustained for prolonged periods, while also reducing the risk for regrowth (Kohn et al., 2017). Quaternary ammonium compounds (QACs) impair membrane permeability by irreversibly binding to phospholipids and proteins of the membrane. One of the commonly used products is benzalkonium chloride, applied to inhibit bacterial growth and the development of mucus in the gills of salmon (Burka et al., 1997), thereby allowing an adequate absorption of oxygen. The capability of the bacterial cell to absorb such molecules influences sensitivity. These compounds are used in finfish and shellfish production to treat bacterial, protozoan, and monogenean infections and as fungicides in shrimp hatcheries (Rico et al., 2012). Ozone is a powerful oxidizing agent widely used in aquaculture for disinfection of pathogens and for water quality enhancements since it oxidizes organic wastes and nitrite. The oxidation of pathogens and other material by ozone occurs extremely rapidly, through reactions of molecular ozone with oxidizable compounds, and also due to the reaction of the oxidizable compounds with reactive oxygen species (ROS) formed by ozone decomposition in water (AQUAVETPLAN, 2008; Gonc¸alvez and Gagnon, 2011; Scarfe et al., 2006; Spiliotopoulou et al., 2018; Summerfelt and Hochheimer, 1997; Tyrell et al., 1995). As an oxidative disinfectant, it is consumed by organic matter; thus, initial ozone concentration decays faster in waters with higher concentrations of organic matter as wastewater. For this reason, it is important to apply an ozone dose that remains in the system for enough time for the oxidation demands, whether inactivation of pathogens or degradation of the organic matter present, without affecting farmed 32 Samantha Eslava Martins and Camila de Martinez Gaspar Martins species. Nevertheless, monitoring ozone performance is still a challenge (Spiliotopoulou et al., 2018). Ozonation is usually used in recirculating culture systems that consume less water per kg of fish produced and allow solid removal and effluent treatment (Gonc¸alvez and Gagnon, 2011). The literature reports a wide range of ozone dosages in recirculating systems, according to the number of animals and feed ratio (Bullock et al., 1997; Summerfelt et al., 2009; Summerfelt and Hochheimer, 1997). A drawback of ozone use is that its application is energy-consuming and requires a trained operator. Even though ozone is more reactive than chlorine in demand-free solutions toward all microorganisms, it has shown to be more efficient against viruses, but less efficient against bacteria than chlorine (Tyrell et al., 1995), although both can inactivate bacterial spores, depending on the conditions of the application and the origin of the pathogen (Broadwater et al., 1973; Rose et al., 2005; Tyrell et al., 1995). Nanomaterials are gaining space as disinfectants for aquaculture purposes. For example, they have been used to improve water quality in shrimp aquaculture, reducing the rate of water exchange, increasing shrimp survival rate and yield (Wen et al., 2003). Silver (Ag) nanoparticles (NPs) (nAg) are the most investigated antibacterial compounds. The nAg was used for the treatment of fungal infections in rainbow trout eggs showing an inhibitory effect on fungi growth ( Johari et al., 2015). Dananjaya et al. (2016) investigated the antibacterial function of chitosan-Ag nanocomposites (CagNCs) against the fish pathogen Aliivibrio salmonicida. The CagNCs inhibited A. salmonicida growth with a minimum inhibitory concentration (MIC) and minimum bactericidal concentration (MBC) at 50 and 100mgL 1, respectively. Similarly, nZnO exhibited antibacterial activity disrupting bacterial cell membrane integrity, reducing cell surface hydrophobicity and down- regulating the transcription of oxidative stress-resistance genes (Pati et al., 2014). Furthermore, Muhling€ et al. (2009) showed that nTiO2 and nAg reduced the buildup of bacteria in estuarine water (). Another disinfectant that has been recognized to be suitable for aquaculture is peracetic acid (PAA) which is considered an alternative sanitizer to formaldehyde. PAA is a highly reactive peroxygen with widespread antimicrobial effects (Liu et al., 2015, 2016; Pedersen et al., 2009, 2013). It degrades entirely within several hours after application, resulting in trace concentration residuals that are not readily measured (Pedersen et al., 2009). The test kits are not sensitive enough to analyze PAA at low ranges (below 0.2ppm). Only a few studies have described analytical measurement of the commercial PAA compound (Pedersen et al., 2013). In this sense, Antifoulants and disinfectants 33

Sudova´ et al. (2010) reported a reduction of the protozoan Ichthyophthirius multifiliis infestation in gills, skin, and fin of carp after 4days of continuous application of PAA at 1.0mgL 1; however, the authors discussed uncertainties in PAA analysis in the work. When PAA is used with ozone, care is needed to minimize the oxidative stress due to residues, since both compounds are highly reactive and in particular the therapeutic index of PAA is small (Pedersen et al., 2013). PAA commercial products are acid mixtures of PAA and hydrogen peroxide (H2O2), acetic acid, H2O, and stabilizers to preserve the formulation (Liu et al., 2015). Usually, PAA concentration is fixed in formulations and H2O2 is variable. The effective concentration of PAA against various pathogens is less than 2mgL 1 (Pedersen et al., 2013). In contrast, the H2O2 needs a much higher concentration (over 20mgL 1) to achieve successful disinfection (Schmidt et al., 2006). Some studies have shown that different ratios of PAA:H2O2 lead to different toxicities. They show an additive effect of H2O2, since the toxicity of the mixture of PAA and H2O2 augmented with increasing in H2O2 concentrations (Liu et al., 2015, 2016; Pedersen et al., 2013). Pedersen et al. (2013) related the time of exposure to PAA and H2O2 and potential toxicity to the microalgae Tetraselmis chuii. The authors reported a rapid degradation of PAA and slower degradation of H2O2, indicating that the toxicity of PAA formulations to T. chuii was due to the effects of the two compounds initially but later only by the H2O2 as it has a longer half-life. As peroxygens, PAA and H2O2 are exogenous sources of ROS that cross the cell membranes, increasing the endogenous ROS concentration. ROS in excess induces apoptosis in algae (Segovia and Berges, 2009; Voigt and Woestemeyer, 2015). Algae have antioxidant defenses to resist against an increase in ROS inter- cellular concentration, but in the face of high doses of pro-oxidants, these defenses may not be enough to keep the algal culture alive. Therefore, the use of PAA formulation in integrated multi-trophic aquaculture (IMTA) systems may be of concern, since in this system, microalgae contribute to the nitrification process of the biofilter (Liu et al., 2016).

3.4 Ecotoxicity and biological effects As defined by the French toxicologist Rene Truhaut in 1969, ecotoxicology is “the branch of toxicology concerned with the study of toxic effects, caused by natural or synthetic pollutants, to the constituents of ecosystems, animal (including human), vegetable, and microbial, in an integral context.” Hence, ecotoxicological 34 Samantha Eslava Martins and Camila de Martinez Gaspar Martins tools such as toxicity tests and the use of biomarkers have been successfully employed to determine toxic threshold concentrations, deleterious effects, and modes of action of toxic compounds, including biocidal products such as antifoulants and disinfectants.

3.4.1 Antifoulants Most antifouling biocides in use are copper-based, usually in the form of copper oxide. Zinc is also frequently used in antifouling paint formulations. Hence, high copper levels have been reported in sediments close to fish farms (Simpson et al., 2013). Similarly, the broad application of antifoulants can act as a source of metals and bioaccumulation of these compounds in cultured fish, representing not only deleterious effects to the exposed fish (Nikolaou et al., 2014) but also a potential risk to human health. In this sense, the use of copper and zinc is not the best alternative for biofouling control in the food producing industries. These metals have been reported to be toxic to aquatic organisms by the European Union and other regulatory boards worldwide (Nikolaou et al., 2014). Therefore, the use of TBT-free organic co-biocides has gained attention from the last few decades. According to Guardiola et al. (2012), the main antifouling biocides used in aquaculture are chlorothalonil, copper pyrithione (CuPT), dichlofluanid, DCOIT (Sea-nine 211), Diuron, Irgarol 1051, TCMS pyridine (2,3,5,6- tetrachloro-4-methylsulfonylpyridine), zinc pyrithione (ZnPT), and zineb. However, these and other co-biocides may also be toxic to nontarget organism. Thus, many studies have focused on determining acute and chronic toxicity of antifoulants to aquatic organisms, including farmed mollusks, crustaceans, and fish species (see recent reviews by Amara et al., 2018; Martins et al., 2018). For the herbicide Irgarol 1051, acute toxicity under estuarine and marine conditions ranges from 1000μgL 1 to the larvae of the killifish Oryzias melastigma to 3730μgL 1 to the adults of the eastern mud snail Tritia obsoleta. Chronic toxicity ranges from 330μgL 1 to the sheepshead minnow Cyprinodon variegatus to 1250μgL 1 to adults of the mummichog Fundulus heteroclitus (Table 3.2). For the herbicide Diuron, acute toxicity ranges from 890μgL 1 to C. variegatus to 7800μgL 1 in larvae of O. melastigma. Diuron chronic toxicity ranges from 0.004μgL 1 in the embryo of Pacific oysters Crassostrea gigas to 1000μgL 1 in larvae of the common prawn Palaemon serratus, showing high Diuron toxicity to the commonly farmed oyster C. gigas in the early stages (Table 3.2). Table 3.2 Toxicity (μgL1 ) of antifouling herbicides (Irgarol 1051 and diuron) to marine mollusk, decapod, and fish species. Table 3.2 Toxicity (μgL1 ) of antifouling herbicides (Irgarol 1051 and diuron) to marine mollusk, decapod, and fish species—cont’d Antifoulants and disinfectants 37

For the fungicide chlorothalonil, marine animals exhibited acute toxicity ranging from 3.6μgL 1 in embryos of C. gigas to 34,780μgL 1 in adult clams Mya arenaria (sand gaper). Chronic toxicity ranged from 4.57μgL 1 in embryos of the common blue mussel Mytilus edulis to 31.3μgL 1 in larvae of daggerblade grass shrimps Palaemonetes pugio (table 12.3). The fungicide dichlofluanid presented acute toxicity ranging from 15.6μgL 1 in the Euro- pean bass Dicentrarchus labrax to 1000μgL 1 in adult brown shrimps Penaeus aztecus. Reported chronic toxicity ranged from 10.6μgL 1 in D. labrax to 98.7μgL 1 in embryos C. virginica (Table 3.3). For the fungicide thiram, acute toxicity from 4.7μgL 1 in embryos C. gigas to 540μgL 1 in C. variegatus has been reported. Only one study of thiram chronic toxicity in marine conditions was found, in which the no-effect concentration (96-h NOEC) of 90μgL 1 was verified in C. variegatus (Table 3.3). Table 3.4 shows the reported ecotoxicity of the broad-spectrum antifoulants DCOIT and TCMTB. DCOIT acute toxicity in the three main farmed aquatic animal groups, bivalves, crustacean and fish, ranges from 11μgL 1 in embryos of M. edulis to 1700μgL 1 in adult decapods Uca pugilator (Atlantic sand fiddler) Only two data for chronic toxicity were found, in which DCOIT is slightly more toxic to early stages of fish C. variegatus (35 d-LOEC¼6μgL 1, LOEC¼lowest observed effect concentration) than to embryos M. edulis (48-h NOEC¼7.1μgL 1). For TCMTB, there are a few toxicity data available. Mun˜oz et al. (2010) exposed the Northern quahog clam Mercenaria mercenaria to TCMTB for 48h and found an effective concentration to 50% of the tested population (EC50) of 13.9μgL 1, while acute and chronic toxicity for C. variegatus was reported as 60μgL 1 (USEPA, 2006) and 36μgL 1 (ECOTOX, 2000-2017), respectively. For the antifouling microbiocide zinc pyrithione (ZnPT), acute toxicity ranged from 2.54μgL 1 in embryos of M. edulis to 4797.2μgL 1 in the adults of M. galloprovincialis. Chronic toxicity for C. variegatus was 200μgL 1 and this was the only data found (Table 3.5). For copper pyrithione (CuPT), acute toxicity ranged from 2.5μgL 1 to the decapod Heptacarpus futilirostris to 43.6μgL 1 to the kuruma prawn Penaeus japonicus (Table 3.5). Data on chronic toxicity were not found. Some emerging compounds have also been regarded as TBT-free alternative co-biocides. Their uses are usually limited to local markets or small scale and so are toxicity tests. Medetomidine showed acute toxicity of 2500μgL 1 to embryos C. gigas and chronic toxicity of 32μgL 1 to fish C. variegatus. For triphenylborane pyridine (TPBP), only acute toxicity data Table 3.3 Toxicity (μgL1 ) of antifouling fungicides (chlorothalonil, dichlofluanid, and thiram) to marine mollusk, decapod, and fish species. Table 3.4 Toxicity (μgL1 ) of broad-spectrum antifouling biocides (DCOIT and TCMTB) to marine mollusk, decapod, and fish species. Table 3.5 Toxicity (μgL1 ) of antifouling microbiocides (ZnPT and CuPT) to marine mollusk, decapod, and fish species. Antifoulants and disinfectants 41

were found, ranging from 6.3μgL 1 in embryos C. gigas to 367μgL 1 in larvae of F. heteroclitus. For tralopyril and capsaicin, only one acute toxicity data were found, in which 48-h EC50 of 3.1μgL 1 and 3868μgL 1 were reported for embryos of mussels M. galloprovincialis, respectively (Table 3.6). Overall, toxicity thresholds for marine mollusk, crustaceans, and fish species are above 1μgL 1 (Fig. 3.1). Exceptions were found for the freshwater fish Oryzias latipes (medaka), whose eggs were very sensitive to the fungicide chlorothalonil (NOEC¼0.06μgL 1) and to the marine oyster C. gigas, for which the herbicide Diuron was highly toxic to the embryo (24-h NOEC¼0.004μgL 1). Nevertheless, the continuous exposure to low concentrations of antifouling co-biocides may lead to deleterious biological effects (Chen and Lam, 2017). Antifouling biocides were deliberately designed to repel or kill target organisms that foul on submerged structures such as ship hulls and aquaculture nets. However, their modes of action may lead to the toxic effects on nontarget organisms including cultured animals. To a lesser extent, some studies have reported adverse effects of antifouling biocides to animals of interest for aquaculture. To cite a few, Guerreiro et al. (2017) exposed adult brown mussels Perna perna to chlorothalonil and found out that chlorothalonil increases phagocytosis and hemocyte adhesion index in concentrations as low as 0.1μgL 1, concluding that this antifoulant modulates the immune system of mussels. In another study, Akcha et al. (2016) observed that Diuron, at environmentally relevant concentration, affected the expressions of genes involved in functions such as DNA methylation, gene transcription regulation, and cytokinesis, during gametogenesis in the marine oyster C. gigas and stated that short exposure to Diuron may be harmful not only to oyster genitors but also to their offspring. Moreover, Diuron and Irgarol were shown to be toxic to the antioxidant defense system in C. gigas, in which realistic Diu- ron concentrations (1μgL 1) reduced Na+/K+-ATPase activity and both Diuron and Irgarol reduced AchE mRNA expression and strongly induced the Hsp70 (heat shock protein) family (Park et al., 2016). In fish, two commercial metal-based antifouling paints were shown to change enzyme activity, as well as cause gill abnormalities, genotoxicity, and were hepatotoxic to the catfish Clarias gariepinus (George et al., 2017). In O. latipes, decreased hatchability, abnormalities of respiration and swimming behavior, were reported after a short exposure to zinc and copper pyrithiones, under freshwater conditions (Ohji and Harino, 2017). Zinc pyrithione was also shown to be toxic to the silver crucian carp Table 3.6 Toxicity (μgL1 ) of emerging antifoulants to marine mollusk, decapod, and fish species. Antifoulants and disinfectants 43

Fig. 3.1 Antifoulant toxicity to marine mollusk, crustacean, and fish species. Mollusk and crustaceans are divided into zooplankton and benthic stages. Modified from Martins, S.E., Fillmann, G., Lillicrap, A., Thomas, K.V., 2018. Review: ecotoxicity of organic and organo-metallic antifouling co-biocides and implications for environmental hazard and risk assessments in aquatic ecosystems. Biofouling 34, 34–52.

Carassius auratus gibelio, in which P-glycoprotein (P-gp) was found to be induced in liver and kidney, and this effect was affected by salinity (Ren et al., 2017). Conversely, some antifouling formulations have been developed to increase safety in aquaculture and prevent adverse effects on caged organisms. For instance, the ascidian metamorphosis inhibitor polygodial was tested for effects on mussels Perna canaliculus and did not affect biochemical parameters such as antioxidant capability and detoxification properties on gills of adult mussels (Cahil et al., 2013).

3.4.2 Disinfectants The risk of disinfection to the aquatic biota is dependent not only on how much disinfectant is being used but also on where it is being released. The composition of disinfectants is also of significant concern and some 44 Samantha Eslava Martins and Camila de Martinez Gaspar Martins disinfectant formulations contain surfactants that may not present in label but that are known to be endocrine disruptors (Burridge et al., 2010). Guidelines for best management practices in aquaculture facilities recognize that the use of disinfectants in the aquaculture waters is toxic and should be neutralized prior to discharge or reuse. However, most of the studies are focused on disinfectant efficacy to aquaculture and a few address the disinfectant toxicity to nontarget organisms. Aquaculture effluents are often directly discharged into the surrounding environment, and the water from the same environment is then pumped back into the aquaculture ponds to breed new stocks. Thus, it is important to determine the optimal disinfectant required as well as its dose of application to balance both antimicrobial efficiency and ecological risks. Regarding the toxicity of disinfectants to fish, Kim et al. (2008) estimated the median lethal concentration (LC50) values of 10 disinfectants to determine toxicity thresholds to flounder (Paralichthys olivaceus), rockfish (Sebastes pachycephalus), and black seabream (Acanthopagrus schlegelii) within 24h of exposure. The authors observed that toxic concentrations depended on each chemical as well as by species, but they suggested that QACs and sodium hypochlorite are more toxic than chlorine and formaldehyde that, in turn, have greater toxicity than H2O2 to the fish. Leal et al. (2018) reported information about formalin (36%–38% formaldehyde) toxicity to some aquatic species produced in aquaculture, where 1 the results of formalin LC50 ranged from 15μLL for the striped bass Morone saxatilis (30–52mm) after 48 or 96h of exposure to 1896μLL 1 for the Atlantic salmon Salmo salar (0.60g) exposed for 3h to formalin. The authors concluded that formalin effects are strongly dependent on several factors such as the fish species and their size/stage of life, and exposure time to formalin and its concentration. It is worth noting that LC50 reports mostly refer to 24 to 96h of exposure, which are longer than the standard bath treatment times recommended. Nevertheless, this should not be ignored, because if formalin is not completely removed from the system after the disinfection, the aquatic species will remain exposed to it in the tank. Then, it may cause toxicity to the caged animals. In addition, the toxicity of formalin to aquatic species is also affected by water chemistry, such as pH, hardness, and the presence of some natural constituents in waters (Bills et al., 1977, 1993; Chinabut et al., 1988; Meinelt et al., 2005). For example, at higher pH toxicity of formalin increased to rainbow trout Oncorhynchus mykiss, channel catfish Ictalurus punctatus (APUD Leal et al., 2018; Bills et al., 1977), and common carp Cyprinus carpio (Chinabut et al., 1988). Antifoulants and disinfectants 45

Some studies have reported possible consequences of fish exposure to residual formalin/formaldehyde to rainbow trout (O. mykiss)(WHO, 1989; Williams and Wooten, 1981), seabream (Sparus aurata), sea bass (D. labrax)(Yildiz and Ergonul, 2010), and olive flounder (Paralichthys olivaceus)(Jung et al., 2003). The most common adverse effect is the perma- nent damage in gills (Shepherd and Bromage, 2001; Williams and Wooten, 1981). However, there are also some studies reporting no significant or limited significant negative effects on fish when exposed to recommended dosages of formalin (Chinabut et al., 1988; Speare et al., 1997). One positive aspect regarding the use of this disinfectant is that formaldehyde is highly soluble in water and its bioaccumulation in aquatic organisms is not expected (WHO, 1989). Concerning the disinfectant PAA, Straus et al. (2018) reported different mortality-based toxicity endpoints to 12 species of fingerling fish: fathead minnow (Pimephales promelas), black nose crappie (Pomoxis nigromaculatus), bluegill (Lepomis macrochirus), blue tilapia (Oreochromis aureus), channel catfish (I. punctatus), golden shiner (Notemigonus crysoleucas), goldfish (Carassius auratus), grass carp (Ctenopharyngodon idella), largemouth bass (Micropterus sal- moides), rainbow trout, sunshine bass (Morone chrysops), and walleye (Sander vitreus). The lethal concentration to 50% of the tested population (LC50) using nominal PAA concentrations ranged from 2.8 to 9.3mgL 1 after 24-h exposure. The fathead minnow was very sensitive and blue tilapia was very tolerant to PAA exposure; The NOEC ranged from 1.9 to 5.8mgL 1 PAA, which is considered as a safe range for culturists. Water chemistry also influences PAA toxicity; there was an increase in toxicity as alkalinity and hardness decreased, while the addition of dissolved organic matter had no effect on PAA toxicity. Aquaculture disinfectants can be moderately to highly toxic to plankton and macroinvertebrate species with acute toxicity ranging between 1 and 100,000μgL 1 (Rico et al., 2012). Formaldehyde might be harmful to phytoplankton and crustaceans at recommended doses to treat bacterial infesta- tions in fish (Tonguthai, 2000). Formalin released into small, stagnant, or slow-moving water bodies may cause the death of phytoplankton and zooplankton population as a result of decreasing oxygen levels (FDA, 1995). The oxidation of formalin to carbon dioxide and water naturally consumes oxygen, which may constitute a severe problem in both the aquaculture systems and natural water bodies. Besides, chlorine disinfectants (e.g., calcium or sodium hypochlorite) react with organic matter, giving rise to significant concentrations of organic chlorine compounds such as halogenated 46 Samantha Eslava Martins and Camila de Martinez Gaspar Martins hydrocarbons, which are highly toxic to aquatic life and are persistent in the environment (Emmanuel et al., 2004; Sanawar et al., 2017).

3.5 Ecological risks and regulation The sections mentioned previously showed that the use of biocidal products, such as organic and organometallic antifouling biocides and disinfectants, causes deleterious effects in nontarget organisms, including farmed animals for aquaculture. Therefore, many nations are concerned with esta- blishing rules for the use of such compounds. Although used intensively and extensively, there is no information on the amounts of disinfectants that have been applied in the aquaculture industry or by the processing plants and the food industry. This makes it very difficult to determine precisely the quantities of these products in the wastewater and environment in order to assess the risk to the fish and nontarget organisms. As far as we know, only the UK reports the required quantities of disinfectants being used in aquaculture activities (SEPA, 2018). Despite that, it is important that the use of disinfectants in aquatic settings include procedures that determine their safe levels in food and the environment. In regard to antifouling biocides, the European Union, North American, Asian, and Oceanian countries have regulations for manufacturing and marketing of these compounds. In the European Union, both the Water Framework Directive (WFD, 2000/60/EC) and the Marine Strategy Framework Directive (MSFD, 2008/56/EC) impose measures against the pollution of surface waters, including biocidal products. The European Union requires ecological risk assessments before approving the use of biocidal products, in accordance with the Biocidal Products Regulation (BPR) No 528/2012. In this sense, the European Commission provides an EU-wide framework and harmonized rules among European countries to ensure that risks are properly assessed before biocidal products are placed in the market. The product must be shown to be safe for human health, animal health, and the environment, besides proving effective for its intended uses. The European Chemicals Agency (ECHA) provides the European Commission with technical and scientific assistance in employing the regulatory processes (ECHA, 2018). In the USA, the chemical germicides formulated as sanitizers, disinfectants, or sterilants are regulated by the Antimicrobials Division, Office of Antifoulants and disinfectants 47

Pesticides Program inserted in the US Environmental Protection Agency (USEPA), under the authority of the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), which also regulates manufacturing and marketing of pesticides (including antifouling biocides) and requires that biocidal products distributed or sold in the country are subject to a registration review program, according to the Food Quality Protection Act (FQPA) of 1996. The Federal Insecticide, Fungicide, and Rodenticide Act also requires the registration of any substance or mixture of substances intended to prevent, destroy, repel, or mitigate any pest before sale or distribution. To get a registration, a manufacturer must submit specific data on the safety and effectiveness of the product. For instance, biocidal formulations should be tested at the manufacturer’s expenses, by using accepted methods for microbiocidal activity, stability, and toxicity to animals and plants. If USEPA concludes the product presents acceptable risks, then its labeling is registered, allowing the manufacturer to sell and distribute the product in the USA. As in Europe, only biocides not causing unreasonable risks to human health or the environment will be approved for re-registration (USEPA, 2020). In regard to disinfectants, the US Centers for Disease Control and Pre- vention (CDC) has the responsibility to guide the healthcare personnel on how to prevent and respond to infectious diseases both in healthcare settings and at home, based on the current scientific evidences about those products on safety and efficacy, and then recommend which chemicals might be the most appropriate or effective for specific microorganisms and settings. As part of their regulatory authority, EPA and FDA support the development and validation of methods for assessing disinfection claims. USEPA also reg- isters the effectiveness of antimicrobial products against common pathogens. The use of the listed EPA-registered products consistent with the product labeling complies with the Occupational Safety and Health Administration’s requirements for Occupational Exposure to bloodborne Pathogens (29 CFR 1910) as well as proper management of any waste when disposed, which is regulated under the Resource Conservation and Recovery Act (RCRA) (CDC, 2017; USEPA, 2018). The Government of Canada also published detailed documents about disinfectant uses (Government of Canada, 2018). In China, the disinfectants are regulated by the (1) Measures on Disin- fectant Administration issued by the Ministry of Health (MOH) in 2002; (2) the Guidance on Application of Administrative Approval License of Disin- fectants and Disinfecting Apparatuses (2006); and (3) the Administrative 48 Samantha Eslava Martins and Camila de Martinez Gaspar Martins

Licensing Procedure for Health-related Products (2006). Under those regulations, any company that plans to place a disinfectant into the Chinese market must obtain a hygiene license for the product first. Foreign companies are required to indicate a Chinese agent to proceed with the application. Imported disinfectants require approval by the National Centre for Health Inspection and Supervision (NCHIS) of the Ministry of Health, while for domestic disinfectants, an application shall be submitted to provincial health authorities (CIRS, 2018). China and South Korea also require the registration of all antifouling biocides used in paints in the country. In Japan, environmental risk assessments of antifoulants are also carried out. Oceania countries also have regulations for antifouling biocides. The New Zealand’s document on the Decision on the Application for a reassessment of Antifouling Paints, adopted in 2013, states Irgarol and chlorothalonil are no longer allowed as antifoulants and that paints containing Diuron, octhilinone or ziram have a time-limited approval (New Zealand, 2013). Australia has 46 products registered for use as antifouling paints under the Australian Pesticides and Veterinary Medicines Authority (APVMA, 2017). In Latin America and Africa, on the contrary, there is a lack of antifoulant regulation. Since TBT ban in the 2000s, it seems that little or nothing has evolved in regard to antifouling biocide regulation. Similarly, the regulation of the use of disinfectants in aquaculture is yet to be implemented or improved. To report a few examples of antifoulants regulation, some antifouling co- biocides commonly used in aquaculture have been already banned from the European Union. For instance, Diuron and chlorothalonil are not allowed in formulations placed on the market since 2008 (Tornero and Hanke, 2016). More recently, the use of Irgarol 1051 is no longer approved as an antifouling biocide since 2016 (CBP, 2011). On the contrary, the fact that a biocidal product is approved to be used as an antifoulant does not assure that it will not pose any risks to the environment. For instance, modeling the environmental distribution of DCOIT and zineb indicated that these compounds or their degradation products might put the marine harbor areas at risk. However, their use showed to be safe for the surrounding marine environment so that the Committee on Biocidal Products approved DOCIT and zineb as active substances for use as antifoulants, subject to pre- cise specifications and conditions. Besides the use of biocidal products, their disposal is also of environmental and health concern. There is no globally uniform disinfectant disposal Antifoulants and disinfectants 49 policy. However, their inherent toxicity prohibits the disposal of disinfectants in open waters or open water systems. Thus, the disinfection can only be reasonably applied to hatcheries and tank holding facilities. The Food Drug Administration (FDA) and Environmental Protection Agency (EPA) regulate the discharge of disinfectants into the environment in the USA. Similarly, in Europe, Canada, China, South America, and Australia, there are also specific regulations concerning the availability, proper use, and forms of disposal for specific disinfectants. However, most regulation is related to freshwater systems and the effects of disinfectants in the marine environment appear to be poorly studied.

3.6 Further considerations This chapter showed that in spite of being necessary for good husbandry, the use of antifoulants and disinfectants may lead to toxic effects on nontarget organisms, including farmed animals. In this sense, many efforts have been put in finding solutions that are more environmentally friendly, as an attempt to ensure high quality in animal production without severe impacts to the natural ecosystems. As an alternative to minimize the environmental impacts resulting from aquaculture and to optimize the production of organisms, the use of recirculating systems has increased as an environmentally friendly mode of cultivation without harming natural ecosystems or fisheries with the release of pollutants such as disinfectants. The reasons for using these systems are mainly biosecurity, environmental, and marketing advantages over conventional extensive and semi-intensive systems. They constitute closed or semi- closed circuits in which the circulating water is treated and reused/ recirculated (Bostock et al., 2010; European Commission, 2012; Fredricks, 2015). Some risks such as pathogen introduction, escape of exotic species, and discharging of wastewater (pollution) are reduced or even eliminated when water is recirculated. Furthermore, these recirculating systems contribute with high productivity and with the possibility to cultivate marine species at inland locations. The examples of recent aquaculture approaches that use recirculation systems are the integrated multi-trophic aquaculture (IMTA) systems that use the “biofloc technology (BFT).” The first requests to biodiversify fed aquaculture (e.g., finfish or shrimps) with extractive aquaculture, recapturing the inorganic (e.g., seaweeds) and organic (e.g., suspension- and deposit-feeders) nutrients from fed 50 Samantha Eslava Martins and Camila de Martinez Gaspar Martins aquaculture for their growth. The combination of fed-extractive aquaculture aims to engineer food production systems both providing biomitigative services to the ecosystem and improving economic farm output through the co-cultivation of complementary species (Chopin et al., 2012). The BFT is based on the growth of microorganisms (biofloc) in the culture medium with two major roles: maintenance of water quality, by the uptake of nitrogen compounds generating “in situ” microbial protein; and nutrition, increasing culture feasibility by reducing feed conversion ratio and a decrease of feed costs. Among the possible types of treatments to ensure the water quality for its reuse/recirculation, there are three processes particularly relevant: aeration/oxygenation, biofiltration, and ozonation. The integration of disinfectants in these processes may have direct effects on their normal mode of action, calling into question the quality of water and the health of animal production. Many solutions aiming at minimizing biological effects and toxicity of antifoulants have been proposed, for example, antifouling coatings based on polyethylene oxide, acrylic resins, and silicones. The latter have some formulations already placed into the market, as they have been proved to be efficient against fouling on high-speed ship hulls. However, more studies are needed to tackle the challenge of their application and fixing on textiles used in aquaculture such as nets and ropes (Amara et al., 2017).

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Metals

Claudia B.R. Martinez, Juliana D. Simonato Rocha, and Paulo Cesar Meletti Laboratory of Animal Ecophysiology, Department of Physiological Sciences, State University of Londrina, Londrina, Parana, Brazil

4.1 Introduction There is a long-standing controversy concerning the scientific use of the terms metals and heavy metals. The term “heavy metals” is commonly used in the literature to refer to metals and metalloids associated with environmental pollution, toxicity, and adverse effects on biota and has been diversely defined, mostly in terms of density, relative atomic mass, and atomic number (Ali and Khan, 2018). However, according to some researchers, this term is imprecise, outdated, and chemically meaningless (Wood, 2012) and The International Union of Pure and Applied Chemists (IUPAC) has advised against its use (Nikinmaa, 2014). Thus, in this chapter, we will use the term metals. The metals can be classified into essential or nonessential metals. The so- called essential metals (e.g., copper, zinc, chromium, nickel, cobalt, molyb- denum, and iron) are required in trace amounts for biological life owing to their participation in several biological processes, such as synthesis and repair of DNA, neurotransmission, molecular metabolism, respiration, cell signaling, and among others (Park and Jeong, 2018). At physiological concentrations, these metals maintain the homeostasis of the animal; however, toxicity occurs either at metabolic deficiencies or at high concentrations. On the contrary, the nonessential metals (e.g., aluminum, cadmium, mercury, tin, and lead) have no proven biological function and are toxic immediately after a limiting concentration has been reached, and their toxicity rises with increasing concentration (Nikinmaa, 2014; Sfakianakis et al., 2015). Metals are neither created nor destroyed, but they are redistributed in the environment. The contamination of aquatic ecosystems by metals may occur naturally, by the erosion of surface deposits of metal minerals, as well as from human activities, such as mining, smelting, fossil fuel combustion, and industrial application of metals (Nordberg et al., 2007). The industrial

Aquaculture Toxicology © 2021 Elsevier Inc. 59 https://doi.org/10.1016/B978-0-12-821337-7.00002-5 All rights reserved. 60 Claudia B.R. Martinez et al. and commercial uses of metals are continuously increasing and aquatic ecosystems, especially freshwater bodies and are liable to metal contamination due to disproportionate discharges of industrial and domestic wastes in their waters (Pandey et al., 2019). Globally, alarming levels of metals were already reported in the water and sediment of fluvial ecosystems around the world. For example, in China, high concentrations of arsenic, cadmium, chromium, nickel, zinc, copper, and lead were detected in the water and sediment of the River Jinjiang (Liu et al., 2018). The concentrations of copper, nickel, and zinc detected in four Korean rivers regularly examined for metals contamination between 2008 and 2015 were found to be high enough to pose risks to aquatic communities (Pandey et al., 2019). In Brazil, the collapse of a dam in 2015 released 60 million tons of mining waste into the environment and caused the burden of anthropogenically derived metals, such as iron, arsenic, mercury, and manganese, to the Doce River basin (Almeida et al., 2018). The transport and release of the tailing plume along more than 650km, down the river to the Atlantic coast, impacted freshwater, estuarine, and coastal marine ecosystems on a scale yet unknown (Gomes et al., 2017; Hatje et al., 2017). However, analyzing metal concentrations of the water body or sediment is not sufficient to determine its toxicity and several biological factors influence the bioavailability of metals (Magalha˜es et al., 2015). Bioavailability is the potential for uptake of a substance by a living organism and is affected by the interactions with the environment (Nikinmaa, 2014). Metal bioavailability in water is determined by the water chemistry. The free metal ion is the bioavailable chemical species and metal speciation in water is greatly influenced by abiotic factors such as pH, and the presence of ligands in the water that may remove free ions, anions that can complex metals or form anionic metal species, or cations that may be competing for biological uptake with trace metals (Shaw and Handy, 2011). In freshwater, the speciation chemistry of different metals varies greatly, but in general lower pH increases the free ion concentration, thereby increasing toxicity, whereas alkalinity and inorganic anions tend to complex metal ions, thereby decreasing toxicity. The hardness cations (Ca2+ and Mg2+) as well as Na+ and K+ (and sometimes H+) may also decrease toxicity by competing for metal-binding sites on the gills. However, in many natural waters, the most effective agent of protection against most metals is the dissolved organic matter (Wood, 2012). In seawater, the importance of metal speciation has been less studied, partly because seawater composition is less variable in comparison with freshwater. The most obvious variable is salinity, with all major ions Metals 61 covarying when salinity changes. For many metals, complexation by the À high levels of Cl dominates speciation in seawater, and this, combined with the greater availability of other anions plus the protective effect of competition by high concentrations of Na+,Mg2+, and Ca2+, means that most metals are far less toxic in seawater than in freshwater (Wood, 2012). Fishes are very sensitive to aquatic pollutants and represent key elements for water quality assessment and the most frequently used organisms in the monitoring of metal contamination. Teleost fishes can accumulate metals in their body through direct uptake from water, via gills, and through diet (Vilizzi and Tarkan, 2016). Fishes are ubiquitous in almost all environments and their displacement ability increases the potential for exposure to contaminated areas. Another aspect is the great ecological importance of these organisms due to their influence in the structure of the food chain, nutrient cycling, and energy transfer. Thus, they are suitable experimental models, since, besides being very sensitive to toxicity and the bioaccumulation of the metals, they also serve as bioindicators of the availability of these chemical elements in water (Atli and Canli, 2010; Kroonetal.,2017). Due to the great diversity of metals, and the consequent variability of their toxic effects on organisms, the use of multiple biomarkers considering different levels of the biological organization should be considered. Thus, many studies evaluate possible alterations from the molecular level— through biochemical and physiological parameters—to histological/morphological analyses, if we consider only the subindividual levels. All these analyses are very important in understanding the effects of metals on fish and, when considered together, allow reliable evaluation. In addition, these analyses usually follow an order of occurrence and, therefore, of sensitivity/ precocity in the indication of effects. Biomarkers have been used extensively to provide the connection between external levels of contaminant exposure, internal levels of tissue contamination, and early adverse effects in organisms (Kroon et al., 2017). Among the biomarkers widely used for evaluation and monitoring of contamination by metals are the biochemical and physiological biomarkers. The behavior component does not constitute a specific level but runs through this organizational structure, so that behavioral alterations may occur due to disturbances at one or more of these levels (Fig. 4.1). In this chapter, we will focus mainly on biochemical, physiological, and behavioral changes related to metal effects on finfish. For more information on metal toxicity in Crustacea and mollusks, the reader is referred to Chong (2020a, b). 62 Claudia B.R. Martinez et al.

Fig. 4.1 Levels of biological organization and the relation with the behavior: disruptions in one or more levels can lead to behavioral changes. However, it may be possible to observe a punctual behavioral signal, such as avoidance, before changes at any of these levels can be significantly detected.

