Shellfish safety and quality
Related titles: Improving seafood products for the consumer (ISBN 978-1-84569-019-9) Consumer health and well-being may potentially be improved by consumption of health-promoting, safe seafood products of high eating quality. This important collection provides a convenient review of significant findings in key areas of seafood research, with chapters authored by experts in the field. Consumer attitude to seafood products is first discussed, followed by chapters analyzing important advances in the area of the health benefits of seafood, for example the relationship between omega-3 