4.2 Biochemical effects Alterations at the biochemical level are typically initial responses to any toxicant. Consequently, biochemical biomarkers may be extremely sensitive indicators of alterations in cellular functions (Benhamed et al., 2016; Schlenk et al., 2008). Among biochemical parameters, oxidative stress parameters are highly associated with metals and are indicative of symptoms of toxicity. For the protection of fishes against oxidative stress, all antioxidant defense pathways such as enzymes, peptides, and metalloproteins need to work together. Thus, responses and effects on these antioxidant lines of defense and oxidative stress indicators should be monitored regularly for early indication of pollution, before a population decline is detected (Di Giulio and Hinton, 2008). The metals present valence characteristics that can be variable or constant, which implies how these metals interfere or interact in cellular processes. Magnesium, strontium, and barium, for example, affect the transport of calcium and zinc, which can compromise energy production and gene expression by replacing them in regulatory enzymes. However, metals with varying valences such as iron, copper, manganese, and chromium, due to their involvement in the Fenton reaction, promote more critical Metals 63 alterations that can lead to increased production of reactive oxygen species (ROS), being the main inducer of oxidative stress. Many of these metals, although essential, may cause ROS generation. However, these ROS will be neutralized by the cell, unless conditions of chronic exposure and/or very high concentrations disturb these pathways, leading to the development of oxidative stress (Lushchak, 2016). Iron is an essential element, a constituent of many proteins (such as hemoglobin and cytochromes) where its ability to go through redox cycles is the key to its function. Under physiological conditions, Fe+2 is soluble and unstable and tends to be oxidized, particularly by molecular oxygen, forming superoxide anion. Thus, iron can initiate processes and propagate nonspecific free radicals that are harmful to cells, being one of the main elements responsible for ROS production in fish (Atli and Canli, 2010). On the contrary, Cu deficiency can lead to oxidative stress, as it is a component of several antioxidant systems. For example, this metal is an element of cytochrome oxidase and Cu, Zn-superoxide dismutase (Cu, Zn-SOD), which play clearly antioxidant roles. Thus, a reduction in the uptake or interruption of the cellular metabolism of copper promotes an increase in the concentration of ROS (Lushchak, 2016). In addition, an increase in the concentration of this metal may also result in the saturation of the cell with this micronutrient and, when the cell is not able to eliminate the excess, it can catalyze the Fenton reaction, resulting in the greater generation of hydroxyl radicals ( Jia et al., 2017). ROS generation induced by copper can lead to the establishment of oxidative stress as observed in the neotropical freshwater species Apistogramma agassizii (Braz-Mota et al., 2018) and Prochilodus lineatus (Simonato et al., 2016), in Danio rerio (Yin et al., 2018), as well as in the marine species Dicentrarchus labrax (Dı´az-de-Alba et al., 2017) and Cyprinodon variegatus (Adeyemi and Klerks, 2012). After copper exposure, these fish species demonstrated increased lipid peroxidation (LPO), defined as a sequence of biochemical events resulting from the action of free radicals, mainly on the polyunsaturated fatty acids of cell membranes. Chromium is an essential trace element, involved in the regulation of a wide range of biological processes, which presents changes in the valence state, being able to be toxic even at low concentrations; furthermore, as it is not biodegradable, it remains in ecosystems for a long time. Chromium participates in the generation of hydroxyl radical and has a role as a catalyst in oxidative processes; in this way, it is accepted that chromium has the capacity to generate reactive oxygen species (Ahmad et al., 2006). This 64 Claudia B.R. Martinez et al. ability was observed in Cr-exposed Nile tilapia (Oreochromis niloticus), which presented both stimulation and inhibition of hepatic and renal enzymes of the antioxidant defense pathway (Atli and Canli, 2010). As with iron, copper, and chromium, the exposure to nickel also promoted and increase in LPO in P. lineatus (Palermo et al., 2015) and in rainbow trout Oncorhynchus mykiss (Topal et al., 2017). Although nickel has no biological function, its widespread use by industry leads to environmental pollution, resulting in large accumulation in animals. The most common oxidation state of nickel is Ni2+, while other valences, such as Ni+ and Ni3+, are also well known, and due to their ability to enter redox processes, they participate in the Fenton reaction, which may lead to increased ROS production. Lead is a nonessential metal, dispersed throughout the environment mainly as the result of anthropogenic activities (ATSDR, 2007). The toxic effects of lead on aquatic animals are associated with the overproduction of ROS, resulting in DNA damage and depletion of cell-antioxidant defense systems (Zhang et al., 2007). In a study conducted by Monteiro et al. (2011), in vivo and in vitro exposures were used to assess the genotoxicity of lead (Pb) to P. lineatus. These authors showed the potential genotoxicity and cytotoxicity of environmentally relevant concentrations of the lead after acute exposures. Moreover, the results clearly indicated that the mechanisms of lead genotoxicity to P. lineatus are related to the generation of ROS. Mercury is not an essential metal and is very widespread in the aquatic ecosystems. Mercury exists as a cation with an oxidation state Hg+1 or Hg+2. In the environment, mercury can be found in the form of methylmercury, principally produced as the result of the methylation of the inorganic form (Hg+2), mediated by microorganisms in soil and water. Among its toxic effects, the main one is the interaction of inorganic or organic mercury with sulfhydryl-containing residues (Rooney, 2007). The effects of mercury can arise by its direct interaction, but also through the induction of oxidative stress, leading to injury and cell death (Lushchak, 2016). In experiments, Atlantic salmon (Salmo salar) exposed to mercury chloride presented an increase in LPO products (Berntssen et al., 2003). Chronic exposure of medaka (Oryzias melastigma) to mercury chloride resulted in changes in proteins related to mitochondrial function in hepatocytes, suggesting that this organelle could be the primary target for mercury attack in cells. Thus, hepatotoxicity is one of the principal toxic mechanisms of mercury in fish and other aquatic organisms (Wang et al., 2013). As already well established for Metals 65 mammals, methylmercury also had neurotoxic effects on zebrafish embryos (Danio rerio), these effects were associated with the induction of oxidative stress, which affected the expression of genes involved in neuronal activity and learning (Ho et al., 2013; Rasinger et al., 2017). Arsenic is a metalloid with different oxidation states, the most common À of which are As5+,As3+, and As3 , and can form inorganic and organic compounds, in the environment and inside cells. Oxidation of As3+ to As5+, under physiological conditions, results in the formation of hydrogen peroxide (H2O2). This indicates its involvement in oxidative processes, particularly oxidation of lipids, DNA, and proteins (Lushchak, 2016), including the induction of apoptosis (Datta et al., 2007; Doganlar et al., 2016; Li et al., 2016). The exposure to As promoted an increase in LPO, with substantial modifications in antioxidant enzymatic defenses, in freshwater fish such as air-breathing catfish Clarias batrachus (Kumar and Banerjee, 2014) and wild brown trout Salmo trutta (Greani et al., 2017) as well as in the marine rockfish Sebastes schlegelii (Kim and Kang, 2015). It is important to point out that several studies also showed the establishment of oxidative stress caused by metals mixtures (Benhamed et al., 2016; Green and Planchart, 2018; Souza et al., 2018; Stankeviciut e_ et al., 2017; Svecevicius et al., 2014), which is the most common scenario in the natural environment. In addition, metals mixtures may act synergistically, resulting in different levels of toxicity (Obiakor and Ezeonyejiaku, 2015; Oliveira et al., 2018). The interactions between different metals may lead to competition for absorption from the environment and distribution in tissues of fish, which results in some metals influencing the accumulation of other metals in these organisms (Svecevicius et al., 2014). Taken together, the studies cited earlier clearly indicate that metals, whether essential or not, have the capacity to promote oxidative stress. Normally, there is a balance between the rate of production and removal of ROS, but if the balance is disturbed, the excess of ROS can alter cellular properties and functions, causing the oxidation of proteins, DNA, and lipids, leading to cell death. Therefore, the balance between the production of ROS and their elimination must be maintained in order to prevent metabolic disturbances or oxidative stress (Di Giulio and Meyer, 2008). In order to keep the redox balance, several antioxidant enzymes, such as glutathione peroxidase (GPx), glutathione S-transferase (GST), glutathione reductase (GR), catalase (CAT), and superoxide dismutase (SOD), and nonenzymatic antioxidants, such as glutathione (GSH) and other thiols (–SH), provide protection against ROS. These antioxidants are commonly used as biochemical 66 Claudia B.R. Martinez et al. biomarkers in fish exposed to various environmental contaminants, including metals (Di Giulio and Meyer, 2008; Kroon et al., 2017). An important characteristic of the biomarkers related to antioxidant defenses is that they are nonspecific for different types of metals and the responses vary greatly depending on the fish species, age, and, especially, tissues. The exposure of the pale chub (Zacco platypus) to Cu promoted an increase in the hepatic activity of SOD and CAT, but GST did not vary (Kim et al., 2014); while in P. lineatus, SOD also increased after copper exposure, while GST and CAT did not vary (Simonato et al., 2016). In addition, in a single work with the goldfish, Carassius auratus, exposed to different copper concentrations, it was possible to observe the variations between organs: gill and renal GST activity decreased, while hepatic and cerebral did not change; SOD activity increased only in the liver at the highest concentration; and finally, liver and gill CAT increased with the decreased renal activity of this enzyme (Husak et al., 2018). Atli and Canli (2010) also observed this variation with respect to organs and different metals in a study with O. niloticus. These authors showed that hepatic SOD activity was induced after acute exposure to Cd, Cu, Cr, Zn, and Fe, whereas CAT was inhibited by Cu and Cr; and GST was inhibited by all the metals; in the kidney, CAT was inhibited by all metals, except by Cd, SOD only by Cr, Zn, and Fe; and renal GST was not altered. Despite the large variations in the responses of antioxidant defenses and oxidative stress products, a slightly more specific biochemical biomarker for metals is the measurement of metallothionein concentration. As already mentioned, some metals participate in biological functions essential to organisms and thus need to be maintained at physiological concentrations within the body. In addition to transport into and out of cells, several lysosome-like organelles participate in metal homeostasis to buffer their cytosolic concentrations and act as storage sites. In both cases, the movement of the metal is carefully controlled by transporters. In response to the excess of metal ions, sequestration by metal-binding proteins occurs, which can subsequently be released under the conditions of deficiency. This control is performed by metal-binding proteins, rich in cysteine residues (Cys), called metallothioneins (MT). These are an important group of proteins involved in metal detoxification due to their chelating ability, maintaining metals at concentrations necessary for the proper functioning of the organism (Carpene` et al., 2007; Fabrin et al., 2018). Thus, without the careful regulation of metals, the normal functions of cells, such as ROS formation, which can promote oxidative damage, occur. The MT may also present Metals 67 affinity for several nonessential metals such as mercury, lead, and cadmium, promoting the displacement of the binding site to the essential metals, damaging the balance of these metals (Park and Jeong, 2018). Several studies have demonstrated alterations in MT against the contamination by metals and for this reason, these proteins are considered important biomarkers in aquatic environments, as observed in the estuarine fish Cen- tropomus parallelus collected in the field, which showed the concentration of MT directly correlated with the concentrations of Fe, Se, Zn, Mn, and As in the water (Souza et al., 2018). In another study that evaluated the mixture of Zn, Mn, and Fe in the freshwater fish P. lineatus, only Zn promoted an increase in the concentration of MT (Oliveira et al., 2018). This same fish species also presented increased MT when exposed to Pb (Monteiro et al., 2011) and Cu (Simonato et al., 2016), but exposure to Zn for 96h induced an increase in hepatic but not gill MT (Palermo et al., 2015). These results demonstrate that MT concentrations may not present a specific pattern (Fabrin et al., 2018) and that many variables should be considered when using MT as a biomarker, such as the tissue studied and abiotic factors such as pH and salinity, which interfere with the bioavailability of the metal.

4.3 Physiological effects Freshwater fish are hyperosmotic with respect to the environment in which they live, which promotes a constant osmotic gradient for water entry into the body and the loss of ions through the gill epithelium. To eliminate the excess of water, the kidney of these animals produces a large amount of dilute urine, and the gills are responsible for the active uptake of ions, mainly À Na+ and Cl . The kidney of freshwater teleosts also has the function of 2+ 2+ À reabsorption of divalent ions such as Mg ,Ca , and SO2 . The activity of ATPases in renal and gill cells has a chief role for the ionic transport through the membranes, maintaining the osmo-ionic balance in these animals. On the contrary, marine teleosts are hyposmotic in relation to the saltwater and consequently lose water through osmosis and also through urine. To replenish this water lost, these fishes drink saltwater, which is absorbed into the blood through the gastrointestinal tract, where Na+ and Cl+ are transported into the blood. Thus, the gills assume the important role of À the excretion of the excess of Na+ and Cl from the plasma to the external medium. The kidney of marine teleosts excretes isosmotic urine in relation to the blood plasma and has the role of the elimination of divalent ions such À as Mg2+,Ca2+, and SO2 . In this way, the gastrointestinal tract, gills, and 68 Claudia B.R. Martinez et al. kidney of these animals are directly involved in the maintenance of the osmo-ionic balance and are the main entrance routes of toxic compounds, which makes them important target organs for ecotoxicological evaluations (see Hwang and Lin, 2014 for revision). The metals are known to interfere in the osmoregulation of teleosts, mainly through direct effects that promote alterations in key enzymes responsible for the ionic balance. In freshwater fishes, the exposure to copper interfered on Na+ homeostasis (Glover et al., 2016) as the result of alterations in the activity of enzymes such as Na+/K+-ATPase, and H+-ATPase in the gills (Chowdhury et al., 2016) and also in the kidney (Moyson et al., 2016). Lead was also able to promote alterations in the gill Na+/K+-ATPase, which triggered an imbalance in plasma concentrations of Na+ and K+ in P. lineatus (Ribeiro et al., 2014). In addition, some nonessential metals have the ability to enter into the organism using ion channels, promoting significant changes in ion concentrations in the fish. Both lead and cadmium are examples of metals analogous to Ca2+ and compete for the same binding sites located in the apical membranes of mitochondria-rich cells and pavement cells in the fish gills (Clemow and Wilkie, 2015). Another important aspect to point out refers to the consensus that abiotic factors, such as hardness and salinity, are able to modulate the effects caused by exposure to metals. Hardness, for example, is a condition that considerably affects the physiology and toxicity of metals in fish. Nile tilapia exposed to copper and cadmium with different degrees of water hardness (soft or hard according to the concentration of Ca2+) presented an increase in the activity of gill and renal Na+/K+-ATPase, in exposure to hard water, but a decrease in this activity when exposed to soft water (Saglam et al., 2013). Water hardness may interfere with the toxicity of metals such as lead and cadmium; these metals become more toxic in soft water, poor in Ca2+, than in hard water, rich in Ca2+, due to decreased competition of these metals (Pb and Cd) for Ca2+-binding sites (Clemow and Wilkie, 2015). An increase in salinity may also affect the physiology and uptake of metals in freshwater fishes. Nile tilapia exposed to copper in increased salinities presented an alteration in the response pattern of ATPase activity depending on the tissue, salinity, and duration of exposure. The activity of Na+/K+- ATPase decreased after exposure to copper in lower salinity and increased after Cu exposure in higher salinity. The gill and intestinal Ca2+-ATPase activity decreased in all salinities, while the activity of this renal enzyme increased (Kulac et al., 2013). The toxicity of aluminum is also known to be affected by several water quality parameters such as hardness, Metals 69 temperature, and dissolved organic carbon (Santore et al., 2018); however, the most important chemical parameter for Al is pH, due to interactions in the speciation and solubility of this metal (Cardwell et al., 2018; Gensemer et al., 2018). For example, increased pH, hardness, and DOC promoted a protective effect against Al toxicity in the freshwater species Pimephales promelas (Gensemer et al., 2018). The exposure of fish to Al may promote osmoregulatory disorders mainly at an acidic pH due to the formation of dissolved monomeric species such as the free form Al3+, and respiratory damage due to the accumulation of precipitated forms in exposure at neutral pH (Cardwell et al., 2018; Gensemer et al., 2018). The intestine is also an important target for toxic agents present in food and water and, given the important osmoregulatory function of the intestine, these compounds directly affect the regulation of the balance of water and ions (Kulac et al., 2013). In addition, the difference of more than 10 times in the rate of water consumption of marine teleosts compared to freshwater fish has important consequences for the exposure of the intestine to metals. While freshwater teleosts present the pharynx as the main site of contact with ambient water, due to the result of respiratory movements, in marine teleosts the entire gastrointestinal tract is exposed, particularly the rectal portion where the absorption of water and ion mainly occurs. The actions of the metals in the intestine are comparable to those of the gills, as they affect the ionic transport, increasing or decreasing the activity of enzymes such as Na+/K+ ATPase, Ca2+-ATPase, and among others (Wendelaar Bonga and Lock, 2008). In the euryhaline fish, Fundulus heteroclitus, both branchial and intestinal Ca2+-ATPase activities decreased in response to Zn at 0ppt and were elevated at 35ppt (Loro et al., 2014). On another work, Baldisserotto et al. (2004) showed that high dietary Ca2+ could reduce the toxic effect of waterborne cadmium exposure in rainbow trout, as it decreases cadmium uptake. Consequently, fish that feed on a Ca2+-rich invertebrate diet could be more protected against waterborne cadmium toxicity in a field situation. These results demonstrate that special attention should be paid to water chemistry in studies with metals and that the alterations in teleosts osmoregulation are sensitive indicators for the evaluation of environmental contamination by metals.

4.4 Behavioral effects Behavior is the visible manifestation of the integrated physiological response of an animal to its environment, so that behavioral integrity may 70 Claudia B.R. Martinez et al. represent the state of health of the organism (Clotfelter et al., 2004). Thus, by integrating effects of multiple physiological processes that can directly affect the survival and reproduction of organisms, behavioral analyses are useful tools to predict the impacts of exposure to contaminants (Marentette et al., 2012; Scott and Sloman, 2004). A xenobiotic can trigger different behavioral responses in aquatic organisms, which can start from rapid responses—as a form of immediate protection—such as avoidance, through intermediate responses, such as alterations in balance, feeding, and looking for a hiding place (burrowing), to more long-lasting responses, such as those related to locomotion, mating, nesting, memory, and risk exposure (Hellou, 2011). These responses may reduce the contact with the substance, provided it is chemically perceived (e.g., odorant) and is in fact avoidable, which depends on the ability of the species to move away from the contaminant plume dispersion or from the area where the toxicant is most present, which is not always possible. Moreover, in a worse scenario, according to Tierney (2016), the substance may not be perceived as harmful, and even worse, the animal may even pre- fer the toxic environment. If it is possible to escape contact with the xenobiotic, effects on individual organisms can be avoided, however not the ecological effects in the absence of a population in that area (Fig. 4.2).

Fig. 4.2 Avoidance behavior from the sensorial detection of a xenobiotic: depending on the locomotion ability of the animal, the exposure to toxic substances can be avoided or not. In the case of continued contact with these substances, further behavioral changes may appear as well as related physiological changes. Both direct avoidance and secondary behavioral alterations can lead to ecological changes, although causal relationships are difficult to establish in field studies. Metals 71

A number of studies indicate that metals may affect the behavior of aquatic organisms (Atchison et al., 1987; Sabullah et al., 2015), with actions on olfactory receptors (Heffern et al., 2018; Hellou, 2011; Scott et al., 2003; Tierney et al., 2010; Williams and Gallagher, 2013) and the central nervous system (Bradbury et al., 2008; Green and Planchart, 2018). A disturbance in metal homeostasis can dramatically impair brain function and cognitive performance, though the exact molecular mechanisms are still not well understood (Carpene` et al., 2017). According to Tierney et al. (2010), olfactory nerve cells play an important role in the behavior of fish as they are in direct contact with the environment and are therefore vulnerable to exposure to neurotoxic xenobiotics. Avoidance can be considered a protective behavior, especially in acute exposures to metals and other substances, but if damage to sensory neurons occurs in chronic exposures, the avoidance response can be atten- uated (Tierney, 2016). For example, cadmium in environmentally relevant À concentrations (up to 30μgL 1) affected olfactory sensory transduction in juveniles of coho salmon (Oncorhynchus kisutch), which was associated with significant metal accumulation in rosettes of the olfactory epithelium (Williams et al., 2016). In this same work, it was observed that the effects of attraction to the co-specific odorant and of repulsion to L-cysteine were À decreased after 16 days of exposure to 2μgL 1 of cadmium and were not restored even after 16 days of recovery in water without the metal. Blechinger et al. (2007) reported damage to the olfactory epithelium of D. rerio with impairment in the predator avoidance response even after 4–6 weeks of recovery in clean water. This study also related the defects caused by cadmium to the olfactory epithelium during the larval period to the deficits in olfaction observed in adult fish. Jin et al. (2015) observed that cadmium and chromium produced alterations in antioxidant defenses, embryonic developmental abnormalities, and also changes in spontaneous swimming behavior in D. rerio larvae. À Copper at ambient concentrations (60μgL 1) decreased the olfactory sensitivity of Ptychocheilus lucius to the warning substance (skin extract), with effects on the olfactory epithelium, found through scanning electron microscopy. However, ciliated cells of the olfactory epithelium were observed again after 14 days of recovery, and more effects were detected at 24-h than at 96-h exposure (Beyers and Farmer, 2001). These authors suggested that there had been a process of adaptation, such as occurs in general adaptation syndrome, a three-phase process that includes an initial loss of response, followed by a period of physiological adaptation, and culminating 72 Claudia B.R. Martinez et al. in the inability of the compensation mechanisms to sustain the level of activity needed to compensate for the effects of the stressor. In addition, the recurring damage and recovery process would represent a high biological cost. The neotropical teleost P. lineatus exposed to copper presented swimming in spurts (shorter swimming time, but at high speed) when exposed to À À 9μgL 1 Cu, but lethargy when exposed to 20μgL 1 Cu. This higher Cu concentration also altered the preference for the layer occupied in the water column, with possible consequences on relevant behavioral parameters, such as grazing (Simonato et al., 2016). The activity of acetylcholinesterase (AChE) may help in the understanding of some behavioral alterations. The inhibition of the activity of this enzyme is well known in relation to exposure to organochlorines and carbamates, but has also been associated with metals (Frasco et al., 2005; Zhang et al., 2017). When AChE is inhibited, there is a longer excitatory action of acetylcholine at the neuromuscular junction, which produces important behavioral alterations in spontaneous swimming, as observed by Zhang et al. (2017) in fish exposed to cadmium and by Simonato et al. (2016) in fish exposed to copper. Bradbury et al. (2008) reported the works in which inhibition of AChE activity caused alterations in spontaneous swimming, both hyper- and hypoactivity, orientation, learning, and, possibly, susceptibility to predation. On the contrary, Senger et al. (2011) observed a decrease in the swimming activity of D. rerio exposed to aluminum, but with an increase in the activity of cerebral AChE. Mercury may also have effects on behavior, as observed by Grippo and Heath (2003) in Pimephales promelas, which presented a decrease in foraging rate, but without significant damages in the consolidation of memories, or alterations in the activity of cerebral AChE. Atchison et al. (1987) reported that mercury may be related to both hyper- and hypoactivity, disturbances in learning, and avoidance of predators and may not trigger avoidance behaviors, while even provoking attraction to the metal. The same authors reported studies involving lead exposure, which caused an increase in swimming activity (contrary to what was observed by Zhu et al. (2016) for D. rerio), damage to learning (as also observed by Xu et al. (2016) for D. rerio), and a decrease in the capacity to capture prey. As demonstrated by these works, metals can induce behavioral responses that may alter the outcome of biological interactions if the effect is expressed in a predator, or its prey, or in competing species. Thus, if a given metal does not directly kill the animal, it can indirectly cause its ecological death by altering its behavior (Scott and Sloman, 2004). However, Fleeger et al. Metals 73

(2003) warn that the majority of these studies are developed in laboratory settings that do not allow for the identification of whether observed behavioral alterations would result in alterations in structure or trophic relationships. Thus, it is important to develop new techniques and approaches in behavioral ecotoxicology to decrease uncertainties as to the real effects, which could be observed in the field.

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Agrochemicals: Ecotoxicology and management in aquaculture

Vania Lucia Loroa and Bárbara Estevão Clasenb aLaboratory of Aquatic Toxicology, Department of Biochemistry and Molecular Biology, Federal University of Santa Maria (UFSM), Santa Maria, RS, Brazil bDepartment of Environmental Sciences, State University of Rio Grande do Sul, Tr^es Passos, RS, Brazil

5.1 Water and soil contamination by agrochemicals In general, agrochemicals are important in agriculture to protect crops and improve productivity. However, uncontrolled use or absence of efficient use, as well as illegal use of banned agrochemicals, can negatively affect the environment, human food, and aquaculture species. A sum of factors contributes to increased soil and water pollution including chemical characteristics of agrochemicals in common use, mixtures of compounds, and interactions with the environment. Contamination of the environment and organisms, especially aquatic organisms, is not a sur- prise nowadays considering the amount and variety of chemicals used. The soil and water frequently suffer contamination by a mixture of different compounds. A study of soil samples in the European Union found the presence of 76 pesticide residues from 317 agricultural samples (Silva et al., 2019). Around 80% of the tested soils contained pesticide residues (25% of samples had 1 residue and 58% of samples had mixtures of 2 or more residues), in a total of 166 different pesticide combinations. The soils originated from 11 European Union Member States and 6 main cropping systems were tested. Glyphosate, aminomethylphosphonic acid, DDT (dichlorodiphenyltrichloroethane), and its metabolites, the fungicide boscalid, epoxiconazole, and tebuconazole were the compounds most frequently found in soil samples and were found at the highest concentrations. This study revealed that the presence of mixtures of pesticide residues (58%) has contaminated the soil and soil organisms (Silva et al., 2019). The water contamination is usually determined in studies considering one or more years of water monitoring. To characterize the transport of neonicotinoids in freshwater ecosystem, a study collected water samples from 10 major tributaries to the Great Lakes, USA, for 1 year (October

2015–September 2016). At least 1 neonicotinoid was detected in 74% of monthly samples. The neonicotinoids imidacloprid (53%), clothianidin (44%), and thiamethoxam (22%) were found in all water samples collected and the most frequently recorded. The concentration of 330ngL 1 was the maximum individual neonicotinoid recorded. The sum of neonicotinoids was 670ngL 1 (Hladik et al., 2018). The neonicotinoids are the class of insecticides mostly used worldwide and their residues are increasing in watercourses near agricultural and urban areas. The neonicotinoids were recorded in soil, water, and sediment during a study considering five neonicotinoids of common use conducted in northern Belize, a region of high biodiversity. The results showed neonicotinoids’ presence in soil, water, and sediment samples. Imidacloprid was the most common residue found, reaching the concentration of 17.1ngg 1 in soil samples. The frequency of neonicotinoids and concentrations were highest in soil (68%) in comparison with water samples (12%). The concentrations in soils differed among crop types, being highest in melon fields and lower in banana and sugarcane fields. Another fact observed is that the distance of the crop fields determines the decline in residues of neonicotinoids. Clothianidin residues were detected at 100m and imidacloprid at more than 10km from the nearest crop fields (Bonmatin et al., 2019). The residues of thiamethoxam, chlorantraniliprole, and tebuconazole were found in the water of an irrigated rice paddy system 53 days after application in southern Brazil (Clasen et al., 2018). The contamination of soil, water, and sediment by agrochemicals is a worldwide problem. A study in Silesia (Poland) monitoring some chemicals of common use during 1 year (2014) verified the occurrence of atrazine products in 41% of sediments, 71% of soil, and 8% of surface water samples. Sulcotrione was determined in 85% of soil samples; its degradation product (2-chloro-4-(methylsulfonyl) benzoic acid) was present in 43% of soil samples and 17% of sediment samples (Barchanska et al., 2017). Monitoring study conducted during four seasons in vineyard region of La Rioja (Spain) covering 12 surface waters and 78 groundwaters to detect agrochemical contamination showed the residues of some pesticides and insecticides. Terbuthylazine and its metabolite desethylterbuthylazine were present in 65% of water samples in all seasons sampled. Other herbicides detected in 50% of samples and at least one season were fluometuron, metolachlor, alachlor, and ethofumesate. The insecticide pirimicarb was detected in 25% of the samples. The sum of results considering compounds detected (mainly herbicides) was higher than Agrochemicals: Ecotoxicology and management in aquaculture 81

0.5μgL 1 in 50% of the samples, especially in the periods with the highest application of these compounds (herbicides in March and insecticides in June) (Herrero-Herna´ndez et al., 2017). Water contamination due to agrochemicals was detected in several countries, and potential risk was identified in Mondego River estuary, located on the North Atlantic Ocean Ecoregion (Portugal), which is a basin affected by agricultural runoff with increasing signs of eutrophication. A study evaluated the amounts and distribution of 56 priority pesticides belonging to distinct categories (insecticides, herbicides, and fungicides) during 1 year (2010–11), in a total of 42 surface water samples. More than 55% of the quantified pesticides were above the maximum amounts established by the European Directives (98/83/EC and 2013/39/EU) and some of the measured compounds were able to cause mortality in fish, invertebrates, and crustaceans (Cruzeiro et al., 2016). The fish that inhabit rice ponds are exposed to a high risk of agrochemical contamination due to the mixture of pesticides (Rossi et al., 2020). Antiox- idant mechanisms failed to prevent oxidative damage in the liver and gills of Markiana nigripinnis and Astyanax lacustris. The latter species showed the inhibition of the antioxidant defenses but without tissue lipid peroxidation. The activity of acetylcholinesterase (AChE) was reduced after spraying in brain and muscle tissues of A. lacustris and in the brain of M. nigripinnis. The results showed that the agrochemicals of common use, like glyphosate; insecticide, like bifenthrin; and the fungicides, like azoxystrobin and cyproconazole, pose health risks on native fish populations inhabiting rice field (Rossi et al., 2020). A study verified 30 pesticide and biocide active substances and metabolites in the estuarine continuums in the Bay of Vilaine area (NW Atlantic Coast, Southern Brittany, France). The following agrochemicals were detected at least once: 11 triazines (ametryn, atrazine, des- ethylatrazine, desethylterbuthylazine, desisopropyl atrazine, Irgarol 1051, prometryn, propazine, simazine, terbuthylazine, and terbutryn), 10 phenylureas (chlortoluron, diuron, 1-(3,4-dichlorophenyl)-3-methylurea, fenuron, isoproturon, 1-(4-isopropylphenyl)-3-methylurea, 1-(4-isopropylphenyl)-urea, linuron, metoxuron, and monuron), and 4 chloroacetanilides (acetochlor, alachlor, metolachlor, and metazachlor). Diuron and Irgarol 1051 exhibited risks to primary producers in Arzal reservoir, close to marina. Diuron was recorded during all measurement periods, whereas Irgarol 1051 showed a clear seasonal pattern, with highest concentrations in June and July (Caquet et al., 2013). 82 Vania Lucia Loro and Bárbara Estevão Clasen

5.2 Environmental contamination by agrochemicals and risk assessment in aquaculture: Effects on aquatic organisms and food for human consumption Ecotoxicological risk assessment is based on measuring contaminant levels in water, soil, and food (fish, meat, shellfish, etc.) and linked to human health risk assessment (e.g., cancer risk) using the metrics of average daily intake of contaminated food items. Usually, fish and shellfish farming should avoid the use of pesticides, except for the control of some parasitic diseases, but antibiotics and antifungals may be used. The persistence of some agrochemicals in the environment increases the likelihood of adverse effects on fauna and flora where the contamination occurs. This requires the assessment of chemical mixtures and the prevailing environmental factors (abiotic) that alter the fate and distribution of agrochemicals. The environmental risk is determined by a balance of soil types, soil organic carbon, pH, and the rates of degradation in the various environmental compartments (Sanchez-Bayo and Hyne, 2011). A study carried out in Bangladesh verified that a weekly application of 1 and 2mgL 1 of the organophosphate sumithion in aquaculture ponds for 120 days did not affect water quality parameters. However, the abundance of benthic invertebrates significantly reduced in both sumithion-treated groups as compared to control group, suggesting that sumithion has a negative impact on aquaculture ponds (Uddin et al., 2016). All abiotic factors are expected to affect the distribution and fate of pesticides in the soil-water-sediment environment. For example, extended dry periods followed by rainfall could be the reason for an increase of pesticide mobility so that they flow to surface and groundwater bodies. Low water levels in surface water during the dry season have the potential to increase agrochemical concentrations. On the contrary, high temperatures are expected to increase the degradation and volatilization of agrochemicals in the soil sediment environment (Masia´ et al., 2013; Sanchez- Bayo and Hyne, 2011). The evaluation of risk assessment also needs to take into consideration the characteristic of agrochemicals, the properties to bioaccumulation, and biotransformation in aquatic organisms. Sometimes, biotransformation produces more toxic compounds. For example, the active ingredient glyphosate has low toxicity compared to POEA (polyoxyethylene), the surfactant present in the commercial formulation. Another possible effect of Agrochemicals: Ecotoxicology and management in aquaculture 83 agrochemicals is the reduction of the capacity of reaction against diseases or activity of the immune system. Litopenaeus vannamei (Pacific white shrimp) exposed to 2μgL 1 endosulfan for 20 days was highly susceptible to Taura syndrome virus (TSV-C, a Belize reference strain) (Tumburu et al., 2012). Neonicotinoids such as imidacloprid (IMI) reach surface waters due to spray drift, runoff from crops and soils, leaching, and foliar deposition. Neonicotinoid exposure induces oxidative damage in shrimp, increasing metabolic rate and energy requirement to maintain normal metabolism. The modifications on cellular energy demand as a result of neonicotinoids exposure could also be impacted negatively shrimp growth rates (Butcherine et al., 2019; Osterberg et al., 2012). Studies of areas of the USA contaminated with agrochemicals recorded declines in crustaceans associated with bifenthrin and fipronil (Weston et al., 2015). Twenty-four-hour acute toxicity assays with the organophosphate ® acephate (Orthene ), the carbamate aldicarb, the chloronicotinyl ® imidacloprid (Trimax™), the pyrethroid lambda-cyhalothrin (Karate with ® Zeon Technology), and the glyphosate-based herbicide Roundup Pro on blue crab (Callinectes sapidus) megalopae and J1–J4-stage juveniles demonstrated that all tested compounds caused toxicity to blue crab megalopae, and to juveniles, the relative toxicity in decreasing order ® was λ-cyhalothrin>imidacloprid>aldicarb>acephate¼Roundup Pro. According to the results, λ-cyhalothrin and imidacloprid were the most toxic to developing blue crabs due to metamorphosis-associated mortality at LC20, suggesting the elevated risk on molting juvenile blue crab (Osterberg et al., 2012). The use of insecticides has increased in Australia in areas adjacent to estuaries where the shrimp farms are located, which is a problem because juvenile black tiger shrimp (Penaeus monodon) maintained in estuarine environment are valued for their large size and rapid growth. The risk assessment and toxicity of some insecticides (imidacloprid, bifenthrin, and fipronil) were evaluated in 20 days post-hatch post-larval stage of this species. The results showed that post-larvae were sensitive to all insecticides tested at concentrations to those that cause mortality to other crustaceans. When the environmentally realistic concentration was tested, bifenthrin and imidacloprid reduced the ability of post-larval shrimp to capture live prey (Hook et al., 2018). An increasing number of studies revealed high levels of contamination occurring in fish and seafood for human consumption (Ernst et al., 2018; Martıńez-Go´mez et al., 2012; Perez-Parada et al., 2018; Rodrı´guez- Hernańdez et al., 2017). In the context of contamination, fatty acids and 84 Vania Lucia Loro and Bárbara Estevão Clasen lipid-adjusted concentrations of dioxins, toxaphene, and dieldrin were determined in 459 farmed Atlantic salmon, 135 wild Pacific salmon, and 144 supermarket farmed Atlantic salmon fillets purchased in 16 cities in North America and Europe. The results showed that the levels of omega- 3(n-3) and omega-6 (n-6) fatty acids considering n-3 to n-6 ratio were about 10 in wild salmon and 3–4 in farmed salmon. The supermarket samples were similar to the farmed salmon from the same region. Lipid-adjusted contaminant levels were significantly higher in farmed Atlantic salmon as compared to those in wild Pacific salmon. In conclusion, salmon, especially farmed salmon, are a good source of healthy n-3 fatty acids. However, they also contain high concentrations of organochlorine compounds such as dioxins and toxaphene. The presence of these contaminants may reduce the net health benefits derived from the consumption of farmed salmon (Hamilton et al., 2005). The study of Rodrı´guez-Hernańdez et al. (2017) revealed high levels of organochlorine pesticides (OCPs) in aquaculture products compared to those obtained from fisheries. The results indicated that a consumer who chose to consume only aquaculture products (ΣOCPs 3.36ngkg 1 day 1) would be exposed to the levels of pollutants investigated about twice higher than if this theoretical consumer had chosen only products from capture fisheries (ΣOCPs 1.85ngkg 1 day 1). The levels of OCPs in salmon were ΣOCPs 5.5ngg wet weight 1 in capture fishing samples and ΣOCPs 10.0ngg wet weight 1 when obtained by aquaculture. Sea bass also showed high levels of OCPs (ΣOCPs 3.5ngg wet weight 1) in aquaculture origin and ΣOCPs less than 1ngg wet weight 1 by capture fishing (Rodrı´guez- Hernańdez et al., 2017). In terms of food safety is important verify if aquatic food as salmon is safe for human consumption. The levels of endosulfan, pentachlorobenzene, and hexachlorobenzene in both wild and farmed Atlantic salmon were well below the maximum levels applicable in the European Union (Lundebye et al., 2017). Chatterjee et al. (2015) analyzed the residues of 119 contaminants in pangasius catfish (Pangasianodon hypophthalmus) in India. Pesticides such as malathion, parathion-methyl, and others were detected in 38% of the samples. The authors concluded that the detection of pesticides commonly used in the region indicates their direct application in aquaculture and contamination by agricultural runoff. The authors emphasize the need for continuous monitoring of residues in aquaculture fish. Information from media reports suggests that pangasius catfish is improper to consume because it is able to survive in the “polluted Mekong River.” Due to the heavy polluted situation of this river, it is sup- posed to contain pesticides and veterinary treatment chemicals. To assess the Agrochemicals: Ecotoxicology and management in aquaculture 85 safety of pangasius consumption, a full toxicological risk assessment study was performed based on the EU Rapid Alert System for Food and Feed (RASFF) database. The toxicological risk assessments did not support any of the toxicological risks suggested in the media. The maximum amount of the fillet that could have been consumed without any adverse effects amounted to between 3.4 and 166.7kg day 1 (lifelong for a 70-kg adult) in the case of pesticide contamination and between 0.613 and 303kg fillet day 1 in the case of preservatives and antibiotics. The authors concluded that the consumption of pangasius available on the European market does not pose any concern for the health of the consumer. The analysis presented in this study illustrates that publicly available independent information could help consumers to develop their own well-informed opinion about food safety issues (Murk et al., 2018). Environmental pollutants such as organochlorine pesticides are a global threat to food safety. In particular, the aquatic biota can bioaccumulate many of these contaminants potentially making seafood of concern for chronic exposure to humans. A study conducted with the aim to evaluate trends of contaminant levels in Norwegian farmed Atlantic salmon using an EU-instigated surveillance program, the Norwegian Food Safety Authority (NFSA) between 1999 and 2011 collected more than 2300 samples of Nor- wegian farmed Atlantic salmon (Salmo salar). The fillets of these fish were analyzed for dioxins, PCBs, metals, and organochlorine pesticides. The results showed a decline in the levels of DDT and its metabolites in Norwe- gian farmed salmon from 2002 to 2011. Other organochlorines do not exhibit any time trends since most of the data are below, or close to, the LOQ (Nøstbakken et al., 2015). Similar studies showed the highest abundance of dieldrin and toxaphene in farmed salmon, apart from the sum of DDT (Hites et al., 2004). However, in the study of Nøstbakken et al. (2015), these two pesticides were found in considerably lower amounts and organochlorines are not a problem to safe salmon consumption. Another factor that compromises the safety of aquatic food is the source of ingredients to be used in feeding aquatic cultures. Examples are fish oil and fish meal used in salmon culture. Fish oil (FO) is the main source of persistent organic pollutants (POPs) in fish feed. Actually, the use of vegetable-origin ingredients in fish feed formulations is considered as an alternative to reduce the levels of lipophilic pollutants in farmed species as European sea bass (Dicentrarchus labrax). A study conducted to evaluate the effect of the use of dietary vegetable oils in the farming of European sea bass on the contents of organochlorine pesticides (OCPs) and other POPs in edible fish showed 86 Vania Lucia Loro and Bárbara Estevão Clasen interesting results. A total of 60 sea bass muscle pools were obtained from fish farmed employing six experimental diets, with different percentages of fish oil (6% vs 3%), and fish meal (FM) (20%, 10%, and 5%). The authors did not observe differences in contamination levels of sea bass muscle. However, the fish farmed using feed which had lower levels of fish oil (3%) showed significantly lower levels of OCPs such as sum of DDTs (30.88% of reduction) and sum of chlordanes (42.85% of reduction). The results of this study indicate that the use of ingredients that allow the employment of low percentage of fish oil in feeds helps to reduce the load of several toxic pollutants in the fillets of European sea bass (Gines et al., 2018). A great risk factor to aquaculture species generally is the proximity to agricultural areas. The application of pesticides and fertilizers in upstream areas may also cause water contamination due to the interconnection of areas, especially during the rainy season (Rasmussen et al., 2016). A recent study performed in the South American rivers Uruguai and Negro analyzed the muscle of fish with different feeding habits and commonly used for human consumption. Three different localities were chosen for the study, all associated with agriculture. The results showed that agrochemical residues occurred in 96% of the muscle of fish sampled. Thirty different pesticides were detected at concentrations ranging from <1 to 194μgkg 1, being trifloxystrobin, metolachlor, and pyraclostrobin the most common. Persis- tence and mobility were factors that contribute to the results found (Ernst et al., 2018). A study conducted in Brazil (Saõ Francisco River) verified the presence of 150 different classes of insecticides, fungicides, herbicides, and acaricides in the muscle of Prochilodus costatus, a fish used for human consumption. These fish contained the residues of 17 different agrochemicals and organophosphorus and carbamate pesticides were the most frequent and exhibited highest concentrations. The muscle of P. costatus also presented chlorpyrifos, diazinon, dichlorvos, disulfoton, ethion, etrimfos, phos- alone, phosmet, and pyrazophos in 61% of samples collected (Oliveira et al., 2015). The bioaccumulation of OCPs in two freshwater fish used for human consumption (Oreochromis mossambicus and Clarias gariepinus) in southern Africa exceeded European Commission Maximum Residue Limit (MRL) in most samples. A health risk assessment indicated potential dietary risk related to exposure to heptachlor, heptachlor epoxide, and dieldrin due to high contamination recorded in the fish sampled (Buah-Kwofie et al., 2018). A study conducted in China, where there is a large consumption of fish, demonstrated high levels of OCP residues in freshwater fish, marine Agrochemicals: Ecotoxicology and management in aquaculture 87 fish, and mollusks from Liaoning Province in Northern China. The results showed that the bioaccumulation rate is higher in freshwater fish than marine fish or mollusks, probably because freshwater fish are raised close to agricultural areas that received residues of agrochemicals. The authors calculated human exposure risks and concluded that considering geographical terms, the residents of inland and coastal regions were exposed to the same total OCP concentrations consuming fish or mollusks (Liu et al., 2010). Many lipophilic pesticides accumulate in animal fat and tend to be concentrated in these tissues. Food intake is a source of exposure to these chemicals for the general population (Clasen et al., 2018; Gerber et al., 2016; Sun et al., 2018). Pheiffer et al. (2018) determined OCP residues in C. gariepinus from South Africa and investigated the risk to human health through the consumption of fish. The largest ΣOC pesticides in the fish muscle in three locations (Fleurhof, Orlando, and Lenasia) ranged from 81 to 1190ngg 1. The DDTs were determined to be from historic use, whereas the chlordane (CHL) levels indicated more recent inputs. The authors found that OCPs in C. gariepinus in the study area pose a high risk to human health when consumed and have a cancer risk of up to 1 in 10. Considering the increasing amount of evidence of pesticide occurrence in edible fish tissues, food safety issues could be addressed by regulatory agencies. The MRL frequently used as a parameter for food quality showed some variations according to the legislation of each country and the global differences in pesticide legislation do not guarantee consumer safety. Some countries, such as the Russian Federation and Canada, have established specific MRLs for DDT and lindane in fish and shellfish at 2mgkg 1. In the European Union, if the precautionary principle is applied, agrochemicals should not be in food at levels above 0.01mgkg 1. However, no MRL for pesticides has been regulated for trout, eels, salmon, sardines, herring, tuna, and 61 other fish widely consumed worldwide in the Codex Alimentarius (Perez-Parada et al., 2018). Organophosphate pesticides such as chlorpyrifos have been detected in various rivers, soils, lakes, and seawater and affect algae, aquatic microorganisms, and aquatic organisms (Huang et al., 2020). The pyrethroids also have high toxicity to fish, around 1000 times higher than to mammals. The main symptoms of the chronic toxicity include neurotoxicity, hepatotoxicity, nephrotoxic, and cell apoptosis. Considering the risk assessment, an alternative to reduce the impacts of some pyrethroids is delivering pesticides in nanoforms. This method offers a modern alternative to control agricultural pests and reduce the release of more toxic substances into the environment 88 Vania Lucia Loro and Bárbara Estevão Clasen

(Zhu et al., 2020). Several studies suggest that fish are particularly vulnerable to pyrethroid pesticides that spill into aquatic environments because of the lack of enzymes that hydrolyze this insecticide. In addition, fish are more susceptible to oxidative stress due to pyrethroid exposure because this class of agrochemicals alters the expression and activity of antioxidant enzymes (Yang et al., 2020). The phenylpyrazole fipronil is one of the most efficient insecticides used widely because it targets the arthropod’s nervous system selectively. Fipronil is the unique insecticide that products of its degradation have been considered to be equal or more potent than fipronil itself. A study using similar water quality as from crayfish culture ponds in south Louisiana determined that LC50s for fipronil to red swamp (Procambarus clarkii) and white river crayfish (Procambarus zonangulus) were 14.3 (95% confidence interval—CI: 1 5.1–23.4) and 19.5 (95% CI: 11.1–27.9) μgL , respectively. LC50s of its products of degradation were: fipronil sulfone (11.2; 95% CI: 9.2– 13.2μgL 1), fipronil sulfide (15.5; 95% CI: 13–18μgL 1), and the photo- product desulfinyl fipronil (68.6; 95% CI: 46–95.2μgL 1). In situ experiments using caged crayfish in cultured ponds that receive effluents from drained rice fields with Icon-treated seeds (Icon is a pesticide containing fipronil as active principle) showed 40% of survival compared to 83% in untreated water effluent (Schlenk et al., 2001). In summary, the frequent exposure to agrochemicals increases the chance of accumulation in seafood or aquaculture products as fish, and this fact could be a risk of food contamination. In addition, exposure to agrochemical mixtures could lead to stress in fish, and consequently, stressed fish are more susceptible to effects of particular chemical in water or changes in environmental conditions such as temperature or pH (Perez-Parada et al., 2018). The evaluations of risk are very important to ensure compliance with the MRL established by each country. Unfortunately, some countries do not have the values of MRL for each agrochemical used actually. So all studies cited here and others considering risk assessment and food safety are relevant to verify the degree of environmental contamination and risk to human diseases due to the ingestion of food contaminated with agrochemicals.

5.2.1 Ecotoxicology and biomarkers of agrochemical toxicity Fish, mollusks, and other aquatic organisms have the potential to be environmental bioindicators for agrochemical exposure. Biomarkers are used to measure exposure and biological effects to environmental stressors such Agrochemicals: Ecotoxicology and management in aquaculture 89 as agrochemicals. Some biomarkers specifically measure susceptibility to adverse outcomes from environmental contaminants. The biomarkers that reflect both exposure and biological effects have been widely incorporated in field and laboratory studies to be considered in the decisions to protect environment and aquatic organisms (Hook et al., 2014). Agrochemical exposure can cause various effects, such as oxidative damage in different tissues of aquatic organisms. The effects are measured by biomarkers, protein carbonyl, AChE activity, antioxidant enzymes, and genotoxic indicators. For example, common carp Cyprinus carpio reared for 90 days in an irrigated rice field system using agrochemicals of rice culture showed 10-fold accumulation of lambda-cyhalothrin and 2.5-fold of tebuconazole in the fillet. Biomarker alterations were lipid peroxidation in liver, gills, and muscle. Protein oxidation was detected by increased levels of protein carbonyl formation in liver and muscle tissues (Clasen et al., 2018). Common carps exposed to environmental relevant concentration of quinclorac (344.6μgL 1) in a rice field condition (irrigated rice) for 90 days showed increased levels of protein oxidation in muscle. The residues of quinclorac were recorded in water 42 days after the first application (Toni et al., 2013). Low concentrations of chlorpyrifos (CPF) in the aquatic environment lead to certain toxic effects on fish, crustaceans, and shellfish, including behavioral changes (Yang et al., 2020), oxidative damage (Bonifacio et al., 2017; Patetsini et al., 2013), and genotoxicity (Ismail et al., 2017). Fipronil and carbamate carbofuran are the agrochemicals commonly used in agriculture and frequently cause biochemical alterations in fish. Carps exposed to carbofuran (3.6μgL 1) or fipronil (0.65μgL 1) at concentrations applied in rice crops for 30 days showed inhibition of AChE in brain and muscle (Fig. 5.1A). The exposure to these agrochemicals caused liver lipid peroxidation determined by the increase in thiobarbituric acid-reactive substance (TBARS) levels. There was also an increase in protein oxidation measured by protein carbonyl concentration (unpublished observations of the authors). Oxidative damage was recorded in mussels (Mytilus galloprovincialis) exposed for 30 days to 0.05μgL 1 chlorpyrifos and penoxsulam. Protein carbonyl increased almost twice in chlorpyrifos exposure as compared to penoxsulam exposure and there was also DNA damage and alteration on antioxidant capacity (Patetsini et al., 2013). According to these authors, the parameters could constitute good mussel’s biomarkers to evaluate agrochemical exposure. An induction of protein carbonyl was also observed in deltamethrin-exposed shrimp (gills) at 0.1μgL 1 (Dorts et al., 2009). 90 Vania Lucia Loro and Bárbara Estevão Clasen 1 − 0.3 Control Carbofuran 0.2 Fipronil mg protein 1 −

0.1

0.0 mmol AcSCh min Brain Muscle (A)

12 14 1

− 12

9 protein 1 10 − 8 6 6 4 3 2 mmol MDA mg protein mmol MDA mmol carbonyl mg mmol carbonyl 0 0 (B)(LiverBrain MuscleC) Control Carbofuran Fipronil Fig. 5.1 Acetylcholinesterase (AChE) activity in brain and muscle (A); lipid peroxidation in liver, brain, and muscle (B); and protein carbonyl in liver (C) of common carps (Cyprinus carpio) exposed to carbofuran and fipronil (n¼10) (unpublished observations of the authors).

Sublethal exposure to synthetic pyrethroid cypermethrin disrupts the oxidant/antioxidant balance in the gills of freshwater mussel Unio gibbus, leading to the induction of oxidative stress and increased protein carbonyl levels. Exposure leads to the upregulation of the activities of the antioxidant enzymes superoxide dismutase (SOD) and catalase (CAT) accompanied by increasing H2O2 and protein carbonyl levels. On the contrary, lipid peroxidation due to increased malondialdehyde (MDA) levels confirms damage to lipids due to insecticide exposure (Khazri et al., 2015). In fish, the gills absorb pyrethroids and promote the influx of the insecticide. Oxidative stress is the main mechanism by which pyrethroid pesticides lead to histological changes in fish. Pyrethroids including deltamethrin, cypermethrin, and lambda-cyhalothrin induce ROS production in the gills, liver, and muscle of fish, which leads to lipid peroxidation (Yang et al., 2020). Investigations using clams (Corbicula fluminea)introducedinthemain drainage channels that receive the effluents coming from agricultural fields in the Ebro Delta (NE Spain) during the main growing season of rice (from Agrochemicals: Ecotoxicology and management in aquaculture 91

May to August) evaluated 46 contaminant levels and 9 biomarker responses. The results showed esterase and antioxidant enzyme inhibition. Lipid peroxidation was higher in clams collected from May to August in drainage channels. Principal component analysis (PCA) and partial least squares to latent structure regression (PLS) analysis allowed the identification of endosulfan, propanil, and phenylureas as being the chemical contaminants causing the most adverse effects in the studied species (Dama´sio et al., 2010). The LC50 of deltamethrin for Chinese mitten crab Eriocheir sinensis was 2.319μgL 1 (48h) and 1.164μgL 1 (96h). Crabs exposed to concentrations above 0.293μgL 1 showed antioxidant enzyme inhibition, genotoxicity, and, in addition, lipid peroxidation that is probably the cause of micronuclei increase and DNA damage. The reduced antioxidant response can also be related to increased lipid peroxidation induced by deltamethrin exposure. Another toxic effect was the accumulation of peroxides (Hong et al., 2018). The inhibition of esterases by carbamates and organophosphates is a fact; however, some authors reported AChE inhibition or induction in brain and muscle of fish exposed to other classes of agrochemicals such as the herbicide clomazone (Pereira et al., 2013). Neonicotinoid imidacloprid (0.5 and 2.0mgL 1) exposure (60 days) increased brain AChE activity, but 2.0mgL 1 nitenpyram inhibited brain AChE activity in Chinese rare minnow Gobiocypris rarus (Tiam et al., 2018). Imidacloprid exposure (20, 200, and 2000μgL 1) also inhibited AChE activity in the digestive glands of clam Corbicula fluminea. Gills showed the inhibition only at 2000μgL 1 (Shan et al., 2020). The AChE activity was significantly reduced in brain and muscle tissues of Astyanax lacustris and in the brain of Markiana nigripinnis inhabiting rice field after spraying of a mixture of agrochemicals (herbicide glyphosate, insecticide bifenthrin, and the fungicides azoxystrobin and cyproconazole) (Rossi et al., 2020). Thus, cholinesterase and esterase measurements could be good biomarkers to measure the neurotoxic effects of agrochemicals other than carbamates and organophosphates. Biomarker approaches are not new, with several biomarkers being in use for decades (Lee et al., 1981; Stegeman, 1978), although the acceptance of these approaches by regulatory agencies has been inconsistent. Biomarkers are very important to monitor agricultural residues in the aquaculture species as well as seafood using sentinel species and biomarkers to detect possible toxicity contributing to improve food quality from aquacultural products and reduce environmental contamination. 92 Vania Lucia Loro and Bárbara Estevão Clasen

5.2.2 Toxicity mechanisms of organophosphorus compounds (OPs), organochlorine pesticides (OCPs), neonicotinoids, and pyrethroids 5.2.2.1 Organophosphorus compounds (OPs) The primary mode of action of organophosphorus insecticides is through acetylcholinesterase (AChE) enzyme inhibition. When AChE inhibition occurs, acetylcholine accumulates in the synaptic cleft and causes repeated stimulation of the post-synaptic axon. Brain AChE inhibition greater than 80% was observed in all fish that survivedtomedianlethalexposures to insecticides. The authors emphatize that AChE level depressed by >13% is a clear indication to OP insecticide exposure. Their findings with AChE inhibition in estuarine fish and invertebrates suggest that brain AChE inhibition levels of 20%–70% in fish could be related to OP insecticide exposure (Fulton and Key, 2001). The amino acid sequence of AChE is highly conserved in animals, and in addition, OPs have a covalent binding to the serine-hydroxyl in the active site of the AChE molecule. Organophosphorus compounds are toxic to most groups of animals with some differences in toxicokinetics and toxicodynamics. The commonly known neurotoxic mode of action of chlorpyrifos involves the inhibition of AChE after conversion of chlorpyrifos to chlorpyrifos-oxon by P450-mediated conversion. The inhibition of AChE by chlorpyrifos-oxon results in acetylcholine accumulation and cholinergic hyperstimulation (Casida, 2017; Silva, 2020). Much of the success of anticholinesterase insecticides results from a suitable balance of bioactivation and detoxification by families of CYP450 oxidases, hydrolases, glutathione S-transferases, and others (Casida and Durkin, 2013).

5.2.2.2 Organochlorine pesticides (OCPs) Organochlorine pesticides (OCPs) have been extensively used in the past, but due to both their environmental persistence and neurotoxicity, their use has been banned or greatly reduced in the last decades. The main effects of OCPs are on the nervous system by an interference with different ion channels. DDT and analogs interfere with sodium channels, causing both central and peripheral nervous effects (Moretto, 2018). OCPs are persistent and frequently can bioaccumulate in fat tissues of aquatic organisms such as fish and crustaceans. According to Martyniuk et al. (2020), the OCPs are endocrine disruptors by suppressing male and female reproductive systems and acting by dysregulation of immune functions. The same study referred Agrochemicals: Ecotoxicology and management in aquaculture 93 also to alterations in lipid metabolism due to OCP exposure. In addition, they also act as teratogens (Martyniuk et al., 2020).

5.2.2.3 Neonicotinoids Since being discovered in the early 1990s, neonicotinoids, a group of neu- roactive insecticides derived from natural toxin nicotine, are used worldwide for destroy insects that cause damage to plant crops. Neonicotinoids are also used in veterinary products and biocides for invertebrate pest control in fish farming (Aaen and Horsberg, 2016). Neonicotinoids are neurotoxins that act as agonists binding to the insect nicotinic acetylcholine receptors (nAChRs). The action mode includes the opening of cation channels such as voltage-gate calcium channels. The opening of channels induced by binding of neonicotinoids to nAChRs leads to insecticidal activity. The agonistic action induces continuous excitation of the neuronal membranes producing discharges that result in cell energy exhaustion and neuron death. The neuronal death toll accumulates as more and more chemical molecules bind to other nAChRs until the organism cannot cope with the damage and dies. This binding potency is due to a unique molecular conformation. It is reported that this interaction between neonicotinoids with nAChRs receptors is irreversible (Simon-Delso et al., 2015; Tomizawa and Casida, 2011). The metabolism of the seven major commercial neonicotinoids can be divided into two phases. Phase I is largely dependent on cytochrome P450. The reactions of phase I involve demethylation, nitroreduction, cyanohydrolysis, hydroxylation nitroreduction, and hydroxylation of imi- dazoline and chloropyridine. Phase I metabolites have been found in plants and mammals. The Phase II metabolism of neonicotinoids involves the formation of conjugates, and there are differences between plants and animals. The differences in properties and structures of the nAChRs subunits between insects and vertebrates explain in part the high selectivity of neonicotinoids to arthropods and relatively low toxicity to vertebrates. Neonicotinoid compounds, such as imidacloprid, are specifically designed to alter the normal activity of arthropods including insects and crustaceans (Casida, 2011; Rondeau et al., 2014; Simon-Delso et al., 2015).

5.2.2.4 Pyrethroids Pyrethroid insecticides are widely used as alternatives to traditional pesticides such as organochlorine or organophosphate because of their low toxicity in mammals including humans. However, several studies suggest that fish are particularly vulnerable to pyrethroid pesticides that spill into the aquatic 94 Vania Lucia Loro and Bárbara Estevão Clasen environment. The biological activity of pyrethroids is caused by the depo- larization of neuronal membranes as excessive sodium ions pass through voltage-gated sodium channels because the pyrethroid prevents the closure of the channel and, due to its lipophilic nature, penetrates rapidly through the membranes of gills, skin, and digestive tracts of aquatic organisms. In addition to its interference with sodium channels, oxidative stress is the main mechanism, by which pyrethroid pesticides lead to a histological change in fish (Yang et al., 2020; Zhu et al., 2020). Pyrethroid pesticides, including deltamethrin, cypermethrin, and lambda-cyhalothrin, induce ROS production in the gills, liver, and muscle of fish, which leads to lipid peroxidation. This same mode of action can be applied to crustaceans and mollusks. In addition, fish are more susceptible to oxidative stress because pyrethroid pesticides alter the expression and activity of antioxidant enzymes (Yang et al., 2020). When crabs were exposed to deltamethrin, results shows that this pyrethroid induce oxidative damage due to reduced levels of antioxidant enzymes in this species. The genotoxicity caused is probably due to lipoperoxidation induced by oxidative products and also the accumulation of peroxides in the crab (Hong et al., 2018).

5.3 Mitigation of agrochemicals 5.3.1 Bioremediation (plants, bacterial strains, and actinobacteria) The use of plants to minimize and influence the transformation and transport of insecticides is a good alternative to mitigate the presence of agrochemicals in water and soil, reducing impacts in aquaculture species (Knok et al., 2008). A recent study revealed the ability of wetland plants in reducing neonicotinoid residues during the early to mid-growing period of the Cana- dian prairie cropping season. The authors sampled 20 agricultural wetlands (10 vegetated; 10 unvegetated) situated in the fields of canola production ® using Prosper (Bayer CropScience) near Alvena, Saskatchewan, Canada. Neonicotinoid was detected in some plants sampled and also the reduction during plant growth confirmed that common macrophytes in prairie wetlands have the potential to significantly reduce the annual inflow of neonicotinoids (Main et al., 2017). Another limiting factor concerning the use of plants to detoxify aquatic environments is the absence of information regarding if the agrochemicals could be released back to the environment during plant senescence (Chagnon et al., 2015). Agrochemicals: Ecotoxicology and management in aquaculture 95

Besides plants being good alternatives to reduce the impact of agrochemicals, different organisms could be used as strategies to eliminate or mitigate the effect of agrochemicals. An example is the use of the white-rot fungus Trametes versicolor to degrade the insecticides imiprothrin and cypermethrin, and the nematicide/insecticide carbofuran. The white-rot fungus promoted the degradation of these agrochemicals with minimal inhibition of fungus activity. The results showed the reduction of the toxic effects and support the potential use of T. versicolor to treat diverse agrochemicals (Mir- Tutusaus et al., 2014). A different alternative to mitigate agrochemicals considered the use of Streptomyces rochei strain AJAG7 to control fipronil sulfone biodegradation. Fipronil is used worldwide and recently has been considered highly toxic to insects, being considered one of the pesticides responsible for the decline in honeybee populations. A low-cost powder formulation of strain AJAG7 prepared with agricultural waste was experimented to reme- diate fipronil-contaminated agricultural areas. The results showed and suggest that strain AJAG7 is an ideal option to mitigate fipronil-contaminated soil (Abraham and Gajendiran, 2019). Bacterial strains Achromobacter xylosoxidans (JCP4) and Ochrobactrum sp. (FCp1), isolated from the contaminated agricultural fields, were able to degrade 84.4% and 78.6%, respectively, of soil samples inoculated with 200mgL 1 of chlorpyrifos within 6 weeks (mean rate of degradation was 5.7mgkg 1 days 1) and also promoted the growth of Vigna unguiculata (Akbar and Sultan, 2016). Another alternative to agrochemical mitigation is the use of actinobacteria. Actinobacteria exhibit cosmopolitan distribution in aquatic and terrestrial ecosystems. They play relevant ecological roles including recycling of substances, degradation of complex polymers, and production of bioactive molecules in the environment. This ability is the reason why actinobacteria are good candidates for bioremediation. Strategies such as bioaugmentation, biostimulation, cell immobilization, production of bio- surfactants, design of defined mixed cultures, and the use of plant-microbe systems were developed to enhance the capabilities of actinobacteria in bioremediation (Alvarez et al., 2017). Streptomyces sp. (A2, A5, A11, and M7) consortium isolated from organochlorine pesticide-contaminated soil and sediment samples was able to grow and remove these pesticides from the different tested soil textures and assay conditions (sterile, slurry formulation) (Fuentes et al., 2010). In liquid systems, good microbial growth was observed, accompanied by the removal of lindane (40.4%), metoxychlor (99.5%), and chlordane (99.8%). A different result was recorded in sterile soils. The best result was using clay silty loam soil with the removal rates 96 Vania Lucia Loro and Bárbara Estevão Clasen of 26% (metoxychlor), 12.5% (lindane), and 10% (chlordane) (Fuentes et al., 2017). The studies about bacterial bioremediation offer complementary efficacy to lower environmental residues of agrochemicals. All strategies proposed to mitigate the environmental impacts of agricultural activities involve the analysis of climatic conditions. Considering different strategies to mitigate agrochemicals impacts, Lizotte Jr et al. (2012) examined the efficiency of a managed riverine wetland in mitigating the effects of pesticides discharged in the rivers near crops. They used a mixture of sediment, nutrients (nitrogen; phosphorus; and three pesticides: atrazine, metolachlor, and permethrin) during an experimental agricultural runoff event. This study provided important information about the use of modification in natural wetlands increasing their natural filtering capabilities when inundated with a mixture of nutrients, sediment, and some pesticides currently occurring in agricultural runoff. The modifications of hydrology in riverine backwater wetland can trap within 48h and attenuate 90–98% of sediments, 40–80% of nutrients, and 80–98% of pesticides in a 300-m2 area. In summary, modifications in riverine and wetlands could be efficiently in retention the discharge of agrochemicals to rivers. 5.3.2 Nanomaterials and use of solar energy to mitigate agrochemicals The toxic effects of some pesticides such as OCPs and OPs still persist in the environment in spite of being banned in many countries. Due to their persistence in the environment, toxicity, and potential to bioaccumulation, their removal from the environment is imperative. The use of nanomaterials with the high surface area could be a rapid and effective alternative to binding with chemicals and reducing them in the environment. Rani and Shanker (2018) published a review of the present status of pesticide mitigation using nanoparticles that combined adsorption with photocatalytic and redox degradation. The review described the performance of various nanomaterials with adsorbent, redox, and photocatalytic properties for the effective degradation of pesticides. Most nanoparticles analyzed in the review were able to degrade more than 80% of different OCPs and OPs completely in a short space of time. In the sustainable perspective, green synthetic nanomaterials would be a cheaper and better alternative to control agrochemical toxicity (Rani and Shanker, 2018). Nanotechnology has shown great potential in precision agriculture, with the development of a range of plant protection products that are termed “nanopesticides.” The term nanopesticide is used to describe any pesticide formulation that Agrochemicals: Ecotoxicology and management in aquaculture 97 involves either very small particles of a pesticide active ingredient or other small engineered structures with useful agrochemical properties. Nanoformulations for agricultural purposes may offer benefits due to their greater solubility, mobility, and durability. They also offer the opportunity to reduce the amount of active ingredients used and the possibility to employ the products releasing less harmful chemicals to nontarget organisms, thus reducing the development of resistance. Among biological sources, plant extracts (leaves, flower, stem, and roots) from a diverse range of plant species have been used in synthesizing nanoparticles to use in agriculture as nanopesticides (Baker et al., 2017; Duhan et al., 2017). Another alternative to remediation of agrochemicals in water was a pyramidal-shaped solar system, a new and cheap tool for pesticide removal from water contaminated with herbicides. This pyramid-shaped solar device was tested in field experimental conditions and was able to remove more than 99.95% of both herbicides from a 10-L solution containing 4.160.94g of imazethapyr and 1.310.17g of imazapic. The pyramid solar system was efficient in removing two fungicides from the surface water of a local river in southern Brazil (Hoff et al., 2019). A simplified model of pyramid solar system is shown in Fig. 5.2.

Fig. 5.2 Pyramid solar system scheme used for the agrochemical removal from contaminated effluents (Hoff et al., 2019). 98 Vania Lucia Loro and Bárbara Estevão Clasen

5.4 Agrochemicals banned from use in agriculture and aquaculture Organochlorine pesticides (OCPs) were largely used in the past to control pests and epidemic and unwanted vegetable species. Persistent and bioaccumulative chemicals such as DDT, hexachlorocyclohexane (HCH), toxaphene, aldrin, and dieldrin were banned by the Stockholm Convention approved in 2002 and have been replaced by environmentally and less bioaccumulative chemicals. OCPs are highly toxic and very dangerous due to their persistence in the environment and capacity of accumulation in human and animal fat tissues; due to this fact, some are banned for use. The different legislation types between countries need to be taken into consideration. For example, some compounds that are environmentally persistent, highly toxic, and banned from agricultural use in developed countries remain popular in developing countries such as Brazil (Daam et al., 2019). Agrochemicals, such as acephate and atrazine, both banned in the European Union (EU), are the third and seventh most sold pesticides in Bra- zil, respectively (Bombardi, 2017). The levels of 19 OCPs were analyzed in surface water and sediments from tanks located nearby agricultural areas of Kanyakumari, Tamil Nadu, India. Some banned pesticides such as hexachlorocyclohexanes (HCHs), dichlorodiphenyltrichloroethane (DDT), and heptachlor epoxide were recorded in the sediment as well as surface water ( Jeyakumar et al., 2014). DDT, in spite of being banned in South Africa, is still used in some places for vector control. Some authors verified the presence of OCPs in edible fish species (Omwenga et al., 2016). A study conducted considering shellfish collected in Xiamen, a coastal city in China found DDTs followed by aldrin and endosulfan. HCHs (hexachlorocyclohexanes), chlordanes, heptachlor, and mirex were recorded in the shellfish (Zhang et al., 2012).

5.5 Regulatory process for new chemicals and good agricultural practices Most potential risks to aquatic organisms due to agricultural activity can be controlled using good agricultural practices and reducing agrochemical use. These practices could be a good alternative to control infectious diseases, reduce environmental contamination, and improve food safety. Agrochemicals: Ecotoxicology and management in aquaculture 99

An example of urgency of actions is reported by Ntsama et al. (2018) that conducted a study with 107 farmers in the central, southern, coastal, and western regions of Cameroon. The results revealed that 51% of the farmers used agrochemical or veterinary products without technical assistance. This result shows that most farmers did not have any assistance to decide what chemicals use in farming. Thus, revisions in the regulation process for existing chemicals and new ones for current use are important to guarantee the protection of the environment, wildlife, and human occupational exposure. Pesticide legislation varies worldwide, probably due to different requirements, guidelines, and legal limits for plant protection in different countries. It is known that developed nations have more stringent regulations than developing countries, such as Brazil, where the lack of resources and expertise to adequately implement and enforce legislation is a limiting factor to improve the regulatory process. Global differences in pesticide legislation act as a technical barrier to trade. International parties such as the European Union (EU), Codex Alimentarius Commission (Codex), and North Amer- ican Free Trade Agreement (NAFTA) were created to harmonize pesticide legislation by providing MRLs. Globally harmonized pesticide standards could be a good alternative to establish limits for the safe use of agrochemicals worldwide (Handford et al., 2015). The use of integrated pest management also represents a good option because it emphasizes the growth of a healthy crop with the least possible disruption to agro-ecosystems and encourages natural pest control mechanisms. Since 2013, the World Health Organization by the International Code of Conduct on Pesticide on Pesticide Management recognizes the importance of the elimination of highly dangerous pesticides and recom- mends good agricultural practices (FAO, 2014). Recent reviews identified substantial differences in current legal requirements for assessing the risk to human health from exposure to multiple chemicals between different countries as well as between regulatory sectors (OECD, 2018). Recent issues pointed out the rules concerning neonicotinoids in some European Coun- tries. The European Commission set restrictions on the use of imidacloprid- based products to reduce the risks for bees and the Dutch national authorities issued a reduction of effluent levels of imidaclopid to protect aquatic life. Future monitoring data will ultimately reveal if these measures are sufficient to meet the newly proposed standards (Smit et al., 2015). These studies rein- force the importance of monitoring the environment and aquatic organisms using all tools available to maintain restrictions and protect aquatic and wildlife as bees for example. Actual regulatory mechanisms of the introduction of 100 Vania Lucia Loro and Bárbara Estevão Clasen new compounds for use in agriculture need improvement to prevent damage to aquatic products and wildlife. An example of the flaws is the neonicotinoids’ approval and introduction in agricultural practices in the 1990s after laboratory and field testing that determined the absence of effects to nontarget organisms. However, neonicotinoids clearly are related to reduced populations of bees and interfere with the pollination process made by the bees, for example (Godfray et al., 2014; Van der Sluijs et al., 2013). Other examples of the interference of neonicotinoids causing toxicity to aquatic organisms were explored in this chapter. Mancini et al. (2019) described a possible approach to complement the existing regulatory process that combines long-term and national-scale data sets on native wildlife with agrochemical use data to understand long-term and large-scale impacts of agrochemicals on wildlife populations. Milner and Boyd (2017) proposed the creation of a pesticide surveillance system similar to the pharmacovigilance program developed to ensure the safety of medicines for large-scale use. In a general way, a program for monitoring agrochemicals would improve decisions regarding the approved use of agrochemicals. The purpose also included the creation of a foundation for studying and defining best practices in the regulatory process for currently used and the future’s new agrochemicals to be used. Nowadays, it is necessary to include biomarkers in local organisms that live near to cultivated areas such as sentinel organisms to perform an ecological risk assessment. According to recent approaches, short- and long-term effects of combinations of agrochemicals combined with different environmental conditions are the main topics of the future research and must consider the organisms of different trophic levels (Akoto et al., 2016; Perez-Parada et al., 2018). In summary, some adjustments need to be made to reduce agrochemical impacts. Future studies on aquaculture development should consider the rapid monitoring of trace contaminants in tissues of fish, crustaceans, and mollusks in order to detect levels that could represent a risk for human health. The development of online alert systems mainly in aquaculture sites can be a valuable technolog- ical advancement for the early detection of any chemical or biological per- turbation in the organisms. 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Pharmaceutical pollutants

Helena Cristina Silva de Assis Department of Pharmacology, Federal University of Parana´, Curitiba, Parana´, Brazil

6.1 Introduction Among the compounds released into the environment from the various anthropogenic actions are pesticides, metals, polycyclic aromatic hydrocarbons, nanoparticles, plasticizers, and pharmaceuticals. This problem has been influenced directly by increasing population growth and urbanization (Kelly and Brooks, 2018). Among these micropollutants, in this chapter, the pharmaceuticals recognized as emerging environmental contaminants are highlighted (Barcel and Petrovicm, 2008) and are currently one of the main concerns at the national (ANVISA, 2017) and international levels (Comber et al., 2018; Daughton, 2016; Ebele et al., 2017). Emerging contaminants are defined as any naturally occurring or synthetic chemicals that are not commonly monitored, but have the potential to enter the environment and cause unknown adverse effects (Ferreira, 2014). Several studies have demonstrated the detection of pharmaceuticals in surface waters such as rivers, lakes, and estuaries (Calado et al., 2019; Fent et al., 2006); effluents (Fang et al., 2012); groundwater (Dougherty et al., 2010); and drinking water (Sodre et al., 2010; Sun et al., 2015). In these studies, they are found at low levels, from ng/L to μg/L, and are therefore termed micropollutants (Ferreira, 2014). The pharmaceuticals differ from other bioactive chemical pollutants, such as pesticides or biocides, because they are generally not intended to kill organisms or regulate populations (with some exceptions, e.g., anthelmintics, antibiotics, and fungicides). Instead, many pharmaceuticals are intended as modifiers of physiology and, in some cases, also behavior (Arnold et al., 2014). Furley et al. (2018) reported that understanding the fate and effects of pharmaceuticals in the aquatic environment is one of the research needs for Latin America.

6.2 Pharmaceuticals in the aquatic environment Chemically, the active compounds contained in pharmaceuticals are often formed by complex molecules with different functionalities and physical-chemical and biological properties. In addition to the active substances, the pharmaceutical formulations may incorporate ingredients such as pigments and dyes and may also increase environmental impacts (Koch et al., 2005). Studies have found that the negative effects of pharmaceuticals presented in the aquatic ecosystem can be transferred within the food chain (Oaks et al., 2004; Ribas et al., 2016). An additional factor of the toxicity of a pharmaceutical compound in the aquatic environment is its persistence. Once released into the environment, the compound is transported and distributed to surface waters, sediments, and biota. The actions and concentrations in each of these compartments are determined by a series of factors and processes, including the concentration of the pharmaceutical and its separation in sediments, degradation, and environmental and habitat climatic conditions. Degradation of the substances may occur in a biotic way by aerobic or anaerobic organisms, or in an abiotic form via photodegradation and/or hydrolysis (Boxall et al., 2004; Morley, 2009). Another problem with these compounds is that many of them, espe- ciallythepoorlysolublepharmaceuticals,haveaveryhighbioaccumu- lation potential that poses an even greater risk to the health of aquatic organisms and the human population in general (Crane et al., 2006; Daughton, 2016). Only a few of the pharmaceuticals that are used in human and veterinary therapy that can reach the water bodies are considered a real risk for the environment and public health, because of their volumes of consumption, toxicity, and persistence in the environment. The most diverse classes are found among antibiotics, contraceptives, nonsteroidal antiinflammatory drugs, beta-blockers, antidepressants, antiepileptics, antihypertensives, hypoglycemic, and lipid-lowering agents (Ebele et al., 2017). There is also great concern about the effluents from hospitals, which have a high load when compared to domestic effluents and may contain pharmaceuticals with high toxicity, such as antineoplastic agents (Langford and Thomas, 2009; Santos et al., 2013). Pharmaceutical pollutants 109

6.3 Pharmaceutical sources and pathway to the environment Pharmaceuticals can enter the environment via several different pathways (Boxall et al., 2003). They enter into the aquatic environment after use in human and veterinary medicine and also from the pharmaceutical industry. For the veterinary pharmaceuticals, the most important routes of entry into the environment are likely the discharge of aquaculture waste effluent, the wash-off of topical treatments, and the excretion of substances in urine and feces of livestock (Boxall et al., 2003; Kemper, 2008). The feces of animals as well as contaminated sludge in a sewage treatment plant may be used as fertilizer on agricultural land, contaminating the soil and can suffer leaching or leaching with rain, contaminating the water bodies (Boxall et al., 2012; Nikolaou et al., 2007). The use of pharmaceuticals in aquaculture also becomes a source of contamination released directly into surface water (Lalumera et al., 2004)(Fig. 6.1). Another possibility is that veterinary pharmaceuticals enter the environment as aerosols or dust, but the significance of these releases is unknown. Also, the impacts of emissions from treating pets and disposing of unused or expired products cannot be established but researchers consider emissions via these routes less relevant than emissions from the treatment of aquaculture and agriculture livestock (Blomberg, 2017). The incorrect disposal also generates the presence of pharmaceuticals in the sanitary landfill contaminating the soil and after a process of leaching due to rainfall runoff, contaminates the groundwater, and can reach the surface water (Glassmeyer et al., 2009). In Brazil and some other countries, sewage treatment plants play a key role in the contamination of surface water by pharmaceuticals. These drugs are present in organic matter, and microbial degradation (biodegradation) under anaerobic or aerobic conditions is inefficient to remove a large majority of compounds (Ferreira, 2014; Trudeau et al., 2005). Bio- degradation involves specific enzymatic reactions that vary according to the chemical structures of the compounds, making it extremely variable. Pharmaceuticals, either in the unchanged form or in metabolites, may have long-range effects depending on the physicochemical properties of 110 Helena Cristina Silva de Assis

Pharmaceuticals pollutants

Human pharmaceuticals Veterinary pharmaceuticals

Pharmaceuticals Incorrect industry Aquaculture Livestock disposal

Sewage

AQUATIC AND TERRESTRIAL ENVIRONMENT

Sewage treatment Sludge Agriculture pants contaminants

Fertilization

Sanitary Plants landfill accumulation

Leaching Ground water Surface water

Effects on non-target organims

Fig. 6.1 Major pollution pathways of pharmaceutical residues in the aquatic environment.

the compound and the characteristics of the receiving environment (Ebele et al., 2017). Generally, pharmaceuticals have low volatility and are highly polar hydrophilic in nature; and their transport in the environment occurs mainly through water and by dispersion in the food chain (Caliman and Gavrilescu, 2009). The pharmaceuticals also accumulate in the sediment by sorption and can be released again to the aquatic environment (Agunbiade and Moodley, 2016). Adverse effects were observed in benthic organisms that are continuously exposed to these chemicals inside the sediments, interstitial water, and overlying water (Ebele et al., 2017). Pharmaceutical pollutants 111

6.4 Pharmaceutical exposure effects in nontarget species The presence of pharmaceuticals in water, sediments, and biota causes toxicity to nontarget organisms at different trophic levels, including algae, fish, zooplankton, and other invertebrates (Donnachie et al., 2016; Nikolaou et al., 2007). Due to their physiological characteristics, fish are especially susceptible to absorption and effects of pharmaceuticals. Absorption may occur through dermal and gill surfaces, orally through diet, and transfer of contaminants through the lipid egg reserve (Corcoran and Winter, 2010). The pharmaceuticals are designed to have therapeutic effects in humans and livestock and could impact nontarget animals and ecosystems due to long-term low-level exposure due to their continual release to the environment, which is likely to lead to chronic effects (Arnold et al., 2014). The effects from the mixtures of pharmaceuticals and other stressors are also a concern. Some of the results of the studies about the main pharmaceuticals found in the aquatic environment and the potential effects on aquatic organisms are going discuss here.

6.4.1 Antiinflammatory/Analgesics/Antipyretic drugs The class of nonsteroidal antiinflammatory drugs (NSAIDs) is often detected in water bodies because they are widely used by the population. It is one of the most consumed and most commonly encountered therapeutic classes in waters from hospital effluents (Santos et al., 2013). Among the pharmaceuticals belonging to this class with detection in the aquatic environment are ibuprofen, diclofenac, naproxen, ketoprofen, and acetaminophen (Ebele et al., 2017). They act by reversible or irreversible inhibition of one or both isoforms of cyclooxygenase, COX-1 and COX-2 enzymes, involved in the synthesis of different prostaglandins from arachidonic acid (Corcoran and Winter, 2010). Ibuprofen is a drug belonging to NSAIDs and is widely used as an analgesic, antipyretic, and antiinflammatory. It has been detected in surface waters and wastewater of many countries at a concentration ranging from 0.001 to 117μg/L (Ebele et al., 2017). The toxicity of ibuprofen has been demonstrated in different aquatic organisms. Oxidative stress was observed in zebrafish exposed to sublethal concentrations (0.0001–25mg/L) for 28 days (Bartoskova et al., 2013) and also in common carp exposed for 112 Helena Cristina Silva de Assis

12–96h at 17.6mg/L (Islas-Flores et al., 2014). DNA damage and oxidative stress were observed in bivalves (Dreissena polymorpha) exposed for 96h to environmentally relevant concentrations (Parolini et al., 2011) and in Daphnia magna exposed for 48h at 2.9mg/L (Go´mez-Oliva´n et al., 2014). Hematological changes were observed in Indian major carp (Saravanan et al., 2012) exposed to 14.2mg/L over 35 days. Studies with zebrafish embryos showed that ibuprofen is metabolized and excreted in a similar way to mammals ( Jones et al., 2012) and interferes in early development, organogenesis, larval growth, and survival (David and Pancharatna, 2009). Ibuprofen also affected embryo locomotor activity and caused neurotoxicity in Danio rerio at 5, 50, and 500μg/L in 6–96h postfertilization exposure (Xia et al., 2017). In teleosts, prostaglandins play important physiological roles, including actions on reproduction (Hayashi et al., 2008). They are involved in processes such as oocyte maturation, ovulation (Lister and Van Der Kraak, 2008), regulation of gonadal steroidogenesis (Wade and Van Der Kraak, 1993), and in the induction of male and female sexual behavior (Sorensen and Goetz, 1993). NSAIDs cause a decrease in prostaglandins and have been shown to cause endocrine-disrupting effects in fish (Groner€ et al., 2017; Guiloski et al., 2017a). Studies with Brazilian fish species Rhamdia quelen, a freshwater Neotrop- ical species, exposed to 0.1, 1, and 10μg/L of ibuprofen for 14 days demonstrated damage in the kidney, but not in the liver and gills, evidenced by biochemical and osmoregulation biomarker responses. The changes did not cause genotoxicity, but perhaps long-term exposure could cause DNA damage (Mathias et al., 2018). Other studies with NSAIDs also showed kidney damage (Ghelfi et al., 2016; Mehinto et al., 2010; Pamplona et al., 2011). This can be related to the mechanism of the action of NSAIDs in humans that inhibit the synthesis of prostaglandins; and consequently, affect renal homeostasis, particularly renal glomerular filtration (Fent et al., 2006). In fish, it was demonstrated that NSAIDs decreased COX 1 and COX 2 expression in the posterior kidney (Mehinto et al., 2010) and decreased prostaglandins (Bhandari and Venables, 2011; Morthorst et al., 2013). Therefore, posterior kidney damage observed by Mathias et al. (2018) can be related to ibuprofen effects on prostaglandins and changes in renal homeostasis. Hematological biomarkers showed an immunosuppressive effect in groups exposed to 0.1 and 1.0μg/L of ibuprofen. A decrease in cell numbers of the immune system predisposes fish to diseases because it alters their ability to combat pathogens. Furthermore, in the cells from freshwater mussel Dreissena polymorpha ibuprofen increased the percentage of hemocytes showing apoptosis and Pharmaceutical pollutants 113 lysosomal membrane destabilization suggesting that ibuprofen was toxic to these cells (Parolini et al., 2011). Paracetamol (acetaminophen) is one of the most commonly used analgesic and antipyretic drugs around the world available alone and also in combination with other medicines in products used to treat colds and pain (Lau et al., 2016). The mechanism of action is complex and includes the effects of both the peripheral (cyclooxygenase inhibition), and central (cyclooxygenase, sero- tonergic descending neuronal pathway, L-arginine/NO pathway, and canna- binoid system) and antinociception processes ( Jozwiak-Bebenista and Nowak, 2014). Paracetamol overdose can lead to hepatotoxicity in humans and experimental animals (Beger et al., 2015). Some of the effects are related to the metabolite N-acetyl-p-benzoquinone imine (NAPQI), which may further cause impaired development of the fetus and the newborn child (Brune et al., 2015). Because of this risk, the Food and Drug Administration (FDA) has suggested a reduction in the maximum daily dosage of paracetamol from 3900–4000 to 3000–3250mg in order to reduce the potential of liver failure (Krenzelok and Royal, 2012). Studies with R. quelen exposed to 0, 0.25, and 2.5μg/L of paracetamol during 21 days demonstrated an anti-androgenic effect, since a decrease in testosterone levels of male exposed fish and an increase in the levels of estradiol were found at the higher concentration (Guiloski et al., 2017a). The gonadal morphology showed that spermatogenesis was affected in some of the animals exposed to 2.5μg/L of paracetamol. Since the animals had similar weight and length and were at the same stage of maturation, it is suggested that due to the testosterone reduction and estradiol increase, spermiogenesis was affected in some animals of this group (Guiloski et al., 2017a). These effects can compromise the reproduction of the species by affecting the production of germ cells, since the paracetamol was listed as a potential endocrine-disrupting pharmaceutical by Caliman and Gavrilescu (2009). Diclofenac is largely regarded as an environmental toxicant in recent times, after accidental exposure via their food chain lead to massive mortal- ities in vulture species on the Asian continent. The use of diclofenac for veterinary purposes was banned in India (Naidoo and Swan, 2009). The analysis of vultures determined that the death of the animals was caused by kidney failure caused by diclofenac (Oaks et al., 2004). In the aquatic environment, diclofenac is suspected of causing damage in nontarget organisms. The weighted-average concentrations reported in surface waters exceed the predicted no-effect concentration (PNEC) of 114 Helena Cristina Silva de Assis

0.1μg/L in 12 countries worldwide, indicating an unacceptable risk in terms of regulatory environmental risk assessment (Kuster€ and Adler, 2014). Diclofenac has been included in the list of priority hazardous substances by the European Commission in 2012 (European Commission, 2013) with a maximum allowed concentration of 100ng/L in drinking water. Some of the adverse effects of diclofenac in humans, such as nephropathy, are due to the inhibition of prostaglandin synthesis. Other studies have reported cytological and histological effects (Mehinto et al., 2010; Memmert et al., 2013) and on gene expression (Cuklev et al., 2012). Exposure of R. quelen to 0, 0.2, 2, and 20μg/L diclofenac for 21 days caused oxidative stress and affected the testosterone levels in male fish (Guiloski et al., 2017b). Oxidative stress and genotoxicity in the liver were not observed in juvenile R. quelen exposed to diclofenac for 96h (Ghelfi et al., 2016). Subchronic exposure (21 days) to environmental concentrations of diclofenac had more pronounced effects than acute exposure (96h). Oxidative stress status of Cyprinus carpio exposed to diclofenac was affected during the initial days of the study (at 4 days), exhibiting an increased response at 24 days in the liver (Saucedo-Vence et al., 2015). Diclofenac-induced DNA fragmentation in hemocytes of the bivalve D. polymorpha exposed for 1h (Parolini et al., 2009), but caused no change in exposure for 96h (Parolini et al., 2011). Industrial effluent containing NSAIDs induced geno- and cytotoxicity in C. carpio (SanJuan-Reyes et al., 2015). The available data about genotoxicity of diclofenac in fish is not conclusive and seems to depend on the exposure time, species, and concentration. Danio rerio males and females exposed to 10 and 100μg/L diclofenac for 14 days did not present any change in the levels of sex steroids, although other NSAIDs such as ibuprofen caused endocrine disruption, raising levels of estradiol and testosterone in females and reducing testosterone levels in males ( Ji et al., 2013). In invertebrates (Daphnia magna and Ceriodaphnia dubia), NSAID residues were relatively higher (median¼20.50ng/g) compared with other classes of pharmaceuticals. The most frequent NSAID detected was diclofenac with a median of 15ng/g, followed by ibuprofen with a median of 83.65ng/g and celecoxib with a median of 24ng/g. In general, while they show relatively higher measured internal concentrations, the risks NSAIDs pose to invertebrates have been reported to be low (Constantine and Huggett, 2010; Miller et al., 2018). Dipyrone sodium or metamizole sodium belongs to the NSAIDs family and is used as an analgesic, antipyretic, and antiinflammatory, albeit its Pharmaceutical pollutants 115 antiinflammatory efficacy has been disputed, compared to its intense analgesic and antipyretic properties (Souza et al., 2002). Its use is controversial and has been forbidden in many countries, such as the United States, the United King- dom, and Sweden, for its relationship with many blood dyscrasias. Due to its strong analgesic effect, availability as parenteral formulation and low cost, the use of dipyrone is still on the rise in Europe and South America. Dipyrone is a prodrug, and after its oral ingestion, it is rapidly hydrolyzed into its main metabolite, 4-methylaminoantipyrine, from which many others are produced by enzymatic reactions (Ergun€ et al., 2004; Pamplona et al., 2011). The presence of its metabolites is also reported in effluents and surface water by many authors (Feldmann et al., 2007; Go´mez et al., 2008; Wiegel et al., 2004). Fish R. quelen exposed to 0, 0.5, 5, and 50μg/L of dipyrone for 15 days showed a decrease of hematocrit, red blood cells and thrombocyte counts at all tested concentrations, and DNA and kidney damage were observed at the lowest concentration. The genotoxicity can be caused by N-nitrosodimethylamine (NDMA), an N-nitroso compound (NOC). These kinds of compounds are genotoxic, and there is evidence that dipyrone may be transformed into NDMA (Brambilla and Martelli, 2007). Dipyrone sodium is considered by many authors a DNA-damaging drug. As reviewed by Brambilla and Martelli (2007, 2009), it is one of the many drugs that can go through an endogenous transformation into the genotoxic and carcinogenic compounds. Those metabolites are also toxic for the posterior kidney with important consequences to excretory systems in fish or other similar systems in vertebrates (Silva and Martinez, 2007).

6.4.2 Hormones Hormones belong to a group of chemicals called endocrine disruptors that are defined by the United States Environmental Protection Agency (USEPA) as “an exogenous agent that interferes with the synthesis, secretion, transport, binding, action, or elimination of the natural hormones in the body that are responsible for maintaining homeostasis, reproduction, development and/or behavior” (USEPA, 1997). Environmental agencies and nongovernmental organizations classify the endocrine disruptors into three main classes: natural estrogens, synthetic estrogens, and xenoestrogens. The hormones estradiol, estrone, and estriol are the three main endogenous estrogens. Progesterone and testosterone are examples of other natural hormones that, like estrogens, are excreted daily in urine and can be detected in waters that receive large amounts of domestic sewage (Shore and 116 Helena Cristina Silva de Assis

Shemesh, 2003). The 17 α-ethinylestradiol (EE2) is a synthetic estrogen found in oral contraceptives, commonly used by women to prevent possible pregnancy, as well as in postmenopausal hormone replacement therapy. Due to the widespread use of this compound, and consequent excretion, there is an increase in its amount introduced into the environment (Laurenson et al., 2014). This hormone has a chemical structure similar to that of 17β-estradiol (E2), a natural hormone produced by oviparous females at the time of reproduction. Thus, it is capable of binding with high affinity to the receptor. The mechanism of action depends on the chemical structure of each molecule, as well as the estrogenic potential and, consequently, the observed effects on organisms. Some endocrine-disrupting pharmaceuticals have been found to have adverse effects on wildlife at very low concentrations, such as feminizing male fish, preventing reproduction, or triggering population collapse (Kidd et al., 2007). One widely studied physiological response likely to impact individual fitness in wildlife is vitellogenin induction. Vitellogenin (VTG) is required for egg-yolk formation, and it is essential for developing embryos and larvae, but overproduction in females can lead to impaired liver and reproductive function, and that in males reduced kidney function and survival (Arnold et al., 2014). The fish fathead minnow Pimephales promelas was exposed for a year to a concentration of 3.2ng/L EE2 and induced vitellogenin production. Al- Odaini et al. (2013) monitored the occurrence of four synthetic hormones (17α-ethinylestradiol, norethindrone, levonorgestrel, and cyproterone acetate) in five treatment plant effluents, and in Langat River water, Malaysia. The results confirmed literature data on the low removal efficiency of these compounds in treatment plants and their presence in the monitored river water. The release of hormonally active substances in water bodies, even at low concentrations, can promote a serious impact on the dynamics and structure of aquatic populations. At concentrations of 5–6ng/L, EE2 has been demonstrated to cause population collapse of fathead minnow as a result of the feminization of male fish in a Canadian whole-lake experiment (Kidd et al., 2007). The feminization of fish from estrogenic pollution of water bodies has already been reported for several countries worldwide (Harris et al., 2011). According to the measured environmental concentration database, the maximum EE2 concentrations reported in surface waters exceed the predicted no-effect concentration of 0.01ng/L (Caldwell et al., 2012)in28 countries worldwide, also raising the possibility of at least temporal adverse ecotoxicological effects on the local fish population at hot spots. Pharmaceutical pollutants 117

Synthetic sex hormones present in low concentrations in surface water can induce, in long term, the feminization of fish. These compounds promote the reduction of the testosterone concentration, because they act as endocrine disruptors, i.e., acting on the development and growth of animal gonads. Yamamoto et al. (2017) verified VTG levels in male fish from the Iguacu River, South Brazil, as well as decreased levels of VTG and estradiol in the plasma of female fish. These findings were associated with immature gonads and lower gonadosomatic index in Geophagus brasiliensis adult females from a reservoir in South Brazil. This indicates the presence of endocrine disruptor compounds with estrogenic action, probably from discharges of urban, industrial, or agricultural waste. In studies on immune function, rainbow trout (Oncorhynchus mykiss) exposed to E2 showed reduced survival following challenge with a pathogen (Wenger et al., 2011); thus, xenoestrogens can increase susceptibility to disease. In a study on migration ability, Atlantic salmon (Salmo salar) injected with E2 were released into a river and their arrival at a downstream migration point was measured (Madsen et al., 1997). Xenoestrogen-treated fish were delayed by approximately one week in reaching the migration point and had significantly increased mortality compared to control fish. The effects triggered by hormones in the environment range from microinvertebrates to large vertebrates, being widely reported in the scientific literature and therefore considered as a matter of global concern (IPCS, 2002).

6.4.3 Antidepressants and other psychoactive pharmaceuticals Psychoactive pharmaceuticals are designed to alter behavior and have side effects that could also influence fitness-related traits in free-living animals (Bean et al., 2014; Brodin et al., 2014; Fong and Ford, 2014). Fluoxetine (FLX) and its active metabolite norfluoxetine is one of the most commonly detected pharmaceuticals in wastewater and bioaccumulates in wild-caught fish, especially in brain, liver, and muscle tissues (Mennigen et al., 2010). It is usually detected in the range below 1μg/L. It has raised concerns over potential disruptive effects of neuroendocrine function in teleost fish, because of the known role of serotonin (5-HT) in the modulation of diverse physiological processes such as reproduction, food intake and growth, stress, and multiple behaviors. According to Brooks et al. (2005), fish collected in urban water bodies, wastewater receivers in North Texas, USA, showed fluoxetine and sertraline values greater than 0.1μg/g in all examined tissues. 118 Helena Cristina Silva de Assis

The most commonly used antidepressants, selective serotonin reuptake inhibitors (SSRIs), and serotonin-norepinephrine reuptake inhibitors (SNRIs), act via the serotonin and norepinephrine reuptake transporters and interact with other parts of the serotonin system (Corcoran and Winter, 2010). Serotonin levels influence both physiology and behavior in a wide range of organisms, including fish, and play a pivotal role in activity, aggression, and reproductive behaviors (Kreke and Dietrich, 2008). Antidepressants have been shown to reduce territorial aggression in coral reef fish (Perreault et al., 2003) and locomotion and aggression in Siamese fighting fish (Kohlert et al., 2012). Therefore, another SSRI, citalopram, even at concentrations a thousand times higher than in the previous studies did not affect rainbow trout (Holmberg et al., 2011), highlighting substance- specific effects of and species-specific responses to SSRIs. Some studies indicated that fluoxetine is pharmacologically active in fish species exerting anorexigenic effects. Fluoxetine reduces the feeding rate in goldfish (Mennigen et al., 2010). While this effect was found at rather high concentrations, more than 1mg/L, studies have also found that fathead minnow (Stanley et al., 2007) and hybrid striped bass (Bisesi et al., 2014) expe- rienced reduced feeding rates after exposure to 3.7 and 250μg/L of fluoxetine and venlafaxine, respectively. Silva de Assis et al. (2013) observed that the VTG in the plasma of the goldfish co-treated with EE2 plus FLX was increased compared to the other groups. The hepatic VTG mRNA levels and the plasma VTG protein levels were therefore correlated. The increased number of proteins in the plasma of goldfish exposed to FLX that were related to endocrine system disorders also supports the hypothesis that FLX has an ability to act upon endocrine systems. The Japanese Medakas exposed for four weeks to different concentrations of fluoxetine (0.1, 0.5, 1.0, and 5.0mg/L) had their fertility affected, and the developing embryos showed several abnormalities such as edema, curved spine, and incomplete development (no pectoral fins, reduced eyes), and the number of changes in the development of these fish was 4–5 times higher than in the control group (Foran et al., 2004). Fluoxetine prevents the increase of cortisol in fish in response to physical stressor stimulus (Abreu et al., 2017) and blocked cortisol response to acute chasing stress in a dose-dependent manner (Abreu et al., 2014), as well as in fish subjected to different forms of housing (Giacomini et al., 2016). According to Flaherty et al. (2001), Daphnia magna exposed for 30 days to 36μg/L of FLX had a significant increase in reproductive rates, and Pharmaceutical pollutants 119 according to Fong (2001), SSRIs can enhance spawning and oocyte maturation in some bivalves and crustaceans. Zebrafish has been used as a model to evaluate the effects of anxiolytic, antipsychotic, and antidepressant drugs on behavioral, biochemical, and molecular parameters (Marcon et al., 2016). Kalichak et al. (2016) showed that fluoxetine, diazepam, and risperidone (RISP) affected the initial development of zebrafish. All drugs increased mortality rate and heart frequency and decreased larvae length. The overall results pointed to the potential of these drugs to cause a negative impact on zebrafish initial development and, since the larvae viability was reduced, promote adverse effects at the population level. It is possible that eggs and larvae absorbed the drugs and the effects on early development may have significant environmental implications. The presence of RISP residues in water can alter the exploratory behavior of zebrafish embryos and larvae. In fact, during the first-time win- dow (0–6min) in the open field test, aversive stimuli and light/dark tests, larvae exposed to 0.0003μg/L RISP increased the immobility time. This low RISP concentration was already detected in natural aquatic environments, indicating the potential risk for populations exposed to this type of contaminant (Kalichak et al., 2017). Idalencio et al. (2015) showed that acute exposure to 170μg/L RISP impaired the stress axis response, since the exposed zebrafish had lower cortisol levels than control fish, when exposed to an acute stress challenge. On the other study, individual zebrafish were evaluated according to their reactivity to novelty and investigated whether fluoxetine modulated these personality traits. Clear differences in behavior in response to fluoxetine were shown and the data reinforced that fluoxetine modulates studied fish personality traits (Fior et al., 2018). Fluoxetine also affected endocrine (Abreu et al., 2014, 2016) and behavioral responses to stress in zebrafish (Giacomini et al., 2016). The study by Mennigen et al. (2010) showed that male goldfish exposed to waterborne fluoxetine had disrupted sperm release. In the same animals, blood levels of testosterone were decreased, whereas E2 levels were significantly increased. Other indirect ecological effects of pharmaceutical exposure in the aquatic environment may arise through changed population sizes (especially extinctions) and subsequently altered community composition and species richness, as these are known to influence ecosystem functioning (Kidd et al., 2014). Such effects may be especially probable if different taxa respond differently to the exposure. 120 Helena Cristina Silva de Assis

Carbamazepine (CBZ) is also among the most common pharmaceutical residues detected in water bodies. CBZ, which is used to treat epilepsy and bipolar disorder, seems to be ubiquitous in sewage-contaminated ecosystems, dominating samples taken from different matrices, species, and at all trophic levels examined (Du et al., 2014). Approximately 72% of orally administered drug is absorbed while 28% is not metabolized and is excreted unchanged in the feces. The metabolites 10,11-dihydro-10,11-expoxycarbamazepine (CBZ-epoxide) and trans-10,11-dihydro-10,11-dihydroxycarbamazepine (CBZ-diol) are as active as the original molecule, and they are present in water bodies at a similar amount and concentration to the parent drug. Brodin et al. (2014) explored how behavioral modification in predators and prey exposed to psychoactive medication can result in alterations to aquatic food chains and ecosystems under different environmental scenarios. The effects of CBZ tested in vitro using the sperm of common carp revealed that the number of mobile spermatozoids and their movement speed decreased. In a review about the pharmaceutical exposure in aquatic fauna (Miller et al., 2018), it was described that the most frequently reported concentrations for invertebrate are to antibiotics and antidepressant therapeutic classes, with concentrations determined ranging from 0.20 to 320ng/g for the antidepressants. Within the antidepressant class, the SSRIs showed relatively higher internal concentrations when compared to the tricyclic antidepressants and the other antidepressant classes including benzodiazepines. Therefore, the data suggest that the SSRI class of antidepressants may have a higher risk potential than other antidepressants. Benzodiazepines are drugs indicated for the treatment of anxiety, emo- tional and sleep disorders, and epileptic seizures. They have also been used as centrally acting muscle relaxants and as analgesia-inducing. Diazepam (DZ) is the best-known drug in this therapeutic class and also the most widely studied as an environmental contaminant. The presence of DZ has been detected in hospital wastewater and also in the effluents from municipal wastewater treatment plants. It has also been found in drinking water in the concentration of 23.5ng/L. Studies conducted in wastewater treatment plants of Germany detected concentrations up to 0.04mg/L. Calisto and Esteves (2009) and Calamari et al. (2003) reported environmental concentrations from 0.04 to 0.88mg/L for DZ. Abreu et al. (2014) showed that acute exposure to DZ (0.88, 16, and 160mg/L) and FLX (1, 25, and 50mg/L) diluted in water impaired the stress axis function, as drug-exposed fish had lower cortisol levels than control fish when exposed to an acute stress test. A fish with an impaired stress response loses its ability to maintain Pharmaceutical pollutants 121 homeostasis against stressors by reducing the ability to promote ionic, metabolic, and behavioral adjustments necessary for the stress response (Barcellos et al., 2007, 2011, 2016; Cericato et al., 2009). Pascoe et al. (2003) demonstrated that acute and chronic exposures of the cnidarian H. vulgaris to 10 commonly prescribed pharmaceuticals did not present an acute lethal risk or adversely affect feeding or bud formation at concentrations up to 1.0mg/L. Therefore, DZ inhibited polyp regeneration at 10μg/L.

6.4.4 Antibiotics The occurrence of antibiotics in the environment is a human health hazard due to the rise of antimicrobial resistance (AMR). The occurrence of trace antibiotics in animals (and other environmental compartments) could drive selection pressures for mobile genetic elements associated with drug resistance (Wellington et al., 2013). Thus, measuring antibiotic occurrence in biota could also help the surveillance of AMR (Le Page et al., 2017). According to Jones et al. (2002), antibiotics could be classified as extremely toxic to microorganisms (EC50 below 0.1mg/L) and very toxic to algae (EC50 between 0.1 and 1mg/L). Most antibiotics used in veterinary medicine are aimed at preventing and treating diseases in livestock production or aquaculture. Even considering their use at subtherapeutic concentrations, many studies suggest the development of bacterial resistance and further potential appearance of cross-resistance between different classes of antibiotics shared with humans (Thacker, 2005). Antibiotics used in livestock production are excreted in the urine and feces of animals and often appear in manure. Rodrigues et al. (2016) studied the toxicity of erythromycin in rainbow trout, in a short (96h) and subchronic study (28 days), and observed changes in the antioxidant system and lipoperoxidation in the gills and genotoxicity in the liver. Monteiro et al. (2016) investigated the relationship between antibiotic residues found in the muscle of cage-farm-raised Nile tilapia (Oreochromis niloticus), the occurrence of resistant bacteria, and the sanitary practices adopted by farmers in a reservoir in south Brazil. Only three antibiotics (oxytetracycline, tetracycline, and florfenicol) were detected in the muscle of Nile tilapia, and their residues were the highest in small fish; however, the multiple antibiotic resistance (MAR) index was higher in large fish. In addition, a direct positive relationship between the MAR index and the concentration of antibiotic residues in Nile tilapia was found. 122 Helena Cristina Silva de Assis

A literature survey covering 236 published reports from 41 countries showed that water contamination by pharmaceuticals including antibiotics was extensive due to their widespread consumption and subsequent disposal to rivers (Hughes et al., 2013). Among antibiotic measurements, the highest detected concentrations were 2390ng/g for erythromycin (Zhao, 2015) and 1600ng/g for ormetoprim (Meador et al., 2016). Median values for macrolides, quinolones, and sulphonamides were 3.60, 5.23, and 7.35ng/g, respectively. The other class of antibiotics quantified in fish was the tetracycline compounds, oxytetracycline (50ng/g), and chlortetracycline (160–590ng/g). In the review by Hughes et al. (2013), the maximum concentration reported for an antibiotic was 6.5mg/L for ciprofloxacin while sulfapyridine had the highest mean detection frequency. Median concentrations of most antibiotics were in the ng/L range, but at least four of the antibiotics listed here had median concentrations >1μg/L. In other studies, the maximum concentration determined among the antibiotic classes was from the compound sulfamethazine, which reached up to 430ng/g (Dodder et al., 2014). Toxicity of selected sulfonamide antibiotics (sulfamethazine included) has been reported in crustaceans with EC50 levels often in the mg/L range (Bialk-Bielinska et al., 2011; De Liguoro et al., 2009). The macrolides and quinolones showed maximum measured concentrations of 132 and 170ng/g, respectively. However, in general, these classes of compounds showed low occurrence in invertebrates with median concentrations of 2.32 and 10.60ng/g. Available effect data for the quinolones are limited, but a study involving the toxicity of ciprofloxacin to fish and invertebrates showed toxicity thresholds of mg/L concentrations on a range of endpoints (mortality, growth, and reproduction) (Martins et al., 2012).

6.4.5 Antiparasitic pharmaceuticals Antiparasitic pharmaceuticals are an important group that is widely used for veterinary purposes. They are administrated to animals to treat or prevent parasitic infections and can be administrated topically, orally, or injected intramuscularly/subcutaneously. The major route of entry into the environment is probably via excretion in urine and/or feces. The antiparasitic compounds and/or their metabolites are then released in the environment either directly from livestock or by the disposal of manure and slurry onto agriculture land. The antiparasitic compounds can reach aquatic environments by runoff to surface water and thereby present a risk for aquatic organisms. Pharmaceutical pollutants 123

Antiparasitic pharmaceuticals such as benzoylphenylurea insecticides and pyrethroids are also used in aquaculture to treat parasitic infections in fish kept for food production. In a Norwegian study (Langford et al., 2014), antiparasitic compounds were detected in wild shrimps at levels at which chronic effects are seen in experimental studies on this species. This suggests that antiparasitic pharmaceuticals from aquaculture are a potential risk to these animals. Some such as malachite green were banned for having carcinogenic and genotoxic effects on humans. One side effect is the accumulation of residues in fish tissues (Srivastava et al., 2004). Yoshimura and Endoh (2005) tested the acute toxicity of five antiparasitic pharmaceuticals used in the veterinary field: amprolium hydrochloride, bithionol, levamisole hydrochloride, pyrimethamine, and trichlorfon. They were tested for acute toxicity on three aquatic species: Oryzias latipes, Daphnia magna, and Brachionus calyciflorus. The least toxic substance to these organisms was amprolium hydrochloride, the LC50 concentration ranged from 227mg/L for the most sensitive species to >600mg/L for the most tolerant one. Bithionol was the most toxic substance for O. latipes (LC50 of 0.24mg/L). Daphnia magna was the most susceptible species to trichlorfon (EC50 0.00026mg/L) while B. calyciflorus was the most susceptible species to bithionol (EC50 0.063mg/L).

6.5 Final considerations The problem concerning the presence of pharmaceuticals and their metabolites in the aquatic environment is significant. The results from various investigations on water bodies in many countries in recent years, as well as some data emerging from ecotoxicological studies on nontarget species, showed that they may alter aquatic ecosystem equilibria. In fact, the larger use of both human and veterinary pharmaceuticals produce an increase in the release of pharmaceuticals into water. This emphasizes the need for further analysis of environmentally relevant mixtures to understand the full impact of sewage effluents on aquatic organisms. There are a number of uncertainties associated with the environmental risk assessment of pharmaceuticals due to the lack of knowledge concerning their fate in wastes and the environment; their uptake, metabolism, and excretion in wildlife; and their target affinity and functional effects in nontarget species. The bacterial resistance to antibiotics is also a concern for public health, and there is a need to investigate and to improve the management of waste. 124 Helena Cristina Silva de Assis

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Oil and derivatives

Helen Sadauskas-Henriquea, Luciana Rodrigues Souza-Bastosb, and Grazyelle Sebrenski Silvac aUniversidade Santa Cecı´lia (Unisanta), Laborato´rio de Biologia de Organismos Marinhos e Costeiros (LABOMAC), Santos, Brazil bInstituto de Tecnologia para o Desenvolvimento (LACTEC), Laborato´rio de Toxicologia e Avaliac¸a˜o Ambiental, Ambiental, Curitiba, Brazil cDepartamento de Morfologia, Universidade Federal do Amazonas (UFAM), Manaus, Brazil

7.1 Oil and derivatives and the aquatic contamination In this chapter, oil and derivatives include all petroleum products like gasoline, distillates (diesel fuel and heating oil), jet fuel, waxes, lubricating oils, and asphalt. Petrochemical feedstocks, like plastic products, will not be considered here. Among the chemical contaminants present in water, oil and derivatives are a significant pollutant group because they are ubiquitous in aquatic environments (Al-Kindi et al., 2000; Connell et al., 1980; Eisler, 1987; Hameed and Al-Azawi, 2016; Irie et al., 2011; McKeown and March, 1978; Medeiros et al., 2017; Rodrigues et al., 2010; Santos et al., 2016; Simonato et al., 2008). The human demand for energy has been increasing during recent decades because of the exponential growth of the human population, industrialization, and modernization of human life. From heating oil to skin creams, the petroleum industry manufactures thousands of products that are essential to our lives (Val and de Almeida-Val, 1999). The exploitation, commercialization, and industrial development due to this activity have negative environmental impacts (Martinez and Colacios, 2017; Teixeira et al., 2012). The pollution of the aquatic environment due to large or small accidental oil spills is of worldwide concern. Since the 1970s until the end of 2018, approximately 2000 small (7–700tons) and large (>700tons) accidental petroleum and derivative spills were reported in aquatic environments by the International Tanker Owners Pollution Federation (ITOPF, 2019). Typically, these accidental oil spills are reported mainly for marine environments, such as Exxon Valdez on the Alaskan coast in 1989; Prestige on the Spanish coast in 2002; the Deepwater Horizon in the Gulf of Mexico in

2010; the Keystone XL spill in the United States in 2017 (Franco et al., 2006; Harlow et al., 2011; Jung et al., 2011; Kirby and Law, 2010; Lawlor and Gravelle, 2018; Martı´nez-Go´mez et al., 2006; Merhi, 2010; Nkpaa et al., 2013; Sole et al., 1996); the Sanchi collision in the East China Sea in 2018 (Yin et al., 2018); and the oil spill in Pernambuco, Northeast Brazil in 2019 (Carmo and Teixeira, 2020; de Arau´jo et al., 2020; Pena et al., 2020; Ribeiro et al., 2020). However, accidental oil spills were also reported in freshwater ecosystems, such as the spill caused by the rupture of pipelines in Parana´ state (Brazil) in 2000, 2001, and 2007 (Boeger et al., 2003); in the Brazilian Amazon in 1999 (Couceiro et al., 2006; Piedade et al., 2014), 2018 and 2019; and in 2014 in Bolivian and Peruvian Amazon (Piedade et al., 2014). There is a significant concern for avoiding accidental oil and derivative spills during the exploitation, transport, and storage of petroleum. However, in spite of the proofs that indicate decreases in oil leaking and spills (ITOPF, 2019), petrochemical pollution of the aquatic environment is still significant (Azevedo et al., 2012; Burgherr, 2007; Huijer, 2004; ITOPF, 2019; Martinez and Colacios, 2017). Accidental point-source spills into the environment occur via extraction, transportation, or processing, and these are the main sources of petrochemical contamination (Al-Kindi et al., 2000; Connell et al., 1980; Eisler, 1987; Galvan et al., 2016; Hameed and Al- Azawi, 2016; IPIECA, 2000; Irie et al., 2011; Martinez-Porchas et al., 2011; McKeown and March, 1978; Medeiros et al., 2017). Petrochemicals also enter water bodies as diffuse-source pollution through surface runoff, effluent disposal (from industries and refineries), licensed discharges of petroleum products used as energy sources in urban areas, etc. (Al-Kindi et al., 2000; Connell et al., 1980; Eisler, 1987; Santos et al., 2016). These have supported several types of research aimed at the characterization, fate, and monitoring of the impacts and effects of oil and derivatives for the aquatic biota (Kirby and Law, 2010; Merhi, 2010). It is a consensus in all these studies that petrochemicals pollute the aquatic environments, as shown by decreases in water and sediment quality causing various acute and chronic effects in aquatic organisms. These effects are found from individual up to population, community, and ecosystem levels (Albers, 1995; Amado et al., 2006; Burton et al., 2001; Duarte et al., 2017; Martinez and Colacios, 2017; Sadauskas-Henrique et al., 2017; Silva et al., 2009). When entering aquatic environments, oil components undergo different physical and chemical processes (e.g., evaporation, dissolution, emulsion, photolysis, and biodegradation) (Fig. 7.1) responsible for generating a Oil and derivatives 135

Fig. 7.1 Physical and chemical processes of oil and derivatives when entering aquatic environments. Adapted from Kingston (2002). water-soluble fraction (WSF) of these elements, allowing them to be transported over long distances as well as rendering them more persistent in the environment and readily available to the aquatic organisms (Al-Kindi et al., 2000; Connell et al., 1980; Eisler, 1987; Galvan et al., 2016; Hameed and Al-Azawi, 2016; Irie et al., 2011; Medeiros et al., 2017; Santos et al., 2016). The effects of aquatic contamination by oil and derivatives are normally linked with their chemical characteristics (Albers, 1995). Oil and derivatives are composed of a complex mixture of elements such as monocyclic aromatic hydrocarbons and polycyclic aromatic hydrocarbons (PAHs), phenols, heterocyclic compounds containing nitrogen and sulfur, metals, among others (Al-Kindi et al., 2000; Connell et al., 1980; Duarte et al., 2010; Eisler, 1987; Irie et al., 2011; Kochhann et al., 2015; Medeiros et al., 2017; Nikinmaa, 2014; Rodrigues et al., 2010; Santos et al., 2016; Wake, 2005). Different sources of petrochemicals contain a diverse mixture of sub- components (Neff, 1979). Although they occur in low concentrations, the PAHs are the main substances causing toxicity (Abdel-Shafy and Mansour, 2016; Barron et al., 1999; Connell et al., 1980; Manoli and Samara, 1999). Based on the condensed aromatic rings, PAHs can be classified into light (2–3 rings) and heavy (4–6 rings); the latter are more stable and toxic. PAHs with a higher molecular weight have higher octanol–water partition 136 Helen Sadauskas-Henrique et al. coefficients (Kow) and toxicity. They are highly lipophilic, tending to adsorb to particulate matter or accumulate in sediments (Barron et al., 1999; Hylland, 2006; Paine et al., 1996). The 16 priority PAHs according to the United States Environmental Protection Agency (EPA) are listed in Table 7.1, as well as their carcinogenicity according to the International Agency for Research on Cancer (IARC), structure, and molecular weight (IARC, 2010). Water is considered to be the compartment where pollutants are dissolved and are consequently bioavailable. The sediments affect the fate of

Table 7.1 Common names, IARC classification, structure, and molecular weight of the 16 EPA priority PAHs Molecular weight Common name IARC classificationa Structure (g/mol) Naphthalene 2B 128 Acenaphthylene NC 152 Acenaphthene 3 154 Phenanthrene 3 178

Fluorene 3 202

Pyrene 3 202

Fluoranthene 3 202

Benzo[a]anthracene 2B 228

Chrysene 2B 228 Benzo[b]fluoranthene 2B 252 Benzo[k]fluoranthene 2B 252 Benzo[a]pyrene 1 252

Indene[1,2,3-cd]pyrene 2B 276

Dibenzo[a,h]anthracene 2A 278

Benzo[g,h,i]perilene 3 276

aIARC classification: Group 1: the agent is carcinogenic to humans; 2A: the agent is probably carcinogenic to humans; 2B: the agent is a possible carcinogen to humans; 3: the agent is not classifiable as carcinogenic to humans. Oil and derivatives 137 xenobiotics in aquatic environments, and they contain the residual concentration of xenobiotics in the aquatic ecosystems (Froehner and Martins, 2008). Moreover, the sediment can be considered an active compartment, which not only accumulates these pollutants, but has also a fundamental role in the redistribution and bioavailability of the contaminants, acting as a potential source of diffuse pollution (Sarmiento et al., 2011). This way, the sediment is a compartment that results in the integration of all processes that occur in the aquatic ecosystem, being the “environmental sink” where all contaminants eventually end up. Monitoring of sediment, water, and aquatic biota contaminants must go together for a complete environmental toxicology assessment (Froehner and Martins, 2008; Sarmiento et al., 2011).

7.2 Aquaculture and the problem of oil and derivative contamination Farmed species are important to provide food security and nutrition. Aquaculture continues to grow faster than other major food and animal production sectors. Fish is an important protein source to the human population, providing around 37% of the total animal protein consumed by the human population worldwide (FAO (Food and Agriculture Organization) Preamble and Annexes, 2000). Global fish production in 2016 peaked at about 171 million tons, and aquaculture represents 80 million tons (47%) and captures fisheries 91 million tons (53%) of production, inclusive of marine and inland/freshwater water species (FAO, 2018). Of the most cultivated animal species in aquaculture, finfish are the most cultivated species (54.1 million tons), followed by mollusks (17.1 million tons), and crustaceans (7.9 million tons) (FAO, 2018), where inland aquaculture has big contributions to these figures, especially for finfish (47.516 million tons). According to FAO (2018), from a list of 21 most cultivated finfish species, the carps (30%) are the most cultivated fish species of the world aquaculture. However, other fish species are also farmed, e.g., Oreochromis niloticus (8%); Carassius spp. (6%); Catla catla (6%); Salmo salar (4%); Labeo rohita (3%); and Pangasius spp. (3%) (FAO, 2018). The main aquaculture food fish production is concentrated in Asia (89.4% or 71,546 thousand tons, of the world production, which was 80,031 thousand tons in 2016), where China, India, Indonesia, Vietnam, and Bangladesh are the main producers (FAO, 2018). Fish, mollusks, and crustaceans can be intensively cultivated in different aquatic environments, where different biotic and abiotic factors influence the culture environment. They can alter the behavior and physiology 138 Helen Sadauskas-Henrique et al.

(growth/survival) so they must be managed properly for successful production ( Jobling, 2010; Wedemeyer, 1996). Farmed species may be exposed to a wide range of contaminants capable of causing adverse health and welfare effects, for example, oil and derivatives (Cole et al., 2009; Jobling, 2010; Wedemeyer, 1996). The uptake of oil and derivatives by aquatic organisms is dependent on the environmental bioavailability of the compounds like PAHs (i.e., par- titioning of the compounds between sediment, water, and food), as well as the physiology of the organisms (Meador et al., 1995), the size of the organisms, ingestion and growth rate, membrane permeability, lipid composition, ventilation rate, and osmoregulation (Cherr et al., 2017), as well as water temperature, salinity, pH, the presence of dissolved organic carbon (DOC), exposure time, and seasonal changes (DePalma et al., 2011; Landrum et al., 1984; Lippold et al., 2008; Moeckel et al., 2013; Ramachandran et al., 2006; Sadauskas-Henrique et al., 2016; Shukla et al., 2007). The accumulation of PAHs in aquatic organisms may constitute a risk to human health because of the consumption of contaminated organisms (Conte et al., 2016; Nasher et al., 2016; Ranjbar Jafarabadi et al., 2019). Besides the PAHs in fish being metabolized, they can be found in fish edible parts. For example, giant sea perch (Lates calcarifer) collected from a fish farm on Langkawi Island (Malaysia) presented a mean concentration of PAHs of 47.56–573.66ngg 1 dry weight, where the benzo(a)pyrene (BaP) concentration was 6.862.10ngg 1 and the sum of BaP, benzo(a)anthracene, and chrysene was 43.269.8 (Nasher et al., 2016). These values are higher than the European Union standard values for BaP (5ngg 1) and for the sum of BaP, benzo(a)anthracene, and chrysene concentrations (30ngg 1), which are indicators of PAHs carcinogenesis (European Union Commission Regulation, 2006). Anyway, the ECR (excess cancer risk) was calculated for three scenarios of giant sea perch consumption (exposure frequency of 365, 156, and 52 days year 1 for people eating fish 7, 3, and once a week, respectively). The ECR for carcinogenic, noncarcinogenic, and total PAHs on giant sea perch muscle ranged from 10 8 to 10 5. This range is within the acceptable criteria of the EPA (10 6 to 10 4)(Sidhu, 2003; Xia et al., 2010), which means that consuming fish from the Langkawi fish farm does not pose a health risk to the inhabitants. On the contrary, Jafarabadi et al. (2020) conducted the human health risk assessment for the consumption of three coral-reef fish species (Lethrinus microdon, Lutjanus argentimaculatus, and Scomberomorus guttatus) from the Persian Gulf, Kharg Island (Iran). Oil and derivatives 139

Results for the human health risk assessment concluded that the proba- bility of PAH intake via fish consumption was considerable in this area (incremental lifetime cancer risk (ILCR)>110 5, hazard quotients (HQs)>1, hazard index (HI)6, and the ECR>110 6), and therefore, comprehensive management and long-term monitoring are needed. Similar results were reported by Tongo et al. (2018), for carcinogenic potency of PAHs in fish and shellfish to cause carcinogenic health risk was evaluated using individual carcinogenic potencies for PAHs. Benzo(a)anthracene had the highest carcinogenic potency (ngg 1) in prawn Penaeus monodon (4.47) and fish Mugil cephalus (4.99) while BaP had the highest carcinogenic potency (ngg 1) in crab Uca tangeri (3.8). The results for individual carcinogenic potencies for BaP in fish and shellfish showed values exceeding the guideline screening value of 0.67ngg 1 (wet weight) (USEPA, 2000), for human consumption indicating high potential carcinogenic risk. In addition, the results of the ECR from lifetime exposure to PAHs through fish and shellfish consumption were calculated and compared to the acceptable guideline value of 110 6 set by USEPA. The ECR for benzo(a)anthracene in fish suggests that lifetime exposure to benzo(a)anthracene through fish consumption would result in cancer risk. Due to the relatively sedentary behavior and, mainly to the eating habits, organisms belonging to the phylum Mollusca and subphylum Crustacea are at risk of high absorption and retention rates for PAHs (Denadai et al., 2015; Neff et al., 1976; Peteiro et al., 2006; Yuewen and Adzigbli, 2018). Neff et al. (1976) have shown that bivalves and shrimps quickly accumulate PAHs in their tissues, due to the highly lipophilic characteristics of these contaminants and also due to a poor detoxification enzymatic system (D’Adamo et al., 1997). The rate of accumulation of these pollutants is also highly variable between species. Juvenile shrimp (Litopenaeus vannamei), for example, when exposed to different concentrations of chrysene, quickly accumulated this pollutant. However, after 21 days, the accumulation process reaches a plateau (Ren et al., 2015). Bivalve mollusks, on the contrary, accumulate hydrocarbons slowly, but with time-dependent increases (Neff et al., 1976). D’Adamo et al. (1997) demonstrated that the bivalve Mytilus galloprovincialis bioconcentrates (i.e., specific bioaccumulation process by which the concentration of a chemical in an organism becomes higher than its concentration in the air or water around the organism) and bioaccumulates (i.e., net accumulation of a contaminant in or on an organism from all sources including water, air, and diet) BaP and 7,12-dimethyl benz(a) anthracene in their tissues; however, the bioconcentration and 140 Helen Sadauskas-Henrique et al. bioaccumulation rates were faster and higher in relation to fish (Dicentrarchus labrax). The authors argue that these different responses among mussels and fish are probably caused by the very efficient detoxification enzymatic system located in the liver of the fish. In the same way, Mya arenaria showed a high bioaccumulation rate after exposure to several PAHs, directly (via sediment) and indirectly (via food intake, i.e., phytoplankton), where high concentrations of these pollutants were found in the digestive glands and gonads (Frouin et al., 2007). Baussant et al. (2009) identified high molecular weight PAHs (BaP, benzo(g, h, i) perylene, and dibenzo(a, h)anthracene) in the digestive gland of Mytilus edulis and Chlamys islandica after exposure to crude oil. On the contrary, Denadai et al. (2015) identified low molecular weight PAHs (naphthalene, acenaphthylene, anthracene, fluorine, and phenanthrene) in cockles (Tivela mactroides) captured in the Caraguatatuba Bay, southeastern Brazil. Although the 16 EPA PAHs are highly relevant to environmental monitoring programs, they are no included in the food legislation for many countries (Zelinkova and Wenzl, 2015). However, as presented earlier, it is known that PAHs present in the diet can be harmful to the human population (Tarafdar et al., 2018). IARC classified the 16 priority EPA PAHs from 1 to 3 regarding their carcinogenicity (Table 7.1). The carcinogenic, mutagenic, and bioaccumulative capacities of PAHs have been reported by the World Health Organization (WHO), the Food and Agriculture Orga- nization of the United Nations (FAO), the European Scientific Committee on Food (SCF), IARC, the European Food Safety Authority (EFSA), and the USEPA (Ledesma et al., 2016). Considering that aquaculture is an important activity to provide food security and nutrition worldwide, and the oil and derivatives are ubiquitous in aquatic environments, it is important to study and understand the main effects of these substances to the aquatic organisms and how the water chemistry can influence the bioavailability and toxicity of these compounds. This is important information for sustainable aquaculture and to ensure food safety. The next topics will address how the environmental contamination by oil and derivatives can impair fish, mollusk, and crustacean health and, as a consequence, the productivity of aquaculture and its product quality.

7.3 Effects of oil and derivatives on fish species As mentioned before the rapid growth of aquaculture has the potential to increase the risk of contamination with organic petrochemical pollutants, Oil and derivatives 141 therefore understanding the toxicology of petrochemicals supports sustainable aquaculture and seafood safety. Cultivable aquatic organisms, when exposed to different environmental contaminations, can be considered important bioindicators. Several biological endpoints can be measured to evaluate the impacts of these contaminants at different levels of biological organization, namely the subcellular, cellular, organ, organism, population, and ecological levels. Bioindicators are organisms or biological systems that respond to environmental changes by modifying their normal vital functions and/or chemical composition, thus reflecting responses to the current environmental situation (Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). Bioindicator responses are called biomarkers. These are detected and quantified in molecular, cellular, biochemical, and physiological processes; structures; and functions of tissues, organs, and organ systems and are therefore classified as molecular, genetic, biochemical, physiological, behavioral, or structural biomarkers (Van der Oost et al., 2003). Biomarkers may be used to indicate exposure, effect, and/or susceptibility of the organism to a stressor. They can provide biological information on the exposure of organisms to environmental contaminants even at low concentrations and for a short time. Thus, they serve satisfactorily as indicators of sublethal changes, allowing a better understanding of the effects of pollution on the animal as well as anticipating subsequent effects on populations and communities of impacted ecosystems. Therefore, together with the chemical assessment of water and sediment quality (which identify and quantify the substances present in environmental samples), they are considered indis- pensable for the evaluation of water pollution. In addition, biomarkers are important instruments for understanding the dynamics of the ecosystem and for choosing management and recovery measures for a given ecosystem. Because fish are totally dependent on water quality for their survival, they may be increasingly subjected to contaminated or poor-quality water, in many different habitats (freshwater, estuarine, and marine), at different stages of their life cycle (Evans, 1987; Evans et al., 2005; Yancheva et al., 2016). As fish are well studied in terms of their physiology and there are ample scenarios where they can be exposed to petrochemical pollutants in exposure experiments that quantify their toxic effects (Al-Kindi et al., 2000; Duarte et al., 2010; Gad and Saad, 2008; Galvan et al., 2016; Hameed and Al-Azawi, 2016; Hannah et al., 1982; Irie et al., 2011; Jahanbakhshi et al., 2014; Kochhann et al., 2015; Langangen et al., 2017; Martinez-Porchas et al., 2011; McKeown and March, 1978; Medeiros 142 Helen Sadauskas-Henrique et al. et al., 2017; Rodrigues et al., 2010; Santos et al., 2016; Shirdel et al., 2016; Simonato et al., 2008; Souza-Bastos and Freire, 2011; Theron et al., 2014). As fish interact at various trophic levels of the biological chain, they are a bioindicator species group, strongly reflecting ecosystem health (Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). Our discussion will focus on the physiological, biochemical, molecular, and genetic responses of fish to petrochemical pollutant exposures. 7.3.1 Physiological responses 7.3.1.1 Behavioral and hormonal responses Behavioral changes are the first defense responses of organisms to a stressful situation. Fish typically move away from poor-quality or polluted water locations and in doing so are able to maintain physiological homeostasis and avoid the negative effects of pollutants. If they are not able to move away, then behavioral effects due to the pollutant may occur. These subsequent misbehaviors can render the fish less able to detect toxicants and to move away quickly (de Magalha˜es and da Filho, 2007; Wingfield, 2003). Despite this, there are few studies that have evaluated the behavior of fish species susceptible to oil hydrocarbon exposure. Kochhann et al. (2015) found that tambaqui (Colossoma macropomum) exposed to low concentrations of insoluble fractions of crude oil and mineral oil had a reduction in swimming activity and perception of the conspecific alarm substance, with subsequent loss of olfactory response after prolonged exposure. Similarly, common carp (Cyprinus carpio) exposed to the soluble fraction of diesel oil for 96h manifested behavioral alterations, such as increased opercular movement, disordered movements, swimming near the surface of the water, sudden jumps, loss of balance in swimming (Hameed and Al-Azawi, 2016). Behavioral responses are usually accompanied by endocrine responses such as the release of catecholamines and corticosteroids (Al-Kindi et al., 2000; Barton, 2002; Mommsen et al., 1999; Van der Oost et al., 2003; Wendelaar Bonga, 1997). Although a few studies have monitored the effect of petroleum hydrocarbons on catecholamine levels (Al-Kindi et al., 2000), several studies have evaluated cortisol levels in fish exposed to these contaminants (Al-Kindi et al., 2000; Jahanbakhshi et al., 2014; Martinez-Porchas et al., 2011; Simonato et al., 2008; Souza-Bastos and Freire, 2011; Stephens et al., 1997). Cortisol is the recognized endocrine marker for the evaluation of stress responses in fish (Al-Kindi et al., 2000; Barton, 2002; Mommsen et al., 1999; Van der Oost et al., 2003; Wendelaar Bonga, 1997). Since cortisol and catecholamines are involved with the Oil and derivatives 143 generalized stress response, increases in catecholamine and corticosteroid levels are expected on exposure to oil and derivatives (Al-Kindi et al., 2000; Barton, 2002). Higher levels of cortisol and catecholamines ensure the mobilization of energy substrates and increased cardiorespiratory functions as required in the face of stress (Al-Kindi et al., 2000; Barton, 2002). The release of these hormones is influenced by fish species (which have different sensitivities and capacity to respond), contaminant type (e.g., crude oil vs petrochemical derivatives), the life stages of fish, and other environmental factors (Al-Kindi et al., 2000; Barton, 2002). Corroborating this information, adults of sole (Pleuronectes flesus) present elevated noradrenaline levels (Al-Kindi et al., 2000), and larvae and adults of this species exposed to the WSF of crude oil presented high concentrations of cortisol in their whole body (Stephens et al., 1997) and plasma (Al-Kindi et al., 2000), respectively. The Brazilian silverside (Atherinella brasiliensis) showed a progressive increase in plasma cortisol concentration over 7 months after an oil spill (Souza-Bastos and Freire, 2011). In contrast, adult specimens of Nile tilapia injected with BaP showed a decrease in cortisol levels (Martinez- Porchas et al., 2011), as occurred with curimba (Prochilodus lineatus) after exposure of 15 days to diesel WSF (Simonato et al., 2008) and with common carp after exposure to crude oil ( Jahanbakhshi et al., 2014). Generally, exposures to petrochemicals produce a cortisol and catecholamine response. The response can be measured in blood or tissues, which may be acute or chronic. However, the reductions in cortisol can also occur because these contaminants can act as endocrine disruptors and impair these stress responses (Al-Kindi et al., 1996; Simonato et al., 2008). Oil and derivatives are also capable of altering the secretion of other hormones, such as the thyroid hormones [thyroxin (T4) and triiodothyronine (T3)] (Al-Kindi et al., 2000; Barton, 2002; Brown et al., 2004; Gad and Saad, 2008; Santos et al., 2016; Shirdel et al., 2016; Stephens et al., 1997). Adult Nile tilapia, for example, showed a reduction in thyroid hormone levels after 16 weeks of exposure to three different concentrations of phenol, a pollutant commonly found in refinery effluents. Such reduction was related to metabolic disturbances (e.g., total serum cholesterol and lipid content increased, micronuclei production increased) and consequent weight loss in this species, directly proportional to both concentrations and time of exposure (Gad and Saad, 2008). Common carp also showed a reduction in plasma concentrations of these hormones when exposed to 100μgL 1 of pyrene (Shirdel et al., 2016). Similarly, juvenile trahira (Hoplias malabaricus) presented with growth reduction (in length, but not 144 Helen Sadauskas-Henrique et al. in weight) after 28-day exposure to 50% of the WSF of petroleum (Santos et al., 2016). However, sole larvae presented with an increase in T4 concentration when exposed to 25%, 33%, and 50% of crude oil WSF (Stephens et al., 1997). On the contrary, the adults of sole presented with a reduction in plasma concentrations of T4 without any change in T3 levels (Al-Kindi et al., 2000). Although variable, the responses expressed by different species of fish demonstrate that oil and derivatives interfere in the endocrine responses of these organisms leading to disturbances in homeostasis, development, fertilization, and reproduction (Al-Kindi et al., 1996; Barton, 2002; Brown et al., 2004; Goksoyr and Forlin, 1992; Irie et al., 2011; Langangen et al., 2017; Rodrigues et al., 2010; Shirdel et al., 2016; Wake, 2005; Wendelaar Bonga, 1997). The examples of homeostasis disturbances include changes in lipid metabolism, glycemic responses (usually hyperglycemia), protein synthesis, or induced catabolism (Al-Kindi et al., 2000; Duarte et al., 2010; McKeown and March, 1978; Shirdel et al., 2016; Simonato et al., 2008). A study by Al-Kindi et al. (1996) showed that the exposure of sole to crude oil WSF produced severely elevated plasma glucose. This response is also seen in common carp ( Jahanbakhshi et al., 2014). Likewise, the same species exposed to different concentrations of pyrene also presented high levels of plasma glucose, but with a reduction in plasma concentrations of total protein and albumin (Shirdel et al., 2016). On the contrary, curimba exposed to soluble fractions of diesel oil presented with the depletion of protein levels, although without any alteration in plasma glucose levels (Simonato et al., 2008).

7.3.1.2 Gills responses The effects of the exposure to the dispersed particles or to the soluble fraction of oil and derivatives cause fish to exhibit different responses ranging from sublethal to lethal depending on the form of exposure, intensity and exposure time, and rate of absorption (Al-Kindi et al., 1996, 2000; Barton, 2002; Hameed and Al-Azawi, 2016; Jobling, 2010; Nikinmaa, 2014; Yancheva et al., 2016). Although the contamination can occur both through the respiratory surface and through the ingestion of contaminated water and food, the main and most direct contact with these pollutants occurs through the gills (Al-Kindi et al., 1996, 2000; Connell et al., 1980; Duarte et al., 2010; Evans et al., 2005; Medeiros et al., 2017; Nikinmaa, 2014; Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). Gills form the main contact surface between the fish Oil and derivatives 145 and the water, respond directly and rapidly to any changes in the surrounding water, and thus may be both morphologically and physiologically affected (Evans et al., 2005; Medeiros et al., 2017; Nikinmaa, 2014; Simonato et al., 2008; Theron et al., 2014; Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). Morphological changes in the gills are usually the primary, and common, lesions generated by waterborne pollutants (Al-Kindi et al., 1996, 2000; Evans et al., 2005; Medeiros et al., 2017; Rodrigues et al., 2010; Theron et al., 2014; Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). Morphophysiological changes caused by oil and derivatives in the gill epithelium include synthesis or destruction of molecular components of the transport system; variations in the morphology; and number of chloride and mucous cells, lesions (hemorrhage, hyperplasia, lamellar fusion, hypertrophy, aneurysms, degeneration/separation of epithelium, necrosis, edema, congestion, lamellar undulation, etc.); vasodilation or vasoconstriction; changes in ATPase activities; increase in the expression of stress proteins; increase in mucus production, and among others (Fig. 7.2). These changes reduce the surface area and increase the gas-exchange diffu- sion distance of the gills with the water altering the blood flow, damaging circulatory and respiratory processes, which if not resolved can trigger the death of the fish (Al-Kindi et al., 1996, 2000; Connell et al., 1981; Duarte et al., 2010; Hameed and Al-Azawi, 2016; Kochhann et al., 2015; McKeown and March, 1978; Medeiros et al., 2017; Rodrigues et al., 2010; Simonato et al., 2008; Theron et al., 2014; Yancheva et al., 2016). Although gill lesions have not been evaluated, the reduction in oxygen consumption and availability was verified in sole exposed to WSF of crude oil (Al-Kindi et al., 1996), and in tambaqui exposed to an insoluble fraction of crude oil and mineral oil (Kochhann et al., 2015). Nevertheless, a high respiratory rate was verified in tamoata´ (Hoplosternum littorale) exposed to the WSF of crude oil (Brauner et al., 1999). Such responses may be related to branchial changes or constitute a generalized stress response, both generated by the presence of these contaminants (Al-Kindi et al., 1996, 2000; Connell et al., 1981; Duarte et al., 2010; Hameed and Al-Azawi, 2016; Kochhann et al., 2015; McKeown and March, 1978; Medeiros et al., 2017; Rodrigues et al., 2010; Simonato et al., 2008; Theron et al., 2014; Yancheva et al., 2016). Gillsarehighlyvascularstructures that have intimate contact with blood and are related to the maintenance of vital processes (osmoregulation, acid–base balance, respiration, and excretion of nitrogenous 146 Helen Sadauskas-Henrique et al.

Fig. 7.2 Representation of common histopathologies in fish gills (Astyanax fasciatus). (A) Normal gill structure; (B) lamellar epithelial hypertrophy (arrow); (C) epithelial hyperplasia and total fusion of several lamellae (arrow); (D) lamellar congestion (★); (E) lamellar telangiectasia (★); (F) lamellar epithelial lifting (arrow); (G) lamellar cell hyperplasia and hypertrophy (arrow); (H) lamellar mucous cell hyperplasia and hypertrophy (arrow); (I) edema (★). L. lamella; F. filament; C. cartilage. Glutaraldehyde, toluidine blue. Scal- e¼20μm. Credit: Sadauskas-Henrique, 2008 residues). Once the structural properties of the gills have changed, the occurrence of functional alterations follows (Albers, 1995; Al-Kindi et al., 2000; Duarte et al., 2010; Evans, 1987; Evans et al., 2005; Hameed and Al-Azawi, 2016; Kochhann et al., 2015; Medeiros et al., 2017; Mommsen et al., 1999; Simonato et al., 2008; Theron et al., 2014; Van der Oost et al., 2003; Wendelaar Bonga, 1997; Yancheva et al., 2016). The lipophilic nature of petroleum hydrocarbons enables them to readily cross cell membranes (Albers, 1995; Al-Kindi et al., 2000; Connell et al., 1981; Duarte et al., 2010; Hameed and Al-Azawi, Oil and derivatives 147

2016; Medeiros et al., 2017; Rodrigues et al., 2010) and be readily distributed systemically leading to a range of biological responses including stress (Al-Kindi et al., 2000; Barton, 2002; Connell et al., 1980; Evans, 1987; Evans et al., 2005; Rodrigues et al., 2010; Theron et al., 2014; Wendelaar Bonga, 1997; Yancheva et al., 2016). Osmoregulation maintains the osmo-ionic concentration of the extracellular fluid, or internal environment of the animal as part of homeostasis and normal cellular functioning (Al-Kindi et al., 2000; Jobling, 1995). Hydromineralimbalanceisaneffectcommonlycausedbyoilsandby- products on different species of marine or freshwater fish (Al-Kindi et al., 1996, 2000; Brauneretal.,1999; Duarte et al., 2010; McKeown and March, 1978; Medeiros et al., 2017; Shirdel et al., 2016; Simonato et al., 2008; Souza-Bastos and Freire, 2011; Theron et al., 2014). Tambaqui exposed to the mixture of crude oil and surfactant presented changes in whole-body Na+ and Cl fluxes (i.e., diffusive loss) and reduction in plasma levels of these ions. Although K+ flux accompanied the major ions in fish exposed to the highest concentrations of crude oil and surfactant, the concentration of this ion increased in tambaqui plasma, possibly indicating a rupture of cells. In addition, tambaqui had a reduction in plasma Ca2+ concentration, but did not show any changes in plasma Mg2+ concentration (Duarte et al., 2010). McKeown and March (1978) also identified diffusive loss and consequently a reduction in plasma Na+ levels in rainbow trout (Oncorhynchus mykiss) exposed to bunker oil C and dispersant (oil disperse 43) alone or as a mixture in freshwater. In contrast, Brauner et al. (1999) found that tamoata´ exposed by immersion to the WSF of crude oil did not show any change in plasma Na+ and K+ levels, but when contaminated through the diet showed a reduction in the plasma concentrations of these ions. In the same way, common carp also did not present alterations in plasma concentrations of Na+,K+,andCa2 + when exposed to different concentrations of pyrene (Shirdel et al., 2016). Sole exposed to the WSF of crude oil showed no alteration in osmolality and plasma concentrations of Na+ and Cl but showed an increase in plasma K+ concentrations (Al-Kindi et al., 1996). Turbot (Scophthalmus maximus) showed increased osmolality and plasma concentrations of Na+ and Cl after exposure to the WSF of crude oil (Theron et al., 2014). Brazilian silverside collected 1 month after the occurrence of an oil spill showed a reduction in osmolality and Cl plasma concentrations. However, these parameters increased progressively up to 7 months after the oil spill (Souza-Bastos and Freire, 2011). 148 Helen Sadauskas-Henrique et al.

7.3.1.3 Hematological responses Accompanying osmoregulatory changes, the effects on hematological parameters are also frequently used to identify physiological changes triggered by oil and derivatives, which have a strong hemolytic effect (Al-Kindi et al., 1996, 2000; Duarte et al., 2010; Simonato et al., 2008). However, hematological changes in fish exposed to these pollutants may occur both as a consequence of changes in the different processes already discussed so far (Al-Kindi et al., 2000), or as manifestations of immunological reactions (Al-Kindi et al., 2000; Wendelaar Bonga, 1997). Immune responses are generally activated in organisms exposed to low concentrations and suppressed in the presence of high concentrations of these pollutants (Nikinmaa, 2014) and both situations may affect hematological parameters (Al-Kindi et al., 2000; Jahanbakhshi et al., 2014). The study with tamoata´ exposed to WSF of crude oil, for example, indicated an increase in hematocrit and hemoglobin levels (Brauner et al., 1999). Conversely, the specimens of curimba exposed to the WSF of diesel, both for 96h and 15 days, had a high incidence of hemolysis and marked reduction in hematocrit and hemoglobin levels (Simonato et al., 2008). Likewise, the sole exposed to the WSF of crude oil also showed a reduction in hematocrit concentration and hemoglobin (Al-Kindi et al., 1996). Tambaqui, on the contrary, presented an increase in hematocrit levels and a reduction in hemoglobin concentration when exposed to crude oil and surfactant (Duarte et al., 2010). Common carp fishes exposed for 96h to crude oil showed an increase in both hematological and immunological responses, including hematocrit, red blood cells, hemoglobin, and neutrophil increase, but a reduction in peripheral blood lymphocyte concentration ( Jahanbakhshi et al., 2014). Turbot, on the contrary, showed a reduction in leukocyte and granulocyte concentrations after exposure to the WSF of crude oil (Theron et al., 2014). The study carried out with lambari (Astyanax altiparanae) suggests that petroleum hydrocarbons may reduce erythrocyte production but cause an increase in the frequency of thrombocytes, neutrophils, and lymphocytes, at the beginning of the recovery/ depuration period (Galvan et al., 2016). Although in some cases there is a reduction in the effects of these pollutants on fish after a certain period of recovery, the total recovery of organisms does not occur (Galvan et al., 2016; Medeiros et al., 2017). Petrochemical pollutants by altering the physiology of fish can induce negative effects on the growth, development, reproduction, and survival of fish whether these are cultured or wild ranging animals (Al-Kindi et al., 2000; Brown et al., 2004; Gad and Saad, 2008; Hannah et al., 1982; Oil and derivatives 149

Martinez-Porchas et al., 2011; Rodrigues et al., 2010; Santos et al., 2016; Shirdel et al., 2016; Wake, 2005). Nevertheless, studies that relate to the effects of these pollutants on the production process are lacking. In addition, studies that focus on Neotropical fish species (e.g., fish species in Central America, including the southern part of Mexico and the Baja California peninsula, southern Florida, all Caribbean islands, and South America) and their responses to petrochemical toxicity are still scarce. Such studies should focus on aquacultured species and consider the toxic effects at different levels of biological organization. In this sense, it is necessary to perform more studies that assess the effects of petroleum hydrocarbons on osmoregulatory processes and on acid–base imbalance. No studies have been found that inte- grate evaluating morphofunctional damage in the gills, endocrine responses, and their implications on osmoregulation. A single study correlated the effects of petroleum hydrocarbons on the response of the carbonic anhydrase enzyme (Souza-Bastos and Freire, 2011), which is of great importance for vital processes, such as respiratory processes and acid–base balance, which are known to be altered by oil and derivatives (Al-Kindi et al., 1996; Brauner et al., 1999; Kochhann et al., 2015; Theron et al., 2014).

7.3.2 Biochemical responses Biochemical biomarkers can be considered important tools to monitor environmental quality because they can reflect the health of the organisms in the environment (Walker and Henderson, 1996). In this sense, they are widely used to evaluate the water quality and the effects of contaminants for the aquatic biota (Van der Oost et al., 2003). The PAHs are the most toxic portion of the oil and derivatives. These low molecular weight petroleum hydrocarbons (4–6 carbons) can be easily absorbed by gill and epidermal tissue and the digestive tract of fishes causing alterations on fish physiology (Sadauskas-Henrique et al., 2017). Several biochemical biomarkers have been used to evaluate the effects of oil and derivatives on fish health. Among these biomarkers, the alterations on the activity of the biotransformation and antioxidant enzymes together with genotoxic damage are commonly identified by biomarkers since these compounds are known for their carcinogenic and mutagenic effects (Martı´nez-Go´mez et al., 2006; Sadauskas- Henrique et al., 2017). However, these effects only become apparent after their metabolization, in a process called bioactivation. For example, epoxide, diol epoxide, and hydrometabolites are primary and secondary metabolites generated after BaP oxidation by cytochrome P450, and 150 Helen Sadauskas-Henrique et al. monooxygenases are formed in Phase I of the biotransformation processes (Miller and Ramos, 2001; Shimada, 2006; Shimada and Fujii- Kuriyama, 2004). Phase I of the biotransformation process consists of oxidation, reduction, and hydrolysis reactions catalyzed by enzymes from the mixed-function oxidase systems (monooxygenases cytochrome P450; cytochrome b5 and NADPH-cytochrome P450). This system biotransforms lipophilic compounds into more hydrophilic metabolites. One way of measuring the cat- alytic activity of the cytochrome P450 is via ethoxyresorufin-O-deethylase (EROD) activity. Several authors have reported the induction of hepatic EROD activity in fish acutely exposed to different types of contaminants, such as the WSF of diesel (Simonato et al., 2011), crude oil (Ramachandran et al., 2004; Sadauskas-Henrique et al., 2016), and chemically dispersed crude oil (Ramachandran et al., 2006; Sadauskas-Henrique et al., 2016). Several field studies have reported increased activity of hepatic EROD in fish collected from recently contaminated areas ( Jung et al., 2012; Martı´nez-Go´mez et al., 2006) or that have a history of contamination by oil and oil products (Buet et al., 2006; Devier et al., 2013; Duarte et al., 2010; Gagnon and Holdway, 2002; Trisciani et al., 2011). Phase II of biotransformation occurs via conjugation reactions of Phase I primary and secondary metabolites with polar molecules, such as glutathione, facilitating excretion and elimination of the final product (Timbrell, 2015). Glutathione S- transferase (GST) is an important enzyme in Phase II of biotransformation, participating in conjugation reactions of the xenobiotic with the exogenous tripeptide reduced glutathione (GSH) and resulting in the efficient elimination of the metabolites from the organism (Rinaldi et al., 2002). GST is a widely used biomarker for a variety of fish species exposed to oil and derivatives in both laboratory (Kochhann et al., 2013; Sadauskas-Henrique et al., 2016, 2017; Simonato et al., 2011) and field studies (Duarte et al., 2017; Tim-Tim et al., 2009). The continuous NADPH consumption by the cytochrome P450 during the metabolism of oil and derivatives, such as PAHs and the BTEX (benzene, toluene, ethyl-benzene, and xylene) inevitably produce reactive oxygen species (ROS) (Wiernsperger, 2003). The organisms possess an antioxidant system, to cope with the excess of ROS, such as peroxyl and hydroxyl radicals. However, when ROS generation exceeds the antioxidant capacity of the organism, oxidative stress and oxidative damage may occur in the cells, tissues, and organs. The antioxidant system consists of enzymes that neutralize the ROS generated, such as superoxide dismutase (Cu-Zn SOD), Oil and derivatives 151 catalase (CAT), and glutathione peroxidase (Se-GPx), as well as by nonenzymatic antioxidants such as glutathione tripeptide (GSH). The SOD is the first enzyme in the line of antioxidant defense and acts by neutralizing the superoxide radicals, generating hydrogen peroxide that, in turn, is neutralized by CAT. The GPx catalyzes the metabolism of oxidizing compounds, a process that involves the ratio between the reduced glutathione (GSH) oxidized glutathione (GSSG) (i.e., GSH:GSSG ratio) (Van der Oost et al., 2003). Low level of the GSH:GSSG indicates oxidative stress, once is considered to be one of the most important scavengers of ROS (Van der Oost et al., 2003). The GPx is also involved in reducing lipid hydroxyperoxides to their corresponding alcohols, making them more water soluble as well as reducing free hydrogen peroxide to water (Van der Oost et al., 2003). When the antioxidant system fails in neutralizing the excess of ROS generation, oxidative damage may occur in the cells, tissues, and organs. The ROS can interact with the biomembranes and disturb the membranes’ delicate structure, integrity, fluidity, and permeability, resulting in a loss of functionality as a consequence of lipid peroxidation (LPO) (Niki, 2009; Wiernsperger, 2003). Damage to the lipid bilayer (i.e., lipoperoxidation) is considered a significant cause of cell injury and death following exposure to contaminants (Modesto and Martinez, 2010).TheROScanalsodam- age proteins and DNA (DNA strand breaks) and are also commonly used as biomarkers of oil and derivative contaminations (Van der Oost et al., 2003).

7.3.3 Molecular and genetic responses The combination of advances in sequencing technology and the development of microarray technology has made measurements of global gene expression in ecologically relevant species possible (Hook, 2010). Genomic approaches, using fish, promise to increase research capability by providing insights into the mechanisms that underlie short-term and long-term environmental adaptations to contaminants. Fish offer many advantages for investigating the organism–environment interface (Cossins and Crawford, 2005). Environmental genomic research using transcriptome and gene expression is a fundamental tool for understanding toxicological susceptibility and responses to toxins, local adaptation of tolerant populations, and identification of refined ecotoxicological biomarkers for environmental regulation at the gene level (Cossins and Crawford, 2005). 152 Helen Sadauskas-Henrique et al.

Aardema and MacGregor (2002) define toxicogenomics as “the study of the relationship between the structure and activity of the genome (the cellular complement of genes) and the adverse biological effects of exogenous agents.” Since gene expression responses represent the primary interaction between environmental contaminants and biota, they provide essential clues to understanding how chemical exposure can affect organismal health (Moens et al., 2007). Currently, there are several studies that have evaluated the molecular and genetic effects of petroleum hydrocarbons and their derivatives in aquatic organisms, and most of these used crustacean, bivalve, and fish species (Bowen et al., 2018; Madison et al., 2017; Zhuang et al., 2017).

7.3.3.1 Fish genes affected by oil and derivatives 7.3.3.1.1 Ras oncogene and Hif-1a gene Rat sarcoma virus (ras) oncogene and hypoxia-inducible factor (hif-1α) have drawn attention in the literature due to their relationship with cancer development (Melstrom et al., 2011; Pratilas and Solit, 2010). Ras consists of a gene family that was first identified in viruses that caused sarcomas in the rat (Bos, 1989; Karnoub and Weinberg, 2008; Lowy and Willumsen, 1993). When normally expressed, ras genes act to regulate various physiological functions, including control of proliferation, differentiation, and cell death. The ras genes are present in all eukaryotes and have been characterized in several species of fish. Fish ras genes have nucleotide sequences similar to mammalian homologs, demonstrating a high degree of conservation (for review, see Rotchell et al., 2001). The first sequence for the ras oncogene to be characterized in fish was from the goldfish (Carassius auratus)(Nemoto et al., 1986); then, other fish species had the sequences for the ras genes, e.g., rainbow trout (Mangold et al., 1991), rivulus (Rivulus marmoratus) (Lee et al., 1998), and zebrafish (Danio rerio)(Cheng et al., 1997). The hif-1a produces the protein HIF-1, which is the major regulator of oxygen-dependent gene expression (Maxwell et al., 1997; Rytkonen€ et al., 2007). The levels of hif-1α expression are associated with tumor genesis and angiogenesis. One of the target genes influenced by hif-1a is the vascular endothelial growth factor (VEGF) that plays important roles in tumor neovascularization, which supplies sufficient amounts of oxygen and nutrients to tumor cells to allow tumor propagation and metastasis (Goda et al., 2003; Zhong et al., 1999). Although hif-1α has been mostly associated with hypoxic responses in fish, tumor cell hypoxia is also a well-studied system (Geng et al., 2014). Oil and derivatives 153

Recently, ras oncogene and hif-1a were identified in tambaqui (Silva et al., 2017), a species that had been largely used as a model in studies with petroleum derivatives (Duarte et al., 2010; Kochhann et al., 2015). BaP injections (4, 8, 16, and 32μmolkg 1) resulted in alteration in the expression of ras and hif-1a genes in the liver. An increase was observed in the expression of ras oncogene in tambaqui exposed for 4, 8, and 16μmol of BaPkg 1 compared to control fish. However, at the highest concentrations of BaP (16 and 32μmolkg 1), the expression of hif-1α was similar to the control group. The highest expression of hif-1a occurred at the lowest concentration of BaP, suggesting that the hepatocytes were capable of activating the transcription of this gene, helping to maintain the cell survival machinery. In fish exposed to the highest concentration of BaP, the cellular machinery was already compromised by cell damage, and the tissue was not efficient in controlling gene expression to keep levels of hif-1a high, since the normal functioning of the liver was impaired by necrosis (Silva et al., 2017). Anguilla anguilla was also exposed to BaP (0.1 and 0.3μΜ BaP over 4 weeks) in laboratory conditions (Nogueira et al., 2006). The BaP-exposed A. anguilla liver samples were screened for ras mutations by direct sequencing of ras cDNA. No point mutations were found at the traditional mutation hot spots, codons 12, 13, and 61, in control or BaP-exposed samples. The analysis of ras gene expression levels in the same samples revealed no difference between control and exposed fish. The lack of ras gene mutations and no apparent change in ras gene expression levels following BaP exposure suggests that hydrocarbon genotoxic endpoints in A. anguilla involve a molecular etiology that does not involve the ras gene. The apparent lack of point mutations in the eel ras gene does not rule out other potential detrimental effects. The ras gene, one of the most important genes involved in carcinogenesis, is mutated in fish from areas of high PAH contamination (Rotchell et al., 2001).

7.3.3.1.2 p53 gene The tumor suppressor p53 (tp53 gene) is a well-known tumor suppressor, which is also involved in organismal aging and developmental control processes. As a transcription factor, the p53 protein regulates the expression of target genes encoding proteins involved in cell growth control (Brady and Attardi, 2010). The p53 gene expression and protein abundance have been studied in other aquatic organisms including rainbow trout (Liu et al., 2011), Japanese medaka (Oryzias latipes)(Min et al., 2003), and a hermaphroditic fish (Kryptolebias marmoratus)(Lee et al., 2008). A study developed by de 154 Helen Sadauskas-Henrique et al.

Souza et al. (2019) identified p53 for tambaqui. The authors analyzed tumor suppressor p53 gene expression after naphthalene injection (50mgkg 1) and subsequent exposure to acute hypoxia. In fish injected with naphthalene, there was an increase in p53 mRNA levels in both normoxia and hypoxia. According to the authors, the aquatic contamination in this species increased, preferentially, the p53 gene transcription, as an attempt to ensure the maintenance of genomic integrity, but this was not as efficient as in fish kept in normoxia, demonstrating that responses at the transcriptional level of the p53 gene in fish may be compromised by hypoxia. Ruiz et al. (2012) also observed an increase in p53 transcripts level in turbot (Scophthalmus maximus) after 14-day exposure to heavy fuel oil and, then, transcription level returned to control levels after 14 days of the recovery period. The authors suggested that turbot responded to hydrocarbon exposure by triggering P53-mediated cell-cycle arrest. On the contrary, Patinõ-Sua´rez et al. (2016) demonstrated that BaP injection (20mgkg 1) in Nile tilapia resulted in a decrease in p53 gene after 24h of exposure, corroborating the toxic effects of BaP on the aquatic biota. The reduction in p53 gene expression may be related to the occurrence of DNA damage and mutations resulting from BaP exposure, since p53 is a uni- versal sensor of genotoxicity that regulates the transcription of genes involved in cell-cycle arrest, DNA repair, and cellular senescence. BaP is a potential inductor of p53 mutation as demonstrated by Sueiro et al. (2000) where BaP-induced adenine transversions in the sole (Platichthys flesus). The authors explained that the adenine mutation is due in part to differing metabolic activation or DNA repair systems for the sole (Sueiro et al., 2000). Williams and Hubberstey (2014) analyzed the expression of p53 in brown bullhead (Ameiurus nebulosus) in the laboratory. They found that BaP-fed fish collected from the contaminated area had a reduction in gene expression, which was observed in both transcript and protein levels. The authors also analyzed the p53 expression in brown bullhead captured at clean sites, whether or not fed with BaP. The results showed that fish collected in a clean site (Pêche Island), when fed BaP-treated food immediately after capture, showed an increase of the p53 protein compared to control. However, fish collected at a contaminated site (Hamilton Harbor), when fed BaP- treated food immediately after capture, showed a reduction in protein level. According to the authors, the inhibitory effects of BaP feeding on fish caught in contaminated environments may be due to the huge increase in BaP exposure in these fish upon feeding. The fish caught from contaminated Oil and derivatives 155 environments were exposed to a multitude of different environmental chemicals and over prolonged exposure; p53 levels have been adapted to protect these fish. Fish can adapt to the pressures of toxic stress in order to survive, acquiring resistance by effectively selected phenotypes that bring more advantage. Yuan et al. (2017) showed that the exposure of Gobiocypris rarus to BaP resulted in increased p53 gene expression. Overall, increased p53 expression was accompanied by increased transcription of p21,ap53 target gene involved in cell-cycle regulation; increased gene involved in DNA damage repair; and an increase in bax gene transcripts, apoptosis-inducing gene by the p53-dependent pathway. The authors suggested that to deal with the stress caused by exposure to BaP, G. rarus invests in cell-cycle block and DNA repair by activating p53/p21/gadd45a, as well as inducing apoptosis through activation of p53/bax. In general, the p53 response is complex and can vary for different organisms exposed to the same contaminant (Patinõ-Sua´rez et al., 2016; Yuan et al., 2017). The changes in gene expression can vary over time, with an increase being reported after 96h (de Souza et al., 2019), 14 days (Ruiz et al., 2012), or even just a week after depuration (Martins et al., 2018). However, the role of p53 in restricting neoplastic progression in fish has already been demonstrated, showing that the functional loss of p53 by mutations induces tumorigenesis (Berghmans et al., 2005; Tu et al., 2016). The mutations within the genome, particularly those occurring in crucial cell regulatory genes such as p53 gene, could have potent consequences on the life and health span of the organisms affected. The activation of mutated p53 may increase the risk of cancer development in fish exposed to petrochemical pollutants because it alters the control of the cell replication cycle in organs (Nadler, 2017). In addition, by acting on the maintenance of genomic integrity, the increase in the expression of p53 in fish exposed to oil derivatives can activate cell signaling pathways related to cell-cycle arrest and programmed cell death, preventing the spread of possible genetic damage and minimizing the risk of carcinogenesis; the same goes for inhibiting gene expression, which can predispose fish to the risk of carcinogenesis (Yuan et al., 2017). Increased expression of p53 in fish can be a differentially modulated response to deal with stressful situations (Williams and Hubberstey, 2014).

7.3.3.2 CYP1A Cytochrome P450 1A (CYP1A) protein converts lipophilic xenobiotics to more water-soluble compounds, promoting excretion and detoxification 156 Helen Sadauskas-Henrique et al.

(Van der Oost et al., 2003). The mechanism of induction of CYP1A involves the compounds that have a high affinity of the enzyme as an aryl hydrocarbon receptor (AhR) present in the cytoplasm of the cell. Aryl hydrocarbon receptor (AhR) is a ligand-activated transcription factor that controls the expression of CYP1A. The members of the cytochrome P450 1A (CYP1A) subfamily are inducible by PAH. The induction mechanism involves the high-affinity binding of inducer to the aryl hydrocarbon receptor (AhR) and subsequent activation of the AhR-dependent signal transduction pathway. The ligand-activated AhR combines with the AhR nuclear translocator (ARNT) to form an active transcription factor that controls the inducible expression of CYP1A1 and other genes. In the inac- tive form, AhR is coupled to proteins hsp90 (heat-shock protein). In the presence of PAHs, the combination between hsp90 and AhR is broken down and the receptor is associated with PAHs, which is then moved to the cell nucleus, where it induces the expression of the cyp1a gene (Goksoyr and Forlin, 1992; Hahn et al., 1998). Studies have shown that the expression of cyp1a in fish is directly related to the levels of PAHs; therefore, the cyp1a gene has been especially highlighted as a biomarker of exposure to xenobiotics (Anjos et al., 2011; Billiard et al., 2002; Hahn et al., 1998; Santos et al., 2018; Simonato et al., 2011). In fish, the CYP1 family consists of four subfamilies, CYP1A, CYP1B, CYP1C, and CYP1D. The cyp1a gene products catalyze the oxidation of environmental carcinogens and are therefore critical determinants in pathways leading either to detoxification and excretion or to DNA adduct formation and carcinogenesis (Gelboin, 1980). Uno et al. (2012) pinpointed 8 CYP families, namely CYP1, CYP2, CYP3, CYP4, CYP11, CYP17, CYP19, and CYP26 in their review, and across all species of fish, 137 genes encoding P450s have been identified. In the Amazon region, another species that has been used as a potential model for monitoring of oil contaminants is the Amazonian cichlid Astronotus ocellatus (oscar). Oscar has economic importance as ornamental and food fish (Val and Almeida-Val, 1995). It is very tolerant of hypoxia (Baptista, 2016). Anjos et al. (2011) verified that the cyp1a expression in Oscar liver exposed to sublethal concentrations of the WSF prepared from a mixture of crude oil, under varying environmental conditions, showed high sensitivity as an indicator of exposure to contaminants. Transcript expression of cyp1a showed a clear concentration–response relationship. A reduced cyp1a induction at hypoxic conditions was identified for the highest WSF concentrations only. Hence, this weak cyp1a induction may indicate Oil and derivatives 157 the toxicity of high concentrations of WSF ingredients at hypoxic conditions leading to interference with the transcriptional response for Oscar (Anjos et al., 2011). The exposure of tambaqui to crude oil induced higher levels of CYP1A protein, especially when combined with humic substances (Matsuo et al., 2006). The curimba has already been shown to be sensitive to various classes of contaminants, including petroleum derivatives that contain large quantities of PAHs (Simonato et al., 2011). Santos et al. (2018) verified cyp1a expression in curimba injected with BaP (20mgkg 1). The relative expression of cyp1a in curimba liver and gill tissues was significantly higher in the BaP group compared to the oil group at 6 and 24h after BaP injection (Santos et al., 2018). The BaP was capable to induce overexpression of cyp1a in curimba. Petroleum derivatives may induce different responses as demonstrated by Woz´ny et al. (2010). Rainbow trout exposed to dibenzothiophene (DBT) (intraperitoneally injected doses of 10 or 50mgkg 1), a common component of crude oil, resulted in lower cyp1a mRNA levels in liver and gills at the end of the experiment (24h), and also a final decrease of CYP1A protein levels. The authors concluded that DBT reduces CYP1A activity in fish (Woz´ny et al., 2010). In general, the authors found that rainbow trout exposure to various xenobiotics resulted in increased expression of cyp1a gene. The increases in gene expression occurred at 8 and 24h after exposure for treatment with BaP, while the exposure with the other PAHs increased cyp1a mRNA only after 24h. Williams and Hubberstey (2014) tried to establish whether there are adaptive genetic/molecular changes occurring in brown bullhead that allow their survival in a polluted area. Fish collected from nonpolluted areas fed BaP-treated food (10mL of the BaP mixture (0.087M)) showed liver cyp1a expression markedly increased at a greater rate than fish caught from contaminated sites. Fish fed BaP-treated food after being cleared for 1 week showed that cyp1a expression increased at a higher rate in clean fish vs contaminated fish from a basal state. Brown bullhead fish from contaminated areas maintain a low cyp1a expression to combat the effects of environmental stress, probably because chronic cyp1a activation is a metabolic cost to the fish. The transcriptional response also reveals the effects of PAHs on the fish immune system as shown by Phalen et al. (2017) in their studies with rainbow trout. Rainbow trout were exposed in vivo with a single intraperitoneal injection of corn oil or 100mgkg 1 of the immunotoxic BaP in corn oil. 158 Helen Sadauskas-Henrique et al.

The cyp1a1 and cyp1a3 were highly expressed in liver, B cells, and thrombocytes in response to BaP. The cyp1a3 expression was higher than cyp1a1 in liver and B cells, whereas the inverse was true for thrombocytes. The cyp1c1 and cyp1c2 had the highest relative expression in B cells and could not be detected at all in thrombocytes. The cyp1b1 had a low relative expression in all tissues. There was no significant difference in the pattern of CYP expression between similar cell types between tissues. This study provides further support for the observation that CYPs in the CYP1 family are differentially expressed in different tissue types. Differential expression of CYP isoforms in the same cell types from different tissues may be important for metabolite-induced cellular toxicity.

7.3.3.3 Transcriptome fish response and perspective for fish population adaptation Research has recently focused on predictive indicators of adverse health effects in aquatic species. A strength of transcriptomics is that one can identify the clusters of genes and pathways that are perturbed by chemical exposure and that are associated with an adverse health effect over time (Loughery et al., 2018a). To better understand effects at the mechanistic level, transcriptomics analyzes were applied to identify molecular pathways that were altered in fathead minnow (Pimephales promelas) after acute exposure to phenanthrene, on both dose and temporal scales. Female fathead minnows were exposed to an average measured concentration of 0, 29.8, 389, or 943μg phenanthrene L 1 for 24, 48, and 72h in a static renewal bioassay. The authors inferred that transcriptome networks associated with hepatic lipid metabolism are rapidly affected by phenanthrene, and this may indirectly reduce the resources available for reproduction (Loughery et al., 2018a, b). Anyway, more studies are necessary to validate this hypothesis. The same authors developed a chronic experiment where adult male and female fathead minnows were exposed to an average measured concentration of 202μg phenanthrene L 1 for 7 weeks. Transcriptomic responses to phenanthrene exposure suggested a reduction in vitellogenin mRNA, and lipid metabolism and immune system pathways. Phenanthrene exposure affects the energy homeostasis pathways, adversely affecting reproductive efforts, as impaired reproduction has been observed in other studies investigating PAHs (Loughery et al., 2018a). Colli-Dula et al. (2018) through transcriptome analyses investigated if a low dose of BaP could alter genes and key metabolic pathways in the liver and testis in male adult tilapia, and whether these could be associated with Oil and derivatives 159 biological disruption. The authors used both high-throughput RNA sequencing to assess whole-genome gene expression following repeated intraperitoneal injections of 3mgkg 1 of BaP (every 6 days for 26 days) and morphometric endpoints (condition factor (k), gonadosomatic index (GSI), and hepatosomatic index (HIS) as indicators of general health). The study showed that GSI, HSI, and K indices were significantly lower in tilapia that had been exposed to BaP as compared to the control group (unexposed). The RNA-Seq analysis also revealed a larger number of altered genes in the liver than in testis as a result of BaP exposure. In the liver, 1444 genes were significantly altered (607 upregulated and 836 downregulated), while in the testis is only 309 showed any changes (167 up and 142 down). Gene ontology (GO) analysis of the liver data indicated potential effects of BaP exposure on mitosis, cell cycle (G2-M transition), microtubule binding, and condensed chromosome kinetochore both of which are associated with a response to DNA damage. In testis, GO analysis suggested that BaP exposure had a significant impact on the regulation of proteolysis, response to cytokine extracellular space, response to cytokine, the inflammatory response, and hemoglobin complex. The adverse effects on adult male tilapia morphometric indices are indicators of overall poor health and decreased reproduction capacity of a fish, and RNA-Seq analysis on liver and testis showed that molecular changes linked to morphometric effects were caused by BaP exposure. The PAHs also induce differential transcriptional responses in fish development as demonstrated by Goodale et al. (2013). They analyzed transcriptional responses to PAH exposure in zebrafish embryos exposed to benz(a) anthracene (BAA), dibenzothiophene (DBT), and pyrene (PYR) at concentrations that induced developmental malformations by 120h postfertilization (hpf). The exposure to DBT, PYR, or BAA caused a significant increase in the incidence of abnormal embryos compared to the vehicle control exposure at 120hpf. All three compounds induced pericardial edema, snout, and jaw malformations in zebrafish embryos. By 24hpf, a disruption of ion transport, muscle function, and metabolism by DBT and PYR were observed in zebra embryos. The ion transport biological process contained the largest number of misregulated transcripts at 24hpf. The genes involved in fatty acid biosynthesis, steroid biosynthesis, and oxoacid metabolism were also primarily underexpressed. At 48hpf, many of the most significant processes misregulated were related to embryonic development, and the region- alization, neurogenesis, and central nervous system development functions together highlight the widespread disruption of neurodevelopment. 160 Helen Sadauskas-Henrique et al.

The gene expression and transcriptome analyses provide new insight into the effects of the petroleum derivatives in aquatic biota (Hook et al., 2018). In summary, gene expression through specific RNA detection has become one of the most robust tools of molecular biology. Gene expression biomarkers are meant to evaluate the effects of the observed pollution on the exposed biota or to provide an early warning in cases in which anthropogenic impact is suspected. They also provide an excellent instrument to monitor the recovery of formerly polluted sites in remediation, using differential expression genes and identifying biological pathways involved in the response of fishes to oil and derivatives.

7.4 Effects of oil and derivatives on mollusks and crustaceans The mollusks and crustaceans comprise an important ecological and economic resource in freshwater, marine, and estuarine environments, representing potential groups for aquaculture. As with fish, the stocks of these organisms are threatened by overfishing. In addition, their cultivation is normally carried out in open waters, which expose these groups to chemical contaminants (Denadai et al., 2015; Mazurova´ et al., 2008). The organisms that have been extensively used in biomonitoring programs are bivalve mollusks, for example, mussels. Due to their sessile life- style, they accurately reflect local environmental conditions. The mussels are benthic filter feeders with a selective mechanism of suspension feeding. They process relatively large amounts of water during feeding, maximizing their exposure to any harmful material within the water column. This can cause pollutants to accumulate (Fenchel, 1991). The digestive gland and gills are the target organs used as biomarkers for mussels. The digestive gland of mollusks is the main organ for metabolic regulation, participating in the processes of detoxification and elimination of xenobiotics (Marigo´mez et al., 2002). On the contrary, the gills are constantly in contact with the surrounding aquatic environment and thus highly exposed to environmental pollutants due to their large surface and their involvement in gas exchange and feeding (Oliveira David et al., 2008). There is ongoing concern regarding the release of PAHs into aquatic environments and the potential harmful effects on bivalves, even when released at low levels (Frouin et al., 2007; Geffard et al., 2003). However, the detailed detoxification processes of the PAH pollutants on bivalve are unclear, especially their overall response at the level of gene transcription. Oil and derivatives 161

Understanding the effects of PAHs on bivalves is required to effectively establish methods of assessing and predicting the toxicity of PAHs (Cai et al., 2014). Studies carried out with organisms of the phylum Mollusca and subphylum Crustacea demonstrated that the exposure to oil and derivatives, like PAHs, can cause several symptoms, such as changes in branchial structures and functions; changes in the rate of respiration and ammonia excretion; decreases in grazing activity; reduction of the immunity; increases in parasites infections; behavioral, hematopoietic, and biochemical changes; reduction in energy availability for growth and reproduction; changes in survival rates, metabolism, genetic damage, tumor formation; and, in extreme cases, changes in the entire population dynamics through the increased mortality rate after long periods of exposure (Baussant et al., 2009; Connell et al., 1981; Eisler, 1987; Martins and Bianchini, 2011; Mcdowell et al., 1999; Neff, 1979; Neff et al., 1976; Perhar and Arhonditsis, 2014; Vaezzadeh et al., 2019; Yuewen and Adzigbli, 2018).

7.4.1 Physiological responses The exposure to oil and derivative contaminants can impair the interaction of the crustaceans and mollusks with their environment (Blaxter and Ten Hallers-Tjabbes, 1992; Connell et al., 1981; Mazurova´ et al., 2008; Olsen, 2010). The exposure to crude oil and derivatives affects food recog- nition capacity in several gastropods and crustacean species (Blaxter and Ten Hallers-Tjabbes, 1992). For example, Pearson and Olla (1980) demonstrated that the blue crab (Callinectes sapidus) presented behavioral changes when exposed to naphthalene even at low concentrations (10 7 mgL 1). How- ever, when exposed to 2mgL 1 of naphthalene, they presented changes in locomotor activity and defensive behaviors. In addition, when exposed to concentrations above 5mgL 1, they did not exhibit any eating behavior. Kra˚ng (2007), in a study with the amphipod species Corophium volutator, demonstrated that males had difficulties in locating females when exposed to low concentrations (0.5 and 5μgg 1) of naphthalene. Some components of crude oil can stimulate nesting behavior and metamorphosis in barnacle larvae (Anger, 2001). On the contrary, the alarm response of Nassarius obsoletus to an injured conspecific was enhanced at 0.1–0.05mgL 1 of fuel-oil WSF (Blaxter and Ten Hallers-Tjabbes, 1992). The disruption of important behaviors has potential knock-on effects for the fitness and survival of the individual. Although some studies have evaluated this type of response, there are still many behavioral aspects of mollusks and crustaceans 162 Helen Sadauskas-Henrique et al. that need to be assessed in the face of exposure to oil and derivatives. Like- wise, there are countless species with potential for cultivation that can be used in this kind of studies. When evaluating physiological aspects, studies with lobster larval stages (Homarus americanus) exposed to sublethal concentration of crude oil showed that normal patterns of storage, utilization, and synthesis of lipids during larval development and metamorphosis were altered, as well increased rates of protein catabolism. Larvae exposed to crude oil also showed lower concentrations of triacylglycerols and higher concentrations of sterols than animals of the same species not contaminated by crude oil (Capuzzo et al., 1984). When testing the effects of the exposure of Macrobrachium borellii to sublethal concentration of crude oil WSF for 7 days, Lavarı´as et al. (2007) identified that adults and embryos at different stages of development also showed changes in lipid metabolism. The changes identified, in both adults and embryos, were increased activity of the microsomal enzyme palmitoyl- CoA synthetase (ACS), increase in endoplasmic reticulum acylglycerol synthesis, and increase in triacylglycerol synthesis of the intestinal gland. Despite this, the fluidity of the microsomal membrane was only increased when the soluble fraction was tested in vitro. These results demonstrate that hydrocarbons interfere with critical enzymes related to lipid metabolism, interfering with their synthesis and on the regulatory mechanisms. The petroleum hydrocarbons also cause changes in other physiological variables such as hyperglycemia, probably due to the excessive release of crustacean hyperglycemic hormone, as seen in crayfish and crabs exposed to naphthalene (Fingerman et al., 1998; Weis et al., 1992). They cause an increase in glycogen reserves in digestive glands and gonads, as seen in bivalves of the species M. arenaria exposed to different sources of contamination by PAHs (Frouin et al., 2007). They also cause a reduction in the respiratory and ammonia excretion rates, as demonstrated by Capuzzo et al. (1984) in an experiment with lobster larvae H. americanus. On the contrary, a reduction in the filtration rate, respiration, and increases in ammonia excretion rate was found for Crassostrea gigas oysters exposed to 200μgL 1 of PAHs (Kim et al., 2007). In addition, several studies have shown that oil and derivatives act as endocrine disruptors, as was evident in juveniles of C. sapidus exposed to benzene, dimethylnaphthalene, and the oil WSF, and crayfish and crabs exposed to naphthalene. These species had increased the length of the molting cycle and in the intervening period, reduction in growth, delay in the regeneration rate of the limbs, reduction in the resistance of the Oil and derivatives 163 carapace, and reduction in the pigmentation rate (Fingerman et al., 1998; Weis et al., 1992). Likewise, larval stages of H. americanus showed developmental delay and reduced growth due to exposure to sublethal concentrations of crude oil (Capuzzo et al., 1984). Specimens of Procambarus clarkii (crayfish) showed a reduction in the size of the ovaries and degeneration of oocytes after exposure to naphthalene (Fingerman et al., 1998). In the same way, the specimens of Mya arenaria (bivalve) showed a delay in gametogenesis after exposure to various sources of contamination by PAHs (Frouin et al., 2007). In addition, Uca pugilator showed the changes in the dispersion of black pigments in chromatophores, probably due to an inhibition of the release of the black pigment-dispersing hormone from the sinus gland, after exposure to naphthalene (Fingerman et al., 1998; Rodrı´guez et al., 2007). Immunotoxicity, characterized by the reduction of hemocyte counts, neutral red uptake, phagocytosis, bacteriolytic, and antibacterial activity, was identified in the specimens of scallop Chlamys farreri, exposed to different concentrations of BaP (Zhang et al., 2009). The specimens of the bivalve M. galloprovincialis grown in the Pindo region (area most affected by the Prestige oil spill) showed a reduction in weight when compared to the specimens caught in the locations of Miranda and Redes (less affected by the Prestige oil spill) (Peteiro et al., 2006). Similarly, when comparing three sites with different levels of contamination by hydrocarbons, Dissanayake and Bamber (2010) identified high concentrations of pyrene and benzo(a)pyrene metabolites in the urine of C. maenas, as well as low-frequency basal heart rate and low neutral red retention capacity in the hemocyte lysosomes. The specimens of C. maenas tested in the laboratory or captured in contaminated estuaries also showed high concentrations of pyrene metabolites in the urine (Dissanayake et al., 2010), which demonstrated the damage, and the attempt of these organisms to metabolize and excrete these pollutants. Amphipod embryos showed abnormal formation after exposure to crude oil (Yuewen and Adzigbli, 2018). Similarly, larvae of oysters Crassostrea virginica showed time- and dose-dependent abnormalities like atrophy; deformities and/or absence of shells; velums abnormally extended; mantle extrusions and indented shell margins, after exposure to different fractions of crude oil. Larvae of crab Cancer irroratus exposed to WSF of refined oil also showed a delay in growth and development and reduced metabolism, probably due to decreased uptake, food absorption, and reduced energy conversion efficiency. On the contrary, larvae of the crab Rhithropanopeus harrisii, exposed to sublethal concentrations of naphthalene, showed 164 Helen Sadauskas-Henrique et al. increased metabolic rates, but with reduced growth (Anger, 2001). Oysters (C. gigas) when exposed to 200μgL 1 of PAHs (acenaphthene, acenaphthylene, benzo(a)anthracene, BaP, chrysene, dibenzo(a, h)anthracene, fluoranthene, fluorene, phenanthrene, and pyrene) showed the inhibition of feeding capacity that resulted in reduced growth, in addition to the damage to immune responses such as destruction of hemocytes (Kim et al., 2007). Frouin et al. (2007) also demonstrated that the specimens of Mya arenaria show the suppression of immune responses (phagocytic capacity in hemocytes) after exposure to different sources of hydrocarbon contamination, especially coke powder and smelting discharge. The large number of examples shows the influence of these contaminants on the physiology of mollusks and crustaceans. However, it is important to note that the physiological effects of exposure to oil and derivatives are dependent on the ability of the organisms to metabolize and excrete these substances. Apparently, organisms with lipid-rich hepatopancreas will suffer a greater accumulation of hydrocarbons and, consequently, greater effects (Capuzzo et al., 1984).

7.4.2 Biochemical responses The investigation of the biochemical detoxification processes has also been carried out on mollusks and crustaceans exposed to oil and derivatives. O’Hara et al. (1985), for example, demonstrated the inhibition of the cholesterol conversion system to 7-ketocholesterol, 7α- and 7β-hydro- xycholesterol, and cytochrome P450 in hepatopancreas of male crabs (C. maenas) exposed to petroleum hydrocarbons, especially BaP. In contrast, the exposure of L. vannamei to different concentrations of chrysene triggered induction, in gills and hepatopancreas, of AhR, EROD, epoxide hydrolase (EH), GST, sulfotransferase (SULT), uridine diphosphate- glucuronyltransferase (UGT), and SOD, while the total antioxidant capacity (TOSC) and the ratios between the GSH:GSSG have been suppressed. This suggests that the shrimp exposed to chrysene presented the induction of detoxification system in its attempt to excrete the harmful chemical from their body. Moreover, the suppression of the TOSC and the GSH:GSSG ration indicates the failure of the detoxification system, if all ROS are not neutralized. Furthermore, increases in LPO levels, protein carbonyl (PC) contents, and DNA damage could also occur (Ren et al., 2015). Likewise, exposure to BaP and 7,12-dimethyl benzo(a)anthracene triggered activation of benzo(a)pyrene-mono-oxygenase (BaPMO) in bivalves Oil and derivatives 165

M. galloprovincialis, demonstrating the presence of activity detoxification in this species (D’Adamo et al., 1997). Chlamys islandica and M. edulis showed an increase in GST activity of the digestive gland after exposure to low (0.2μgL 1) and high (3.7μgL 1) oil concentrations, respectively. However, only C. islandica showed the changes in the other biochemical biomarkers, like increases in the activity of CAT and TOSC, certain digestive glands, in addition to the reduction in lysosomal membrane stability, determined in hemocytes, which demonstrates greater sensitivity of this species to these contaminants (Baussant et al., 2009). In the study by Baussant et al. (2009), an increase in DNA damage was identified in C. islandica and M. edulis exposed to oil. Despite this, this marker was more sensitive to M. edulis. Carcinus maenas also showed low cell viability and phagocytic capacity of hemocytes (a general indicator of impaired immune function) after exposure in situ and in laboratory to pyrene. Pasquevich et al. (2013) used the prote- omic approach to demonstrate the effects of the WSF of crude oil on the expression of different proteins extracted from the hepatopancreas of prawn Macrobrachium borellii and which are involved in carbohydrate and amino acid metabolism, detoxification, transport of hydrophobic molecules, cellular homeostasis, and among others. Such findings demonstrate that hydrocarbons act at different levels of biological organization, as well as corroborating the variability of responses between species, both in terms of tolerance and responses, in the face of exposure to these contaminants. Dissanayake et al. (2008a, b) demonstrated that the nutritional status influences the susceptibility to hydrocarbons. The adults of C. maenas deprived of food and exposed to 200μgL 1 of pyrene for 7 and 14 days eliminated high concentrations of metabolites of this contaminant and protein in the urine. In addition, in the first 7 days, they showed an increase in antioxidant status and cell function (increased cell viability and decreased phagocytosis); however, after 14 days, these responses significantly decreased. Dissanayake et al. (2008b), in turn, demonstrated that the organism’s life stage also interferes with the response capacity and, consequently, the toxicity of these contaminants. These authors identified that juveniles of C. maenas have less immunocompetence (phagocytosis and cell integrity), metabolic energy (glucose levels in hemolymph), basal heart rate, breathing (at rest), and smaller scope of growth compared to adults after exposure for 7 days at 200μgL 1 of pyrene. In addition, Dissanayake et al. (2011) demonstrated that seasonality completely affects the physiological responses of specimens of C. maenas captured in three estuaries with different levels of PAH contamination. The authors stated that in order to understand the 166 Helen Sadauskas-Henrique et al. effects of hydrocarbons on different species, it is essential that the “normal” seasonal variability of physiological conditions be established. Based on these and other results, there is a need to assess the effects of hydrocarbons in association with different factors that may influence the responses of organisms. Most of the studies carried out with mollusks and crustaceans were done with specimens captured in contaminated areas, such as ports and oil explo- ration sites, and not with cultivated species (Martins and Bianchini, 2011). In any case, the knowledge about the cause and effect relationships between these contaminants and the biological consequences that can occur in these organisms needs further studies. Likewise, studies with cultivable species represent a wide gap in the literature, as well as studies that evaluate the responses of biomarkers in larvae are also rare. There is also a lack of information about the effects of interactions between petroleum hydrocarbons and other organic and inorganic compounds in the environment on species of mollusks and crustaceans, whether they are cultivable or not.

7.4.3 Molecular and genetic responses There is a large group of genes studied in vertebrates and invertebrates in the response of PAH contamination. Although in fish cyp1a gene is widely accepted as an environmental biomarker, there are no data about its expression in mollusks suggesting that P450 isozymes other than CYP1A may be induced after exposure to contaminants (Simpson, 1997). There is a large diversity of cyp genes that also occurs in invertebrates (Guo et al., 2013; Zanette et al., 2013). For bivalves, complete cyp4 cDNAs have been identified in Chlamys farreri (Miao et al., 2011) and Venerupis philippinarum (Pan et al., 2011), whereas partial sequences have been cloned in Unio tumidus (Chaty et al., 2004), Perna viridis (Zhou et al., 2010), and Pinctada martensii (Du et al., 2015). The genes of the cyp family have a different pattern of expression according to the type of contaminant, time of exposure, tissue, and species. For example, in the clam Ruditapes philippinarum quantitative real-time (RT) PCR analysis revealed that there was no notable change in cyp4 mRNA expression in gill after exposure to BaP, while the mRNA expression was induced significantly in the digestive gland of the clam by 0.2μgL 1 BaP (Pan et al., 2011). In P. martensii exposed to different concentrations of pyrene (0, 4, 8, 16, and 32mgL 1) for 7 days, the highest expression level of cyp4 was observed in the 8mgL 1 exposure group after 5 days, which then decreased to normal level after 7 days of exposure. The cyp4 transcripts Oil and derivatives 167 were detected in different tissues (mantle, gill, hepatopancreas, and adductor muscle) examined, the highest band intensities cyp4 were observed in the gill (Du et al., 2015). Other authors observed that the cyp1a-like gene expression patterns are different from cyp3a-like and cyp4 in the ovary of C. farreri exposed to BaP (0.025; 0.5; 10mgL 1) for 10 days (Tian et al., 2014). The authors demonstrate that BaP exposure could induce upregulation of cyp3a-like gene expression in the scallop ovary, and in all groups of BaP treatments induced CYP4 gene expression in a dose- and time-dependent manner (Tian et al., 2014). To understand the mechanism of phenanthrene (PHE) biotransformation, scallops (Nodipecten nodosus) were exposed for 24 or 96h to 50 or 200μg of PHE L 1 (equivalent to 0.28 and 1.12mM, respectively), followed by depuration in clean water for 96h. Increased transcriptional levels of cyp2ui-like, cyp2d20-like, and cyp3a11-like genes were detected in the gills of organisms exposed to PHE for 24 or 96h. The data indicated that N. nodosus is able to induce an antioxidant and biotransformation-related response to PHE exposure, counteracting its toxicity. Depuration caused a decrease in PHE concentration by 2.9- and 2.5-fold after 96h of exposure to 50 and 200μg of PHEL 1, respectively, as well as the restoration to basal levels for most of the parameters affected by PHE exposure (Piazza et al., 2016). Studies with cyp family genes are the most popular, but AhR can be bound and activated by a variety of chemicals (Denison and Heath- Pagliuso, 1998). The study performed by Liu et al. (2010) showed that the AhR mRNA expressions in gills of the clam Ruditapes philippinarum correlated with BaP concentrations, which suggested it can be used as a molecular biomarker of BaP exposure. Genes as p53 and ras have also been cloned in mollusks where they show a degree of homology and conservation with the vertebrate p53 and ras genes, suggesting that the functional role may also be conserved (Ciocan and Rotchell, 2005; Muttray et al., 2010; Sˇtifanic et al., 2009). Studies with the ras oncogene in aquatic organisms exposed to oil products have demonstrated that ras gene is a reliable biomarker to indicate that the contaminant is a potent cancer inductor. The challenge to establish a particular gene as a specific biomarker, however, most genes have different patterns of expression even though their sequences are very conserved. For example, Lima et al. (2008) observed a significant level of polymorphic variation in the M. galloprovincialis in the ras gene that may indicate the presence of a second ras gene in these species. When mussels were exposed to 168 Helen Sadauskas-Henrique et al. different dilutions (0%, 6.25%, 12.5%, 25%, 50%, and 100%), water- accommodated fraction (WAF) of fuel oil over a period of 21 days. The expression of ras gene of M. galloprovincialis decreased in PAH-contaminated samples and 100% WAF-exposed samples compared with reference samples. The ras gene expression was higher in digestive gland samples compared with gonad samples from the same individuals. The effects of BaP in ras gene and p53 gene were also demonstrated in M. edulis mussel exposed to final nominal BaP concentration of 5.6, 56, and 100μgL 1 for 6 and 12 days. Wide interindividual variation of p53 expression was detected in all tissues (gills, adductor muscle, mantle, and digestive gland). After 12-day exposure, significant increased p53 expression was detected in both the mantle and adductor muscle with 5.930.83, and 3.281.23-fold higher expression levels, respectively. For ras gene expression after 12-day exposure, the mantle showed significantly (2.181.14-fold) higher ras expression, but the adductor muscle and digestive gland showed no significant increase in expression. The different patterns of the responses of the genes involve different BaP uptakes and accumulations in tissues that also show interindividual variability (Di et al., 2011). Another group of invertebrates affected by the PAH pollutants are the crustaceans. Emphasizing that the mechanisms of PAH toxicity to crustaceans have been rarely studied, while more information about the mechanisms of PAH toxicity is reported in fishes and mollusks (Liu et al., 2014; Pan et al., 2006). Many studies have shown that crustaceans living in PAH- polluted environments attain high body burdens of PAHs, for example, Carcinus maenas exposed to pyrene resulted in significantly elevated pyrene metabolite concentrations in the urine at 7 and 14 days compared with control individuals (P<0.001), validating contaminant bioavailability. After 14 days of pyrene exposure, antioxidant status and cellular viability of C. maenas were significantly decreased in pyrene-exposed compared to unexposed crabs (Dissanayake et al., 2008b). Recently, the genome sequencing projects together with the increasing use of high-throughput methods generated data that accelerated the identification of crustacean metabolism-related genes and can be used for insights into molecular responses of crustaceans to PAH stress (Hansen et al., 2008). For example, Ren et al. (2014) demonstrated that white shrimp (L. vannamei) exposed to different BaP concentrations (0.03, 0.3, and 3μgL 1) increased cytochrome P450(CYP)1A1, GST, SULT via AhR mRNA expression in a dose-dependent manner, which suggests that they could be potential targets of BaP that disrupt the detoxification system. The Oil and derivatives 169 consistency of their responses to BaP exposure implies that AhR action may be involved in invertebrate CYP regulation (Ren et al., 2014). Zheng et al. (2015) also demonstrated that there is a tendency of positive correlation in the expression levels of cyp4v28 and cyp4v29 genes with the concentration of BaP in Marsupenaeus japonicus. The data showed that the deduced amino acid sequences of the two CYP4s shared high homology with other proteins of the CYP4 family. The two CYP4s were most abundantly expressed in the hepatopancreas and showed dose- and time-dependent expression patterns in response to BaP. Most of the studies with crustaceans important for aquaculture involves only the effects of petroleum derivatives on detoxification enzymes and bioaccumulation (Ren et al., 2015). Genetic studies are scarce and mostly with crustaceans like copepods (Hansen et al., 2008). Zhuang et al. (2017) observed that in the group of copepods, sensitivity to PAHs varies among species. Their results showed that Pseudodiaptomus poplesia was more sensitive to pyrene (LC50-48h: 62.21mgL 1) than the calanoid copepod Acartia tonsa (LC50-48h: 129mgL 1). Copepods are remarkable species as they bridge producers and consumers in transferring energy and toxicants in trophic levels and many studies have reported in the field of environmental genomic and marine ecotoxicology (Lee et al., 2006). Marine copepod Calanus finmarchicus, cyp301a1, cyp302a2, and cyp330a1 genes were cloned, while partial cyp4 genes were identified in the copepod Calanus helgolandicus. Particularly, cf-cyp330a1 gene associated with ecdysteroid synthesis and lipid storage regulation was modulated in response to naphthalene, dispersed oil, and WSF (Hansen et al., 2008, 2009). Tigriopus japonicus and Paracyclopina nana showed a high tolerance in survival in response to WSF. However, they showed the increased expression of several cyp genes (Han et al., 2014). The analysis of the expression patterns of cyp genes upregulated in response to WSF exposure could be used to detect WSF exposure and to compare chronic endpoints such as development rate and hatching rates (Han et al., 2014).

7.5 Interaction of oil and derivatives with water characteristics 7.5.1 Fish Water chemistry can influence the toxicity of contaminants. Recent studies have shown alterations in the final toxicity of oil and derivative compounds, such as PAHs, for organisms living in environments with high DOC 170 Helen Sadauskas-Henrique et al. concentrations, low ion compositions, and pH (Landrum et al., 1984; Moeckel et al., 2013; Sadauskas-Henrique et al., 2016). High PAH bioavailability was reported for fish species (Fundulus heteroclitus and rainbow trout) exposed to PAHs in water with low dissolved ion composition (Ramachandran et al., 2006; Shukla et al., 2007). Higher bioaccumulation of these compounds was observed, promoting activation of physiological and biochemical responses (Ramachandran et al., 2006). Similarly, DOC can influence PAHs’ bioavailability (Landrum et al., 1984; Moeckel et al., 2013; Sadauskas-Henrique et al., 2016) and solubility (Lippold et al., 2008). DePalma et al. (2011) demonstrated that the presence of DOC at low concentrations (0.023g of CL 1) increased the transference of hydrophobic compounds, like PAHs, from a nonaqueous layer (oil) to the water, increasing the bioavailability of the toxic compounds from the oil to aquatic organisms. The DOC can also promote the activation of some biotransformation and antioxidant enzymes in organisms that together with the presence of contaminants as oil and derivatives can impair or overwhelm the detoxication systems and cause damage to organisms (Matsuo et al., 2006; Timofeyev et al., 2006). For example, the acclimation of tambaqui for 10days to both commercial humic acid (Aldrich; AHA) and natural organic matter from Rio Negro (Amazon, Brazil) induced the hepatic EROD and benzyloxyresorufin-O-debenzylase (BROD) activities (Matsuo et al., 2006). Sadauskas-Henrique et al. (2016) observed increased high EROD activity, LPO, DNA damage, and increased absorption of PAHs in tambaqui exposed to crude oil in natural Rio Negro water (DOC concentration of 10mg of CL 1). Fish species (Acarichthys heckelii and Satanoperca jurupari) collected in a harbor (Sa˜o Raimundo, Manaus, Brazil) affected by a Petroleum Asphalt Cement (PAC) spill in Rio Negro (Amazon, Brazil) presented with the increases of EROD activity, PAH metabolites in bile, erythrocyte DNA damage, together with the inhibition of the brain acetylcholinesterase (AChE) activity. The multivariable (PCA-factorial) approach indicated positive correlations between the DOC and PAH concentrations in water and the analyzed biomarkers in both fish species (Sadauskas-Henrique, unpublished data).

7.5.2 Mollusks and crustaceans The association between contamination by hydrocarbons and other biotic and abiotic factors seems to potentiate the effects of hydrocarbons, interfering with functions that normally allowed physiological compensation and Oil and derivatives 171 resulting in subideal conditions or even mortality. Thus, the combination of the effects of stressors will, sooner or later, increase the impacts on populations and communities (Whitehead, 2013). Among the environmental factors, salinity was shown to be a factor of great influence, since the solubility of PAHs increases with decreasing salinity and, consequently, increases bioavailability and toxicity (Whitehead, 2013). Laughlin and Neff (1979) corroborated this information by demonstrating the increased mortality of larvae of the mud crab Rhithropanopeus harrisii exposed to phenanthrene and naphthalene at low salinity. The increase in temperature triggered greater sensitivity of shrimps (Eualus spp. and Pandalus goniurus) exposed to toluene and naphthalene, probably due to the increase in metabolism and the accumulation of these toxic elements in these organisms (Korn et al., 1979). Physiological effects normally triggered by hypoxia can also be enhanced by the presence of hydrocarbons. These pollutants, which are known to cause several morphophysiological changes in gills, block gas exchange at the air–water interface, which in hypoxic conditions can result in mortality (Whitehead, 2013).

7.5.3 Climate change In the last few decades, climate change has been the focus of discussion involving biodiversity (Tao et al., 2018), species distribution (Chen et al., 2011), food production (Thoai et al., 2018), and aquaculture (Lauria et al., 2018). According to the Intergovernmental Panel on Climate Change (IPCC and Pachauri, 2014), climate change will amplify existing risks and create new risks for natural and human systems. Risks are unevenly distributed and are generally greater for disadvantaged people and communities in countries at all levels of development. Increasing magnitudes of warming increase the likelihood of severe, pervasive, and irreversible impacts for people, species, and ecosystems. Continued high emissions would lead to higher negative impacts for biodiversity, ecosystem services, and economic development and amplify risks for livelihoods and for food and human security. In addition to the effects of climate change, the advanced technologies and human development have also increased the release of pollutant products in nature, contaminating many ecosystems, including air, ground, and water pollution. Today’s industrialized society is threatening water quality, releasing a vast amount of harmful xenobiotics, including metals, pesticides, petroleum derivatives, and industrial chemicals, which bioaccumulate in organisms and cause toxicity to aquatic fauna and flora (Vor€ osmarty€ et al., 2010). 172 Helen Sadauskas-Henrique et al.

There is evidence that climate changes are modifying the physical and chemical quality of the ecosystems. These modifications can enhance the capacity of bioaccumulative compounds to enter the food chains. This scenario should bring policy makers to adopt further measures for the reduction and elimination of bioaccumulative compounds from all sources of emissions. At the same time, the assessment of the impacts will require the identification of a suitable set of indicators of the changes in climate, with particular reference to the selection of biota species (Carere et al., 2011). Fish biota are influenced by climate change (Rijnsdorp et al., 2009), and also by pollutants such as oil derivatives (Oyibo et al., 2018; Santana et al., 2018). Oliveira and Val (2017) evaluating the effects of three climate scenarios (A1: current, water temperature of 24.70.2°C and water CO2 of 8.80.9ppm; A1B: intermediate, 2.5°C and 400ppm CO2 above current levels, and A2: extreme, 4.5°C and 850ppm CO2 above current levels) as predicted by the Intergovernmental Panel on Climate Change (IPCC) for 2100, on the growth and physiology of tambaqui. It was observed that the food intake by the fish increased in A1B and A2 scenarios. However, these climate scenarios (A1B and A2) caused lower food usage for growth, so as to endure higher metabolic demand, and the temperature was a limiting factor for fish growth. Scenario A2 also affected the transcriptional responses of tambaqui (Prado-Lima and Val, 2016). The list of differentially expressed genes showed several genes that play roles in energy production, protein folding, maintaining cellular homeostasis, and among other functions. Silva (2016) investigated the effects of climate change in tambaqui injected with BaP (8 and 16μmolkg 1) and subjected to an extreme scenario fore- casted for the Amazon region for 2100. The extreme scenario magnified the fish responses to BaP, increasing the histological damages on the fish liver and DNA strand breaks in blood cells. Implementation of new studies will help to predict how climate changes can affect the fish metabolism and molecular response added to the effects of pollutants such as petroleum products.

7.6 Future perspectives on oil and derivative contamination and aquaculture The contamination of the aquatic environment by oil and derivatives is a reality. Besides the significant increasing effort to prevent oil and derivative spills during the exploitation, transport, and storage, current evidence indicates the decreases in oil leaks and spills. However, the contributions of Oil and derivatives 173 these pollutants to the aquatic environment are still significant. On the contrary, aquaculture continues to grow faster than other major animal food production sectors, and fish is an important protein source consumed by the human population worldwide. The uptake of oil and derivatives by fish is dependent on the environmental bioavailability of the compounds like PAHs, as well as the physiology of the organisms as well as physical and chemical characteristics of water (temperature, salinity, pH, presence of DOC, exposure time, and seasonal changes). The accumulation of PAHs in aquatic organisms may constitute a risk to human health due to the consumption of contaminated fish. The carcinogenic, mutagenic, and bioaccumulative capacities of PAHs have been reported by the WHO, FAO, SCF, IARC, EFSA, and USEPA (Ledesma et al., 2016). In this sense, it is necessary to maintain a high degree of environmental quality because of multiple ecologic feedback loops linking human health and food production.

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Ecotoxicological effects of microplastics and associated pollutants

Fábio Vieira de Araújoa, Rebeca Oliveira Castrob, Melanie Lopes da Silvac, and Mariana Muniz Silvab aFaculdade de Formaçaõ de Professores, Universidade do Estado do Rio de Janeiro, Saõ Gonçalo, RJ, Brazil bPrograma de Po´s Graduaçaõ em Biologia Marinha e Ambientes Costeiros, Universidade Federal Fluminense, Niteroí, RJ, Brazil cLaborato´rio de Ecologia e Dinâmica Bêntica Marinha, Faculdade de Formaçaõ de Professores, Universidade do Estado do Rio de Janeiro, Saõ Gonçalo, RJ, Brazil

8.1 Introduction Due to its properties of lightness, low cost, impermeability, and thermal and electrical insulation, plastics are present in all sectors of human life. This diversity of use leads to higher consumption and consequently higher production of plastics, increasing from 1.5 million tons in the 1950s to 335 million tons in 2016. However, most plastics have only single-use purposes and are promptly discarded (Alimba and Faggio, 2019). Durability combined with the intense use and mismanagement of these items makes it the most common residue found among solid wastes that pollute terrestrial and aquatic environments around the world (Walker, 2018). Contamination by plastic debris in the marine environment was first reported 50 years ago (Law et al., 2010) when the presence of these materials was detected in the Sargasso Sea by Carpenter and Smith (1972). Since then, many studies have emerged and the scientific community has sought ways to analyze its impact on the environment. Although plastic is rarely biodegradable, it undergoes physical, chemical, and biological processes that fragment it into smaller pieces. Based on their sizes, plastics are classified as macroplastics (>5mm), microplastics (5–100nm), and nanoplastics (<100nm) (Cole et al., 2011; Rezania et al., 2018; Thompson et al., 2004). The plastic debris are ubiquitous in the marine environment and have been reported on the sea surface (Suaria et al., 2016; Zhu et al., 2018), along coastlines (Browne et al., 2011; Liebezeit and Dubaish, 2012) and on the seabed (Woodall et al., 2014),

Aquaculture Toxicology © 2021 Elsevier Inc. 189 https://doi.org/10.1016/B978-0-12-821337-7.00009-8 All rights reserved. 190 Fábio Vieira de Araújo et al. accumulating in different ecosystems (Co´zar et al., 2014, 2017; Suaria et al., 2016; Van Sebille et al., 2015), including remote regions. The small size of micro- and nanoplastics increases their adsorption area and their ability to cross cell membranes and organelles and together with their wide distribution in the marine environment make them highly harmful pollutants (Barnes et al., 2009; Browne et al., 2008; Engler, 2012; Thompson et al., 2004). Zooplankton (Desforges et al., 2014), benthic animals (Boerger et al., 2010), and filter-feeding animals (Germanov et al., 2018) may ingest such small-sized plastics because they are similar in size to their normal food items. Studies have shown that microplastics can enter at different levels of the trophic chain (Anderson et al., 2017; Kanhai et al., 2018; Obbard, 2018; Schmidt et al., 2018; Wen et al., 2018a; Xiong et al., 2018) and suggest that organisms can accumulate them (Li et al., 2015). In addition, they carry chemical additives in their composition and are able to adsorb the pollutants of the environment, such as the persistent organic pollutants (POPs) (Koelmans et al., 2015) and metals (Wang et al., 2017), potentializing their toxic effects in organisms that feed on them. Throughout their dispersion by the marine environment, plastics, microplastics, and nanoplastics are colonized by prokaryotic and eukaryotic microorganisms, forming a microbial biofilm, a community known as “plastisphere” (Curren and Leong, 2019; Miao et al., 2019). Thus, the microplastics are complex structures of polymers with chemical additives, as well as having adsorbed organic pollutants and colonized microorganisms (Galloway et al., 2017). In vitro studies show that microplastic ingestion may cause potent effects on the physiology of different organisms, including those grown in aquaculture as demonstrated by Hanachi et al. (2019) in showing that the microplastics are present in fish meal used as feed and can be transferred to cultured organisms raising concerns for aquaculture and also on human health (Barboza et al., 2018a; Galloway, 2015; GESAMP, 2016; Wright et al., 2013a, 2017), which has aroused global concern (Holmes et al., 2012; Teuten et al., 2007). Therefore, this chapter aims to show the current state of knowledge of the issue of microplastics and associated pollutants in aquatic environments, as well as some of the possible ecotoxicological effects caused by these associations on biota.

8.2 Impacts of microplastic on marine animals Biota can be physically, biologically, and chemically impacted by microplastics. Physical and biological impacts result from intentional direct ingestion (when animals eat plastic instead of food), unintended direct Ecotoxicological effects of microplastics and associated pollutants 191 ingestion (performed by filter animals), or indirect ingestion through trophic transfer between prey and predators. As a result, internal damage to animals reported includes the accumulation of microplastics in the gut cavity, in the digestive tubules, and in the circulatory fluid. These accumulations may result with blockages throughout the digestive system; ulceration, and rupture of the digestive tract; modification of intestinal function; anomalous embryonic development; inhibition of larval locomotion; and decreased feeding capacity and energy reserves (Wright et al., 2013b). Chemical impacts are caused by chemical compounds adsorbed by microplastics during their residence in the environment (Mato et al., 2001; Teuten et al., 2007, 2009) or added to plastics during their manufacturing (Berge et al., 2012; Net et al., 2015).

8.3 Plastic additives Plastic additives are chemical compounds added to plastic that function as plasticizers, flame retardants, stabilizers, antioxidants, and pigments. Phthalates, bisphenol A (BPA), nonylphenol (NP), and brominated flame retardants (BFRs) are examples of plastic additives (Berge et al., 2012; David et al., 2009; de Boer et al., 1998; de los Rı´os et al., 2012; Mackintosh et al., 2004; Net et al., 2015; Xie et al., 2005, 2007). The type of additive used depends on the plastic polymer and the requirements of the final product (Hermabessiere et al., 2017). Besides leaching from macro- and microplastics, plastic additives are directly discharged into the marine environment from industrial and municipal wastewater and are thus common in the sea (Berge et al., 2012; David et al., 2009) and sediments (Klamer et al., 2005; Verslycke et al., 2005). Several authors have been addressing the contamination of marine organisms by different types of plastic additives such as phthalates in plankton and macroalgae (Mackintosh et al., 2004) and fish (Cheng et al., 2013) and NP in oysters (Cheng et al., 2006), mussels (Isobe et al., 2007), and fishes (Ferrara et al., 2008). Laboratories and field studies have shown that the ingestion of microplastics is the main source of plastic additives transfer to organisms. The consequences observed include a higher mortality rate; decrease in the reproduction capacity; physical effects like immobilization; behavioral effects; carcinogenesis, reproductive toxicity, abnormal inflammatory, or immune responses; and developmental disorders of the brain or nervous system (Murata and HunKang, 2018). Table 8.1 summarizes the studies described in this topic. Table 8.1 Summary of studies reporting microplastics and their effects in biota Type of Biota plastic Effects observed Studied area References Invertebrate Mussel PS Accumulation in gut cavity, digestive tubules, United Browne et al. Mytilus edulis and in the circulatory fluid, oxidative stress Kingdom (2008) Mytilus edulis PS Production of pseudofeces, reduction in The Wegner et al. filtering activity Netherlands (2012) Mytilus edulis HDPE Accumulation in lysosomal tissue, Helgolang, Von Moos et al. granulocytoma formation, lysosomal Germany (2012) membrane destabilization Mytilus galloprovincialis PE and PS Toxic effects related to immune response, Italy Avio et al. oxidative stress, genotoxicity (2015) Scrobicularia plana PS Neurotoxicity, genotoxicity, oxidative stress Portugal Ribeiro et al. (2017) Crustacean PS Reduction of reproductive fitness, highest South Korea Lee et al. (2013) Tigriopus japonicas mortality rate Calanus helgolandicus PS Reduction of reproductive fitness, highest United Cole et al. mortality rate Kingdom (2015) Hyalella azteca PE and Reduction of reproductive fitness, highest United States Au et al. (2015) PP mortality rate Amphibalanus amphitrite HDPE, Highest mortality rate United States Li et al. (2016) LDPE, PP, PVC, PC, PET, and PS Paracyclopina nana PS Developmental delays, reduced fecundity South Korea Jeong et al. (2016) Amphibalanus amphitrite and Artemia PS Alteration of swimming activity, oxidative Italy Gambardella franciscana stress, neurotoxicity et al. (2017) Gramma ruspulex, Hyalella azteca, PS Reduction in growth The Redondo- Asellus aquaticus, Sphaerium corneum, Netherlands Hasselerharm Lumbriculus variegatus, Tubifex spp. et al. (2018) Daphnia magna PS Oxidative stress China Zhang et al. (2019) Daphnia magna PP, PE, Highest mortality rate, immobilization Italy Renzi et al. PVC, (2019) PVC/PE Polychaeta Inflammatory response, depletion of lipid United Wright et al. Arenicola marina PVC reserves Kingdom (2013a, b) Arenicola marina PS Reduction of feeding activity The Besseling et al. Netherlands (2013) Arenicola marina PE and PS Increase in energy consumption French, Van Belgian, and Cauwenberghe Dutch North et al. (2015) Sea coast Echinodermata PS Oxidative stress, genotoxicity, growth delay, Italy Della Torre Paracentrotus lividus embryotoxicity et al. (2014) Tripneustes gratilla PE Impact on larval growth and development Australia Kaposi et al. (2014) Lytechinus variegatus (larvae) PE Anomalous embryonic development Brazil Nobre et al. (2015)

Continued Table 8.1 Summary of studies reporting microplastics and their effects in biota—cont’d Type of Biota plastic Effects observed Studied area References Vertebrate Fish PE Neurotoxicity Portugal Oliveira et al. Pomatoschistus microps (2012) Pomatoschistus microps PE Neurotoxicity (reduced acetylcholinesterase Portugal Oliveira et al. activity) (2013) Pomatoschistus microps PE Neurotoxicity Portugal Luı´s et al. (2015) Pomatoschistus microps PE Reduction of feeding activity Portugal de Sa´ et al. (2015) Dicentrarchus labrax PE Highest mortality rate France Mazurais et al. (2015) Danio rerio (larvae) PS Genotoxicity, inhibition of larval locomotion, China Chen et al. oxidative stress, alterations on body length, (2017) nervous, and visual system (nanoplastics) Danio rerio (larvae) PS Immune response, toxicity pathways for lipid The Veneman et al. metabolism, oxidative stress, growth delay Netherlands (2017) Danio rerio and Caenorhabditis elegans PA, PE, Intestinal damage (separation of villi and China Lei et al. (2018) (nematode) PP, PVC, dehiscence of enterocytes), growth delay, and PS highest mortality rate, reduction of reproductive fitness, reduced calcium levels, oxidative stress Danio rerio and Scenedesmus obliquus PS, APS, No significant effects China Liu et al. (2019) and CPS

APS, amine-modified; CPS, carboxyl-modified; HDPE, high-density polyethylene; LDPE, low-density polyethylene; PA, polyamide; PE, polyethylene; PET, polyethylene terephthalate; PP, polypropylene; PS, polystyrene; PVC, polyvinylchloride. Ecotoxicological effects of microplastics and associated pollutants 195

8.4 Microplastic and persistent organic pollutants (POPs) POPs are chemical compounds capable of persisting in the environment, moving through the air, water, soil, and sediment and accumulating at levels that can harm terrestrial and aquatic life. These pollutants may be of natural or anthropogenic origin and have a specific combination of physical and chemical properties that, when released into the environment, remain unchanged for long periods of time, being resistant to photolytic, and chemical and biological degradation. The physicochemical properties of POPs allow these compounds to be adsorbed to particles, such as microplastics (El-Shahawi et al., 2010). The adsorption of POPs to microplastics may occur during the manufacture with the addition of chemical additives or in the environment through sewage, surface runoff, and landfill leaching. These associated pollutants can be carried over long distances, be ingested by marine animals, and then passed along the food chain. The microplastics can also transport POPs from surface water to the sediment, thus increasing the exposure of benthic organisms to the pollutants (Anderson et al., 2016; Besseling et al., 2013; Germanov et al., 2018; Hartmann et al., 2017; Herzke et al., 2016; Mato et al., 2001; Rodrigues et al., 2019; Syberg et al., 2015; Wang et al., 2015, 2017). The influence of microplastics on the accumulation of POPs in living organisms is a current issue. Some studies indicate that microplastic ingestion does not provide a significant contribution to the transfer of chemicals from water to biota through the feed, especially in environments with high concentrations of POPs. The authors who defend this theory point out that experiments with alarming results tend to use a high concentration of microplastics, which is not representative of the reality found in the environments (Bakir et al., 2016; Lohmann, 2017). One of the first studies to analyze the toxicological effect in the laboratory of microplastic (polystyrene) ingestion by marine organisms was carried out by Besseling et al. (2013) exposing a polychaete species (Arenicola marina) to sediments contaminated with polystyrene and PCBs. The results showed a positive correlation between: (1) the microplastic concentration and the bioaccumulation of the PCBs and (2) reduction in feeding activity and weight loss in the animals. Some authors considered seabirds as environmental sentinels and have studied their organs and tissues to find POPs and to correlate with the 196 Fábio Vieira de Araújo et al. ingestion of plastic. These studies are important because they point to the health of marine animals. Tanaka et al. (2013) found toxic compounds in the adipose tissues of dead birds and in microplastics present in the stomach contents of these birds, without, however, finding these compounds in their natural prey (pelagic fish), indicating in this way the microplastics as vectors of these compounds. Herzke et al. (2016), also studying dead birds, did not find a positive correlation between the presence of POPs in the bird’s tissue and plastic in the stomach contents. However, these authors pointed out that plastics may adsorb substances other than POPs and be potentially hazardous to marine animals. The experiments with fish and other marine animals showed that the animals fed with diets containing plastics associated with POPs presented higher concentrations of these compounds in the tissues and toxicological effects more prominent when compared to the animals fed with only virgin plastics or only POPs. It is because digestive fluids can help in removing sorbed POPs on contaminated microplastics. Thus, the longer that ingested microplastics stay in the gut of organisms, the more likely it will be that any sorbed POPs will transfer and become incorporated into body tissue (Chua et al., 2014). Rochman et al. (2013) offered to a group of fish (Oryzias latipes), a diet-containing virgin polyethylene, and to another group, polyethylene associated with chemical pollutants (polycyclic aromatic hydrocarbons, PAHs; polybrominated diphenyl ethers, PBDEs) adsorbed from the Bay of San Diego (San Diego, CA, USA). The fish exposed to the polyethylene associated with the chemical compounds presented hepatic stress, including the reduction of glycogen, the formation of fatty vacuoles, cell necrosis, and in a specimen of this group, a hepatocellular adenoma was found occupying 25% of the liver. Among the fish that ingested the virgin polyethylene, signs of hepatic stress were also observed, but at lower levels than the others. In this group, one specimen presented an eosinophilic focus of cell alteration, a precursor to tumorigenesis. Corroborating the hypothesis that the plastic besides serving as a vector of POPs increases the bioaccumulation, Chua et al. (2014) exposed amphipods (Allorchestes compressa) to different diets: (1) virgin microplastic (polyethylene); (2) POPs (PBDEs); (3) microplastic+POPs; and (4) common diet. As a result, they found that PBDEs adsorbed in microplastic were assimilated into the tissue of marine amphipods, demonstrating that the microplastic can transfer POPs from the water to the organism. Jeong et al. (2018) conducted an experiment with marine zooplankton (Brachionus koreanus) to estimate the impacts of nanoplastic and microplastic Ecotoxicological effects of microplastics and associated pollutants 197 ingestion associated with BDE-47 (PBDE—flame retardant) and TCS (triclosan—antimicrobial). The authors found a greater accumulation of nanoplastic than the microplastic in the tissue of the animals, and as a consequence, it was observed the oxidative stress in cell membranes. Nanoplastics are known to become more “bioreactive” than microplastics, once the increased total surface area facilitates chemical reactivity with higher surface area to volume ratios (Nel et al., 2009). The contact with the nanoplastic generated the inhibition of resistance proteins to multiple drugs, indicating the toxicity of nanoplastics at the molecular level and the importance of the synergistic effects of microplastics when associated with other POPs of the environment. The results suggest that toxicological interactions between microplastics and POPs present higher toxicity than microplastics and POPs separately. PBDE was also assimilated by bivalves (Bellas et al., 2014; Johansson et al., 2006; Ramu et al., 2007), fish (Peng et al., 2007), crustaceans (Verslycke et al., 2005), and mammals (de Boer et al., 1998), reducing physiological fitness and inducing antioxidant activity. Paul-Pont et al. (2016) exposed four groups of mussels (Mytilus spp.) to different diets for 7 days: (1) control (diatoms—Chaetoceros muelleri); (2) diatoms with fluoranthene (PAH); (3) diatoms with micropolystyrene; and (4) diatoms with micropolystyrene and fluoranthene; afterward, the animals were placed in clean water for 7 days to depurate or clear the contaminants from their tissues. Although fluoranthene presented a high affinity with polystyrene microparticles, this did not influence the bioaccumulation of POPs in the analyzed organisms. After depuration, higher fluoranthene concentration was observed in mussels exposed to the microplastic and fluoranthene than in mussels exposed only to fluoranthene. The greatest histopathological damages were observed in group 4. Diet 3 led to an alteration in cellular oxidative balance and significant increases in hemocyte mortality, in reactive oxygen species production in hemocytes, in antioxidant enzymes, and in glutathione (related enzymes in mussel tissues). The authors concluded that the presence of microplastics induced toxic effects on tissues at the cellular and molecular levels. Batel et al. (2016) fed a crustacean (Artemia sp.) with fluorescent microplastics (polyethylene) containing PAH (benzo(a)pyrene and ethoxyresorufin) and afterward offered it as food to zebrafish (Danio rerio), simulating a trophic chain to verify the transfer of POPs along the chain. The use of fluorescently labeled PAH molecules demonstrated the release of PAH from microplastic to Artemia and consequently PAH was ingested by the fish. In this experiment, the microplastics played the role of vectors and donors of POPs along the trophic chain. 198 Fábio Vieira de Araújo et al.

Batel et al. (2018) also studied the patterns of accumulation and transfer of POPs (PAH—benzo(a)pyrene) associated with microplastics to zebrafish by the gills and to their embryos, through fluorescently labeled polymers indicating that microplastics did not accumulate permanently in large quantities in the adult gills after 6 or 24h, adhering only to the layer of mucus in the filaments, which is excreted. On the surface of the embryos, there was an accumulation of smaller and heavier microplastics (1–5mm) from the cho- rion or water column. In both the exposure scenarios, it was possible to observe the transfer of POPs from the microplastic to the animals through fluorescence microscopy; however, no morphophysiological disorders or changes in developmental patterns were detected. Rainieri et al. (2018) also conducted a study with zebrafish in which they provided four types of food for 3 weeks to groups of animals: (1) untreated food; (2) ration supplemented with microplastics; (3) ration supplemented with microplastics and a mixture of PCBs, BFRs, PFCs (perfluorinated compounds), and methylmercury; and (4) food supplemented only with the contaminant mixture. The feed supplemented with microplastics and contaminants produced the most evident effects, especially in the liver, besides significantly altering the homeostasis of the brain, intestine, and muscle. Microplastics and contaminants alone did not produce relevant effects on zebrafish under the experimental conditions tested. The accumulation of POPs and changes observed in the fish muscle tissues deserve to be highlighted in this experiment since this is generally the portion of a fish eaten by people. Granby et al. (2018) showed that Dicentrarchus labrax fed with contaminated microplastic presented a greater accumulation of PCBs and BFRs, increasing up to 40 days the detectable residue of chemical compounds in the fish tissues. Analysis of liver gene expression after 40 days of exposure indicated that microplastics increased the toxic effects related to liver metabolism, immune system, and oxidative stress. Peda et al. (2016) submitted three groups of European sea bass (Dicentrarchus labrax) to different types of diet for 90 days: (1) control; (2) polyvinyl chloride (PVC); and (3) PVC removed from a site contaminated by PAH, hexachlorocyclohexane (HCH), dichlorodiphenyltrichloroethane (DDT), and PCB. Intestinal histological analyzes performed after 30, 60, and 90 days showed a structural and functional alteration in the intestine, especially in the distal part in groups 2 and 3. After 30 days of exposure, the intestine provided the first defense strategy, mucus secretion, and an increase in the number of goblet cells. Subsequently, the apical vacuolation of enterocytes and the fusion of villi occurred. The enterocyte damage increased proportionally to increased Ecotoxicological effects of microplastics and associated pollutants 199 exposure. After 60 days, circulatory changes and worsening of inflammatory processes were observed. The last analysis, after 90 days of exposure, showed severe histopathological lesions, reduction of visceral fat, reduction of intestinal functions, and generalized intestinal dysfunction, with group 3 presenting more severe cases. Gerdes et al. (2019) evaluated the effect of microplastic on the removal of PCB (18, 40, 128, and 209) from planktonic cladocerans (Daphnia magna), exposing them to high PCB loads (Phase 1); to a mixture of microplastics and algae (Phase 2-group 1); and only to algae (Phase 2-group 2-control). In cladocerans fed with microplastic, PCB 209 was removed more efficiently than in cladocerans fed with algae alone, while there was no difference for the other types of PCBs. The reduction of POP load on microplastic-fed animals after PCB (group 1) resulted in an improved reproduction. The authors showed that the effects of microplastic on lipid metabolism seem to affect the accumulation of contaminants and growth performance. Microplastics and nanoplastics, once in the tissues of an animal, can travel through the gastrointestinal membranes via mechanisms similar to endocy- tosis, then enter the circulatory system, cause dysregulation in gene expression, induce oxidative stress, eliciting immunological responses, genomic instability, disruption in the endocrine system, neurotoxicity, and reproductive abnormalities (Alimba and Faggio, 2019; Bakir et al., 2016; Germanov et al., 2018; Prokic et al., 2019; Rodrigues et al., 2019). Due to these characteristics along with their persistence in the environment, bioaccumulation, and long-range transport, Lohmann (2017) emphasized that microplastics and nanoplastics should also be considered as POPs. Excessive use and rapid and inappropriate disposal make plastic impacts similar to POPs such as DDT or PCB. After their arrival in the ocean, microplastics are hardly removed, accumulate in organisms and sediments, and persist for a long time, negatively impacting the environment and organisms. Primary microplastics also deserve special attention, since they are used in several countries in the products of cleaning and exfoliating cosmetics and have a great ability to accumulate POPs on their surface, as demonstrated by Napper et al. (2015) in an experiment performed with polyethylene present in cosmetics. Thetypeofplasticpolymerusedas raw material should also be taken into account, since polyethylene particles can absorb and accumulate more organic chemicals than polypropylene and PVC pellets, for example (Lohmann, 2017; Napper et al., 2015; Teuten et al., 2007; Worm et al., 2017). Table 8.2 summarizes the studies described in this topic. Table 8.2 Summary of studies reporting POPs in microplastics and their effects in biota. Biota Type of plastic POPs Effects observed Studied area References Invertebrate Rotifera Microbeads PBDEs Oxidative stress in lipid membranes, South Korea Jeong et al. Brachionus koreanus PS (BDE-47 and multixenobiotic resistance (2018) (0.05, 0.5, and TCS—triclosan) conferred by the P-glycoprotein; 6μm) inhibition of multiple drug resistance proteins Polychaeta Microplastic PCBs Loss of weight and reduction in Tanaka Arenicola marina PS feeding activity The Netherlands et al. (400–1300μm) (2013) Amphipod Microplastic PBDEs Assimilation of the pollutant into Queenscliff, Chua et al. Allorchestes compressa (11–700μm) the tissue Australia (2014) Mussel Microplastic PAH Histopathological damage Bay of Brest, Paul-Pont Mytilus spp. PS France et al. (2–6μm) (2016) Crustacean Microplastic PCBs Alteration in lipid metabolism and Sweden Gerdes Daphnia magna 4Æ1μm reproductive gain et al. (2019) Vertebrate Fish Pellets PAHs, PCBs, Hepatic stress including reduction San Diego Bay, Rochman Oryzias latipes LDPE PBDEs of glycogen, formation of vacuoles United States et al. in lipid cells, cell necrosis, and (2013) hepatocellular adenoma Dicentrarchus labrax Italy Table 8.2 Summary of studies reporting POPs in microplastics and their effects in biota—cont’d Biota Type of plastic POPs Effects observed Studied area References Pellets PAH, HCH, Worsening of inflammatory Peda et al. PVC DDT, PCB processes, reduction of visceral fat, (2016) (0,5 mm) reduction of intestinal functions, and total bowel compromise Danio rerio Plastic and PAH (benzo[a] Adhesion of the microplastic to the Germany Batel et al. microplastic pyrene) mucus layer in the filaments of the (2018) PE gills; accumulation of microplastics (1–5mm) on the surface of the embryos Dicentrarchus labrax Pellets PCBs, BFRs Increased permanence of chemical Denmark Granby LDPE compounds in the body of animals, et al. (125–250μm) increased hepatotoxic effects, (2018) changes in the immune system, and oxidative stress Danio rerio Microplastic PCBs, BFRs, Change in homeostasis of the liver, Denmark Rainieri LDPE PFCs, brain, intestine, and muscle et al. (125–250μm) methylmercury (2018) Bird Plastic and PBDEs Plastic like vector of pollutant Japan Tanaka Puffinus tenuirostris microplastic (3– et al. 10mm) (2013) Fulmarus glacialis Plastic and PCBs, DDTs, Negative correlation between Norway Herzke microplastic PBDEs POPs and plastic in the stomach et al. (1mm) contents (2016)

Continued Table 8.2 Summary of studies reporting POPs in microplastics and their effects in biota—cont’d Biota Type of plastic POPs Effects observed Studied area References Trophic chain Benthonic Microplastic DDT, PHE, Low correlation between the United Kingdom Bakir et al. invertebrate, fish, and PVC e PE DEHP presence of microplastic and the and the (2016) seabirds (200À250μm) transfer and accumulation of Netherlands pollutants in the food chain Crustacean (Artemia sp.) Plastic and PAH Microplastics as a vector and donor Germany Batel et al. and Zebrafish (Danio microplastic of POPs along the trophic chain (2016) rerio) PE (1–5mm and 10–20mm)

BFRs, brominated flame retardants; DDT, dichlorodiphenyltrichloroethane; DEHP, di-2-ethylhexyl phthalate; HCH, hexachlorocyclohexane; LDPE, low-density polyethylene; PAH, polycyclic aromatic hydrocarbons; PBDEs, polybrominated diphenyl ethers; PCBs, polychlorinated biphenyls; PE, polyethylene; PFCs, perfluorinated compounds; POPs, persistent organic pollutants; PS, polystyrene; PVC, polyvinylchloride. Ecotoxicological effects of microplastics and associated pollutants 203

8.5 Microplastics and metals The early research focusing on the investigation of plastics associated with metals began in 2010 since it was previously believed that plastics would be inert to these materials (Ashton et al., 2010). Nowadays, it is known that the microplastic itself functions as a source of metals, releasing adsorbed metals. Metals such as Al and Fe can be adhered to the pellets due to contamination during abrasion manufacturing which is caused by friction in the equipment (Hoffman et al., 1991). Other metals such as Cd, Mn, Pb, Sn, and Zn can be used as additives to provide different properties to plastic products such as antioxidants, polymer stabilizers, flame retardants, pigments, and plasticizers (Michaeli, 1995; Rodriguez, 1996). Although most metals present in the microplastics are inherent to the manufacturing process, other studies have shown that they can also come from the environment, increasing the risk of contamination (Wang et al., 2017). The sources of metals in the aquatic environment are several, such as antifouling paints that diffuse in water and industrial waste effluents (Almeida et al., 2007; Deheyn and Latz, 2006), more common in environments with strong anthropogenic pressure such as ports and marinas (Brennecke et al., 2016; Vedolin et al., 2018). Cu is a metal that is the basis of a biocidal pigment present in coatings used in ship hulls and other structures that hinder the biofouling process (Canning-Clode et al., 2011; Claisse and Alzieu, 1993). The adsorption capacity of metals to microplastics is due to changes in their surface such as cracks, burrs, crushing, and also the presence of bacterial biofilms, which increase the surface of anionic sites for adsorption of metals. This process is driven by the need for the electrical charge of the surfaces to remain neutral (Brennecke et al., 2016; Richard et al., 2019). The different plastic polymers are environmentally friendly and interact with metals in different ways. Although properties of plastics may influence metal adsorption, the weathering process is the main driver (Holmes et al., 2014), as described in the following discussion. Rochman et al. (2014) observed no significant difference in metals’ adsorption by distinct plastic polymers although in other experiments, the polymers adsorbed metals differently (Brennecke et al., 2016; Gao et al., 2019). Ashton et al. (2010) and Turner and Holmes (2015) demonstrated that the adsorption capacity of the metals is higher in the aged plastics when compared to the virgin (new) plastics (Brennecke et al., 2016; Holmes et al., 2012, 2014; Marsˇic-Lucic et al., 2018; Prunier et al., 2018; Turner and 204 Fábio Vieira de Araújo et al.

Holmes, 2015; Vedolin et al., 2018), probably due to weathering and accumulation of organic matter. Much of the work developed to date has focused on quantifying the presence of metals in microplastics and few have focused on the toxicological effects of these metals on organisms. Some experiments carried out in the laboratory investigated the bioaccumulation and the possible effects of metals associated with microplastics in small crustaceans of the species D. magna (Kim et al., 2017); in the bivalves Corbicula flumıńea (Oliveira et al., 2018) and Mytilus galloprovincialis (Rivera- Hernańdez et al., 2019); and in the fishes D. labrax (Barboza et al., 2018b) and Symphysodon aequifasciatus (Wen et al., 2018a, b). Kim et al. (2017) tested the effects of nickel (Ni) associated with polystyrene plastic polymers with one carboxyl group (PS-COOH) and another without this functional group (PS). The authors observed greater immobilization of D. magma exposed to Ni combined with PS-COOH than Ni combined with PS, since this functional group influences the bond with the metal. Barboza et al. (2018b), besides demonstrating that the presence of the microplastic mixture with the metal (Hg) influenced the increase of Hg concentration in the fish tissues, also pointed to the significant inhibition of acetylcholinesterase (AChE) brain activity and a significant increase of lipid oxidation levels in brain and muscle. These are considered to be the combined effects of mercury-microplastic exposure. In Corbicula flumıńea, the microplastic metal mixture caused a reduction of the filtration capacity in the animal, induced stress, and lipid peroxidation that led to lower animal activity (Oliveira et al., 2018). However, the works of Wen et al. (2018a, b) with juvenile fish and Rivera-Hernańdez et al. (2019) with mussels mention that the adverse effects of metals were lower when they were associated with microplastics. Wen et al. (2018a, b) observed an inverse relationship between cadmium concentration and microplastics concentration (Cd accumulation decreased when microplastics concentration increased), without causing adverse effects on growth and survival of the studied organisms. Rivera- Hernańdez et al. (2019) working with the mercury retention rate (Hg) in mussels observed that this metal was less retained when associated with microplastics because: (1) some of the microplastics may not have been ingested; (2) the ingested microplastics were eliminated by the feces because they are nonnutritive particles that can be rejected in the stomach preventing their entry into the digestive gland; and (3) high affinity of Hg on the surface of microplastics, which means that Hg was mainly eliminated in conjunction with microplastics. Although these results do not demonstrate adverse effects Ecotoxicological effects of microplastics and associated pollutants 205 on the organisms studied, the risk of microplastics as metal vectors should not be overlooked. Compared with other topics, studies with metals in microplastics are still limited. The importance of microplastics analyzed as a reservoir of metals on beaches is relatively low compared to other contaminants such as POPs, even in beaches that are heavily polluted (Gao et al., 2019; Vedolin et al., 2018). It is worth mentioning that the number of plastics can increase certain metals and pollutants in the environment through the leaching process of their additives (Gouin et al., 2011). Table 8.3 summarizes the studies described in this topic.

8.6 Microplastics and microorganisms: The plastisphere According to Zettler et al. (2013) and Oberbeckmann et al. (2015), plastic particle colonization by microorganisms in the marine environment occurs rapidly. This is driven by the biological adhesion of the microorganisms to the hydrophobic surface of the plastic. Biofilm formation on the surface of plastics is important for the degradation process of the plastic (Artham et al., 2009), since it can protect it from UV radiation and consequent direct action of photocatalysis (Andrady, 2011). Biofilm can also indirectly protect plastic particles from photocatalysis as it increases their size and weight, thereby decreasing their buoyancy, accelerating their sedimentation to the seabed (Salta et al., 2013), and moving it away from the action of UV rays (Andrady, 2011). However, studies have shown that microorganisms themselves can accelerate the degradation process of plastic (Balasubramanian et al., 2010; Zettler et al., 2013). Biofilm formed on floating plastics is composed mainly of diatoms, cyanobacteria, and coccolithophores (Briand et al., 2012; Oberbeckman et al., 2014; Reisser et al., 2014). However, Zettler et al. (2013), de Tender et al. (2017), and Oberbeckmann et al. (2018) state that several prokaryotic families are responsible for forming the general community of plastic biofilm. Among these families, we can mention Flavobacteriaceae, Erythrobacteraceae, Hyphomonadaceae, and Rhodobacteraceae found in several environments studied (North Sea, Baltic Sea coast, multiple sites in the North Atlantic, and freshwater systems) (Zettler et al., 2013; de Tender et al., 2017; Oberbeckmann et al., 2018). Furthermore, Zettler et al. (2013) observed that the microbial communities present in the biofilm formed in the plastic debris differed from the microbial communities found in the surrounding waters, calling these communities of “plastisphere.” Amaral-Zettler et al. (2015) and Oberbeckman Table 8.3 Summary of studies reporting metals in microplastics and their effects in biota. Studied Biota Type of plastic Metals Effects observed area References Invertebrate Bivalve Microplastics Mercury (Hg) Low filtration rate, cellular stress due Portugal Oliveira et al. (2018) Corbicula flumı´nea (0.13mg/L) (30μg/L) lipid peroxidation Mussel Microplastics Mercury (Hg) Hg incorporated through Spain Rivera-Herna´ndez Mytilus (mean size 10mg Hg/L microplastics was quickly eliminated et al. (2019) galloprovincialis 12.78μm) in the biodeposits Vertebrate Fish Microplastics Nickel (Ni) Immobilization of D. magna exposed United Kim et al. (2017) (PS and PS- to Ni combined with PS-COOH was States COOH) higher Dicentrarchus labrax Microplastics Mercury (Hg) Microplastics influence the Portugal Barboza et al. (2018a) (0.26 and (0.010 and bioaccumulation of mercury, causing 0.69mg/L) 0.016mg/L) neurotoxicity, oxidative stress, and changes in the activities of energy- related enzymes Symphysodon Microplastics Cadmium (Cd) Oxidation damage, increase in MDA China Wen et al. (2018a, b) aequifasciatus (0, 50, or (0 or 50mg/L) and PC contents. Stimulates innate 500mg/L) immune responses due to increased LZM, ACP, and ALP

ACP, acid phosphatase; ALP, alkaline phosphatase; LZM, lysozyme; MDA, malondialdehyde; PC, protein carboxyl; PS, polystyrene. Ecotoxicological effects of microplastics and associated pollutants 207 et al. (2014) observed it too and stated that the composition of the plastisphere is directly influenced by spatial and seasonal effects and varies with the composition of the substrate. Webb et al. (2009) showed that the microbial community is able to colonize plastics of different compositions, such as PET (polyethylene terephthalate), PVC (polyvinyl chloride), and PS (polystyrene). Among the different populations forming the plastisphere, several studies (Barboza et al., 2018a; Lagana et al., 2018; Silva et al., 2019) point to the presence of pathogens and bacteria carrying antimicrobial resistance genes, which, due to the widespread distribution of plastic waste and its constant drift along with seawater, can pose a serious risk to human health and marine biota.

8.6.1 Pathogens dispersed by floating plastic debris Large and small plastic debris can act as a substrate for pathogenic microorganisms and parasites, as demonstrated in several studies carried out in both marine and freshwater environments (Vethaak and Leslie, 2016). However, most studies that mention pathogens in plastic waste did not have the main objective to detect these pathogens, but to describe the microbial community (plastisphere) present in these debris, correlating it with the geographic location, chemical composition of the residue, and seasonal variation (Oberbeckman et al., 2014; Zettler et al., 2013). If the 16S rRNA gene sequencing technique, commonly used in these works, allows determining the composition of these communities, it is limited in its ability to provide the necessary taxonomic resolution for the detection of human pathogens (Harrison et al., 2018)(). The presence of virulence genes is one of several virulence factors investigated to evaluate the degree of pathogenicity of a particular strain. Silva et al. (2019) used this tool through the PCR technique to determine different pathotypes of Escherichia coli isolated from plastic samples collected in Guanabara Bay in Brazil. Thus, most works performed with the plastisphere reported the presence of potential pathogens without actually assessing the degree of virulence of these pathogens. Despite this, Keswani et al. (2016) argue that plastic debris can increase the overall risk of human and animal diseases since several microorganisms associated with diseases have already been described as being part of the plastisphere. Maso´ et al. (2003) were the first authors to suggest that plastic waste would be one of the main problems to marine environments, when 208 Fábio Vieira de Araújo et al. observing the colonization of this debris by the dinoflagellates Ostreopsis sp. and Coollia sp. and by temporary cysts and vegetative cells of Alexandrium taylori, all responsible for harmful algal blooms (HABs). Casabianca et al. (2019) also investigated the presence of toxin- producing microalgae in plastic debris. In their samples, the diatoms were the dominant group, being Pseudo-nitzschia spp. the most abundant genus. Species of this genus produce domoic acid (DA), a neurotoxin that causes amnesia poisoning (ASP) (Trainer et al., 2012). The dinoflagellates Ostreopsis cf. ovata, producer of ovatoxins (OVTXs) and pal- iotoxins (PLTX); and Alexandrium pacificum and Alexandrium minutum, producers of PSP neurotoxin (paralytic shellfish poisoning), were also observed. According to Kiessling et al. (2015), these toxins can bind to other chemical compounds that are adsorbed by the plastic debris and enter the trophic chain via filtering or grazing organisms, causing the contamination of these shellfish, with a potential of being eaten by people and causing illness. Corals can also be affected by contaminated plastic debris. Goldstein et al. (2014) found in the samples of plastic debris collected in the western Pacific large amounts of the folliculinid ciliate Halofolliculina spp., a coral pathogen. Potentially pathogenic bacteria can also be found in plastic debris. These include those that cause diseases in marine organisms and those that cause diseases in humans (Barboza et al., 2018a; Silva et al., 2019). Photobacterium rosenbergii, also associated with diseases and coral bleaching, was found by Curren and Leong, (2019) in the southern strait of Singapore, endangering coral communities (Pu, 2016). These corals are of significant value because of their high biodiversity and conservation priority (Pu, 2016), but could be negatively impacted by the plastic-bacteria contamination. Cyanobacteria Phormidium sp. and Leptolyngbya sp. are other examples of potential pathogens to marine invertebrates, having been found in greater numbers in plastic debris than in the surrounding seawater (Dussud et al., 2018). Potential fish pathogens have also been described in the plastisphere. Oberbeckmann et al. (2016) and Dussud et al. (2018) detected the sequences of the 16S rRNA gene from Tenacibaculum spp., the etiological agent of tenacibaculosis, a bacterial disease that affects many commercial marine fish species, associated with plastic waste in seawater. Aeromonas spp., a genus that has several fish pathogens species, has also been detected in the fragments of microplastics. Virsˇek et al. (2017) detected in these debris Aeromonas salmonicida, a pathogen that affects populations of salmonids, causing furunculosis, a disease that causes sepsis, hemorrhages, muscle injuries, inflammation of the small intestine, enlargement of the Ecotoxicological effects of microplastics and associated pollutants 209 spleen, and death. McCormick et al. (2014) detected the presence of this bacterium in plastic samples in rivers. These same authors found a great abundance of bacteria of the family Campylobacteraceae, known to cause human gastrointestinal (GI) infections, colonizing microplastics near a sewage treatment plant. The fact that these pathogens have been detected in samples of plastic debris present in rivers emphasizes the risk that they can be brought to marine environments, especially microorganisms resistant to saline water, due to the high dispersion of this debris (McCormick et al., 2014). Among all the bacteria detected in the plastisphere, the genus Vibrio spp. seems to be dominant. This genus comprises several species pathogenic to humans, fish, and mollusks, some of them (V. coralliilyticus, V. harveyi, V. splendidus, V. parahaemolyticus, V. alginolyticus, and V. fluvialis) already detected in microplastics (Dussud et al., 2018; Foulon et al., 2016; Kirstein et al., 2016; Schmidt et al., 2014). Vibrios spp. pathogens cause severe losses in fish, mollusk, and shrimp farming (Austin and Austin, 2012), limiting the development of industry prepared to preserve wild fisheries (FAO, 2016). Some species such as V. cholerae, potential human pathogens, have also been detected in plastic waste (Silva et al., 2019). The dominance of this genus in the plastisphere can be explained by its rapid growth in optimal conditions (Franzellitti et al., 2019), suggesting that microplastics can constitute a niche for vibrios, influencing its population dynamics (Foulon et al., 2016). E. coli, Stenotrophomonas maltophilia, Bacillus cereus, Arcobacter, Colwellia, and Pseudomonas were also found in abundance in microplastic samples, all of which may have possible pathogenic strains in humans (Curren and Leong, 2019; Harrison et al., 2014; Silva et al., 2019; Vethaak and Leslie, 2016). The higher concentration of pathogens in the plastic debris in relation to surrounding waters, as observed in several studies, results in the favorable conditions that the biofilm offers (Bryant et al., 2016; Dussud et al., 2018; Ferrara et al., 2008). Although densely populated and polluted areas show an increased risk of contact with contaminated floating plastic debris, their high dispersion in aquatic systems and the increase of plastic pollution worldwide, especially in developing countries with poor waste management, make microplastics colonized by pathogens a possible problem to the population and organisms present in sites with no history of pollution (Keswani et al., 2016; Silvaetal.,2019). Table 8.4 summarizes the studies described in this topic. Table 8.4 Summary of studies reporting biofilm formation with the presence of potential pathogens in plastics and microplastics. Organisms Sample type Type of plastic Pathogens infected Studied area References Sampling at Plastic debris Dinoflagellates Various La Fosca Beach, Maso´ et al. sea surface Not specified Ostreopsis spp. and Coolia sp. Spain (2003) Sampling at Plastic debris Bacteria Various Northwestern Schmidt et al. sea Not specified Vibrio scophthalmi, V. ichthyoenteri, V. Atlantic (2014) anguillarum, V. splendidus, V. gigantis, V. coralliilyticus, V. harveyi Sampling at Microplastics Bacteria Various The North McCormick surface water Not specified (Aeromonas salmonicida) Shore Channel et al. (2014) in urban river in Chicago, USA Sampling at Plastic debris Protozoan Coral Pacific Ocean Goldstein et al. sea surface Not specified Halofolliculina spp. (2014) Incubation in Plastic debris Bacteria (Tenacibaculum spp.) and Fish North Sea off the Oberbeckmann seawater PET Archaea UK coast et al. (2016) Microbial eukaryote, fungi Sampling at Microplastic Vibrio parahaemolyticus, V. fluvialis, Various North and Baltic Kirstein et al. sea surface PP, PE, PS V. alginolyticus Sea (2016) Sampling at Microplastic Bacteria Fish Slovenian coast Virsˇek et al. sea surface PP, PS, PE (Aeromonas salmonicida) (from Piran Bay (2017) to Koper Bay) Adriatic Sea Table 8.4 Summary of studies reporting biofilm formation with the presence of potential pathogens in plastics and microplastics—cont’d Organisms Sample type Type of plastic Pathogens infected Studied area References Incubation in Plastic Polyester (PHBV); PE; Bacteria and Archaea Fish; Mediterranean Dussud et al. seawater PE additivated with (Tenacibaculum spp. crustacean Sea (2018) prooxidant, and artificially Phormidium sp., and Leptolyngbya sp) and aged invertebrate Sampling at Microplastics Vibrio aestuarianus, V. splendidus Oyster Bay of Brest, Ferrara et al. sea surface PE, PP France (2008) Sampling at Plastic Bacteria Various Guanabara Bay, Silva et al. sea surface PE, PP, PET (Vibrio spp. Brazil (2019) Escherichia coli) Sampling at Plastic debris Microalgal Various Mediterranean Casabianca et al. sea surface PE, PP (Pseudo-nitzschia spp., Sea (2019) Ostreopsis cf. ovata; Alexandrium pacificum) Sampling at Microplastics Bacteria Various Coastline of Curren and sediment Not specified (Photobacterium rosenbergii; Singapore Leong (2019) surface Vibrio spp. and Pseudomonas)

LDPE, low-density polyethylene; PE, polyethylene; PET, polyethylene terephthalate; PHBV, poly(3-hydroxybutyrate-co-3-hydroxyvalerate); PP, polypropylene; PS, polystyrene; PVC, polyvinylchloride. 212 Fábio Vieira de Araújo et al.

8.6.2 Microplastics and antimicrobial resistance genes (ARGs) One aspect that has just begun to be studied in relation to plastic debris is their ability to disperse resistance genes. If the presence of pathogens in the plastisphere still leaves doubts as to the spread of infectious diseases to humans (Barboza et al., 2018a), studies have shown that bacteria associated with microplastics have a greater capacity to transfer genes between bacteria taxonomically distant when compared to free-living bacteria (Imran et al., 2019). The growth of multidrug-resistant pathogens (MDRs) has decreased the efficiency of antibiotics, making them a major threat to human health (Knapp et al., 2017). Studies (Hu et al., 2016; Knapp et al., 2011; Su et al., 2015) suggest that this growth, previously credited only to the indis- criminate and erroneous use of antibiotics, is enhanced by the impact of various anthropogenic activities that release antibiotics and metals into the environment, since Stepanauskas et al. (2006) and Bednorz et al. (2013) suggest that metals play a key role in the selection of antibiotic resistance. Plastic particles can adsorb on their surfaces various compounds such as DDT, PCBs, PAHs, drugs, and metals (Guo et al., 2018; Hahladakis et al., 2018; Li et al., 2018; Revel et al., 2018). Thus, the presence of metals and antibiotics and a high bacterial density in the biofilm, which favors gene transfer mechanisms, make this substrate a real “hot spot” for the evolution of mul- tiresistant pathogens to antibiotics (Imran et al., 2019). A study carried out at Antarctica showed that bacteria present in the plastisphere of polystyrene stranded on the banks of King George Island, presented multiple resistance to antibiotics such as cefuroxime, cinoxacin, ampicillin, amoxicillin+ clavulanic acid, among others. It has been suggested that plastic may serve as a vector for the dissemination of multiple antibiotic resistance genes (ARGs) in marine environments and, consequently, may be dispersed to other environments (Lagana et al., 2018). During their stay in aquatic ecosystems, the microbial community present in the microplastics can come into contact with natural populations, promoting processes of horizontal gene transfer. This poses increasing risks to human health and marine organisms, since multi- resistant pathogens present in the microplastic can reach unpolluted environments and transfer their resistance genes to nonpathogenic microorganisms, rapidly spreading resistance to antibiotics (Arias-Andres et al., 2018). The presence of plastic debris in marine ecosystems is a real, and increasing problem and their persistence and dispersal capacities in these environments are widely known. Although current evidence indicates Ecotoxicological effects of microplastics and associated pollutants 213 an important role of microplastics as vectors for animal and human opportunistic pathogens, transfer between them is still speculative, and understanding the risks of the transmission of pathogens and infectious diseases through seafood consumption will require further investigations. Without such studies, conclusions about the possible responsibility of plas- ticdebrisasavectorforthedissemination of disease-causing organisms may be considered alarmist (Barboza et al., 2018a). The use of techniques that allow the detection of genes associated with virulence may be an interesting option to address this issue (Kirsteinetal.,2016). Therefore, caution should be taken, since evidence of pathogenicity in marine animals in relation to the plastisphere is currently lacking ( Jacquin et al., 2019). Thus, studies on the composition of biofilms, identification of virulence factors among potential pathogens, transmission of pathogens to the food chain, and effects on marine and human biota, as well as the study of resistance gene transfer, are extremely important to minimize risks to human health and marine diversity.

8.7 Microplastics and other compounds Various substances such as pharmaceuticals, antibiotics, and radionuclides are present in aquatic environments leaching from terrestrial environments, carried by sewage or by other way. Like POPs and metals, microplastics can adsorb these substances and dispersed it to different regions of the world, endangering the health of aquatic biodiversity (Brennecke et al., 2016). Carbamazepine (Cbz) is an anticonvulsant drug commonly detected in the environment (Archer et al., 2017; Huerta et al., 2018) and has already been reported to cause physiological and biochemical alterations in marine organisms (Luis et al., 2016; Oliveira et al., 2017). Analyzing the effects of nanoplastic polystyrene (PS) individually and combined with Cbz on the Mediterranean mussel (M. galloprovincialis), Brandts et al. (2018) found that Cbz when associated to PS caused a weakening of the mussels immune defense as well as a significant downregulation in gene expression responsible for the repair of cell tissue and DNA strand. Antibiotics are other substances often found in the aquatic environment ( Jiang et al., 2011; Li et al., 2012) and studies have been demonstrated that they may pose a relatively high ecological risk to the relevant aquatic organisms (Xu et al., 2013). Trueba et al. (2008) and Storteboom et al. (2010) showed that exposure to residual trace antibiotics may increase antibiotic 214 Fábio Vieira de Araújo et al. resistance of bacteria and accelerate the development and spread of ARGs. Li et al. (2018) evaluated the adsorption capacities of five types of commonly used antibiotics (sulfadiazine (SDZ), amoxicillin (AMX), tetracycline (TC), ciprofloxacin (CIP), and trimethoprim (TMP)) on five types of microplastics (polyethylene (PE), polystyrene (PS), polypropylene (PP), polyamide (PA), and polyvinyl chloride (PVC)) in the freshwater and seawater systems. They observed that polyamide particles showed the strongest adsorption capacity for antibiotics tested. Higher toxic effects on aquatic life may be expected due to the combined effect of antibiotics and MPs. Lu et al. (2019) demonstrated in their experiment that microplastic is an important reservoir of ARGs in recirculating aquaculture system fish farms that use plastic or fiber- glass fish tanks and that the microbial community diversity of microplastics was higher than that of water. Radionuclides such as Cs and Sr have been widely found in the ocean (Buesseler, 2014). Their association with microplastics is another source of concern for the spread of these contaminants in the environment. Tazaki et al. (2015) detected on the surfaces of plastics recovered from a contaminated freshwater lake, elevated 137Cs, thought to be derived from the Fukushima accident. Johansen et al. (2018) showed that Cs and Sr adsorptions to plastics are generally 2–3 orders of magnitude lower than those for sediments. Although low, this result showed that like sediments, microplastics provide an environmental sink for radionuclides. Besides, these authors suggest that plastics as an absorbing media may be of more importance in the open ocean where mobile and buoyant plastics have the potential to interact with atmospherically deposited 137Cs without competition from sediments. Studies evaluating associations between microplastics and pharmaceuticals, antibiotics, and radionuclides are still scarce and need to be expanded.

8.8 Final considerations Marine biodiversity has a great relevance on our planet. It is the main source of the global primary production; it has many species exploited as fishery resources and organisms with biotechnological potential application. Plastic debris is increasing in our ocean, and consequently, its impacts on biota. Fig. 8.1 presents the main impacts on aquatic biota. More studies are needed to better understand the toxicological effects of microplastics and associated compounds on marine organisms. Future studies should focus on: (1) the analyses of animals collected from the environment; (2) adsorption and effects of other compounds still not studied; (3) exposing the Ecotoxicological effects of microplastics and associated pollutants 215

Fig. 8.1 Ecotoxicological effects of microplastic and associated pollutants on aquatic organisms. microplastics used in the experiments to natural environments so that the levels of contamination are closer to reality; and (4) evaluation of the potential negative impacts on physiological functions of exposed aquatic animal species, e.g., respiration, reproduction, assimilation, and the quality of the offspring. This will improve the understanding of the ecotoxicological effects. Besides carrying out studies, management strategies that reduce plastics in the environment should be a priority in the public policy of countries. Although some projects believe that macroplastics are removable from the marine environment (plasticoceans.org; theoceancleanup.com), mitigating the presence of plastic, especially micro- and nanoplastics (because they are hardly removable in the environment), is essential to prevent future damage to marine biota. 216 Fábio Vieira de Araújo et al.

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Note: Page numbers followed by f indicate figures and t indicate tables.

A crustacean aquaculture, 10 Acetylcholinesterase (AChE), 72, 81, 91–92, farmed species, 4–5t 97f, 204 finfish aquaculture, 6–9 Actinobacteria, 94–96 global aquaculture industry structure, 2–9 Aeromonas spp., 208–209 global production, 3 Agrochemicals governance of, 11–12 aquaculture risk assessment, 82–94, 90f mollusk aquaculture, 9–10 banned from use in agriculture and risk assessment in, 82–94, 90f aquaculture, 98 techniques, systems, and facilities, 5–6 environmental contamination, 82–94, 90f world production, 6 good agricultural practices, 98–101 Aquatic animal species, 3–5 mitigation of, 94–97 Astyanax lacustris,81 bioremediation, 94–96 Atlantic salmon (Salmo salar),64–65, 83–85, nanomaterials, 96–97 117 regulatory process, 98–101 soil contamination, 79–81 B water contamination, 79–81 Bacterial biofilms, 203 Ammonia treatment, in sanitation, 31 Bacterial strains, 94–96 Amphipods (Allorchestes compressa), 196 Benzodiazepines, 120–121 Amprolium hydrochloride, 123 Bioaccumulation, 196, 204 Analgesics, 111–115 Biochemical effects, 62–67 Antibiotics, 121–122, 213–214 Biofilm, 205 Antidepressants, 117–121 Biofloc technology, 8–9 Antifoulants, 26–27 Biofouling, 25–26 ecological risks and regulation, 46–49 Biological effects, 33–34 ecotoxicity and biological effects, 34–43 antifoulants, 34–43 mode of action, 28–29 disinfectants, 43–46 Antiinflammatory, 111–115 Biomarkers, 61 Antimicrobial resistance (AMR), 121 of agrochemical toxicity, 88–91 Antimicrobial resistance genes (ARGs), characteristic of, 66 212–213 Bioremediation, 94–96 Antioxidant mechanisms, 81 Biotic ligand model (BLM), 21 Antiparasitic pharmaceuticals, 122–123 Bithionol, 123 Antipyretic drugs, 111–115 Brazil, 59–60 Apistogramma agassizii,63 Aquaculture, 1–2 C agrochemicals banned from use in Campylobacteraceae, 208–209 agriculture and, 98 Capsaicin, 29 aquatic animal species, 3–5 Carassius auratus,66 chemicals in, 10–11 Carbamate carbofuran, 89

229 230 Index

Carbamazepine (CBZ), 120, 213 EU Rapid Alert System for Food and Feed Carp farming, 6–7 (RASFF) database, 83–85 Chlorantraniliprole, 79–81 European sea bass, 85–86 Chlorine, 30 European Union (EU), 98–99 Chlorothalonil, 28–29 Euryhaline fish, 69 Chlorpyrifos (CPF), 89 Eutrophication, 79–81 Chromium, 63–64 Chronic toxicity, symptoms of, 87–88 F – Clams, 90 91 Fallowing, 7 – Closed containment, 7 8 Finfish aquaculture, 4–5t,6–9 – Clothianidin, 79 81 Fipronil, 89 Codex Alimentarius Commission (Codex), Fishes, 61 – 98 99 Fish meal (FM), 85–86 Coho salmon, 71 Fish, oil and derivatives Contaminated microplastics, 196 biochemical responses, 149–151 – Corbicula fluminea,90 91 fish oil (FO), 85–86 Crustacean aquaculture, 10 molecular and genetic responses Cyprinodon variegatus,63 cytochrome P450 1A (CYP1A) protein, 155–158 D fish genes, 152–155 Danio rerio,64–65 transcriptome fish response and fish Daphnia magna, 118–119 population adaptation, 158–160 Diazepam (DZ), 120–121 p53 gene, 153–155 Dicentrarchus labrax, 63, 85–86 physiological responses Dichlofluanid, 28–29 behavioral and hormonal responses, Dichlorodiphenyltrichloroethane (DDT), 142–144 98 gills responses, 144–147 Diclofenac, 113–114 hematological responses, 148–149 Dipyrone sodium/metamizole sodium, ras oncogene and hif-1a gene, 152–153 114–115 Floating plastics, 205 Disinfectants, 27–28 Fluoxetine (FLX), 117–119 ecological risks and regulation, 46–49 Food safety, 83–85 ecotoxicity and biological effects, 43–46 Formaldehyde, 30 efficacy and characteristics, 28t Formalin, 44 mode of action, 30–33 Dissolved organic matter (DOM), 19 G Generalized additive modeling (GAM), 22 E Global aquaculture industry structure, 2–9 – Ecotoxicology/ecotoxicity, 33–34 Good agricultural practices, 98 101 of agrochemical toxicity, 88–91 antifoulants, 34–43 H disinfectants, 43–46 Harmful algal blooms (HABs), 207–208 Emerging compounds, 29 Heavy metals, 59 Emerging contaminants, 107 Hepatotoxicity, 64–65 Environmental pollutants, 85 Heptachlor epoxide, 98 Essential metals, 59 Hexachlorocyclohexanes (HCHs), 98 17 α-ethinylestradiol (EE2), 115–116 Hormones, 115–117 Index 231

Human consumption, aquatic organisms and pathogens dispersion, by floating plastic food for, 82–94, 90f debris, 207–211 Hydrogen peroxide (H2O2), 30 Microplastics, ecotoxicological effects of, 192–194t, 213–214 I antibiotics, 213–214 carbamazepine (Cbz), 213 Ibuprofen, 111–112 durability, 189 Imidacloprid, 79–81 marine animals, 190–191 Industrial aquaculture, 7 marine biodiversity, 214–215 International Code of Conduct on and metals, 203–205, 206t Pesticide on Pesticide Management, microorganisms, plastisphere, 189–190 99–101 antimicrobial resistance genes (ARGs), International Union of Pure and Applied 212–213 Chemists (IUPAC), 59 pathogens dispersion, by floating plastic Iodophors, 31 debris, 207–211 Iron, 63 persistent organic pollutants (POPs), 195–202 L plastic additives, 191–194 Land-based farming, 8 zebrafish, 197–198 Lead, 64 Mollusk aquaculture, 9–10 Levamisole hydrochloride, 123 Multidrugresistant pathogens (MDRs), Lipid peroxidation (LPO), 63 212 Litopenaeus vannamei, 10, 82–83 Multiple antibiotic resistance (MAR) index, 121 M Macroplastics, 189–190 N Marine animals, 190–191 Nanomaterials, 96–97 Marine ecosystems, plastic debris in, Nanopesticide, 96–97 212–213 Nanoplastics, 189–190, 196–197, 199 Marine zooplankton (Brachionus koreanus), Nanotechnology, 96–97 196–197 Neonicotinoids, 79–81, 83, 93 Markiana nigripinnis,81 Nicotinic acetylcholine receptors Maximum Residue Limit (MRL), 87 (nAChRs), 93 Medetomidine, 29 Nile tilapia, 63–64, 68–69 Mercury, 64–65, 72 N-nitroso compound (NOC), 114–115 Metallothioneins (MT), 66–67 N-nitrosodimethylamine (NDMA), Metals 114–115 behavioral effects, 69–73, 70f Nonsteroidal antiinflammatory drugs bioavailability of, 60–61 (NSAIDs), 111–113 biochemical effects, 62–67 North American Free Trade Agreement biological organization, 61, 62f (NAFTA), 98–99 microplastics, ecotoxicological effects of, North Atlantic Ocean Ecoregion (Portugal), 203–205, 206t 79–81 physiological effects, 67–69 Norwegian Food Safety Authority (NFSA), Microorganisms, plastisphere 85 antimicrobial resistance genes (ARGs), NSAIDs. See Nonsteroidal antiinflammatory 212–213 drugs (NSAIDs) 232 Index

O PCA. See Principal component Oil and derivatives analysis (PCA) aquatic contamination, 133–137 Peracetic acid (PAA), 32–33, 45 biochemical responses, 149–151 Persistent organic pollutants (POPs), 85–86, contamination problem, aquaculture and, 189–190, 195–202, 200–202t 137–140 Pharmaceuticals, 107 fish species, effects on, 140–160 in aquatic environment, 108 interaction with water characteristics exposure effects, in nontarget species, climate change, 171–172 111–123 fish, 169–170 antibiotics, 121–122 mollusks and crustaceans, 170–171 antidepressants, 117–121 molecular and genetic responses antiinflammatory/analgesics/ cytochrome P450 1A (CYP1A) antipyretic drugs, 111–115 protein, 155–158 antiparasitic pharmaceuticals, 122–123 fish genes, 152–155 hormones, 115–117 transcriptome fish response and fish psychoactive pharmaceuticals, 117–121 population adaptation, 158–160 pathway to environment, 109–110 mollusks and crustaceans, effects on Pharmacovigilance program, 99–101 biochemical responses, 164–166 Phenylpyrazole fipronil, 88 molecular and genetic responses, Pimephales promelas,68–69, 72, 116 166–169 Plants, 94–96 physiological responses, 161–164 Plastic perspectives on, 172–173 additives, 191–194 physiological responses debris, 214–215 behavioral and hormonal responses, marine environment, 189–190 142–144 potential pathogens, biofilm formation, gills responses, 144–147 209, 210–211t hematological responses, 148–149 pollution, 209 Olfactory nerve cells, 71 Plastisphere Oncorhynchus kisutch,71 microplastics, ecotoxicological effects of, Oncorhynchus mykiss,64 189–190 Oreochromis niloticus,63–64 antimicrobial resistance genes (ARGs), Organochlorine pesticides (OCPs), 83–87, 212–213 92–93, 98 pathogens dispersion, by floating plastic Organophosphate pesticides, 87–88 debris, 207–211 Organophosphorus compounds (OPs), 92 Polyamide particles, 213–214 Osmoregulation, 68 Polybrominated diphenyl ether (PBDE) Oxidative damage, 89–90 bivalves, 196–197 Oxidative stress, 62–63, 65–67, 90 Polyculture, 6–7 Ozone, 31–32 Polyethylene particles, 199 Pond-based carp culture, 6–7 Principal component analysis (PCA), 22, P 90–91 PAA. See Peracetic acid (PAA) Prochilodus costatus,86 Pacific white shrimp, 82–83 Prochilodus lineatus,63 Pale chub, 66 Prostaglandins, 111–112 Paracetamol, 113 Psychoactive pharmaceuticals, 117–121 Partial least squares (PLS), 90–91 Pyramid solar system scheme, 97, 97f Index 233

Pyrethroids, 93–94 of microbiocides, 40t Pyridine-triphenylborane (TPBP), 29 Toxicokinetic-toxicodynamic (TK-TD) Pyrimethamine, 123 models, 21–22 Toxicology, 17 R intraspecies variation, 20 Radionuclides, 213–214 prediction models, of contaminants Rainbow trout (Oncorhynchus mykiss), 117 biotic ligand model (BLM), 21 Reactive oxygen species (ROS), 62–63 generalized additive modeling Recirculating aquaculture system (RAS), 8 (GAM), 22 Risperidone (RISP), 119 principal component analysis (PCA), 22 S toxicokinetic-toxicodynamic Salmo salar,64–65, 85 (TK-TD) models, 21–22 Seed bivalves, 9–10 water quality criteria/guidelines, 18–20 Selective serotonin reuptake inhibitors Tralopyril, 29 (SSRIs), 118, 120 Trametes versicolor,95 Sewage treatment plants, 109 Trichlorfon, 123 Shellfish, 89, 98 Soil contamination, 79–81 V Sulcotrione, 79–81 Veterinary pharmaceuticals, 109 Vitellogenin (VTG), 116, 118 T Tebuconazole, 79–81 W Teleost fishes, 61 Water contamination, 79–81 Terbuthylazine, 79–81 Water quality criteria/guidelines, 18–20 Thiamethoxam, 79–81 dissolved organic matter (DOM), 19 Thiobarbituric acid-reactive substance dissolved oxygen, 19 (TBARS) levels, 89 hardness and alkalinity, 18 TK-TD models. See Toxicokinetic- pH, 18 toxicodynamic (TK-TD) models salinity, 19 Toxicity temperature, 19 of broad-spectrum antifouling biocides, total dissolved solids, 19 39t emerging antifoulants, 42t Z of fungicides, 38t Zacco platypus,66 of herbicides, 35–36t Zebrafish, 64–65, 119 This page intentionally left blank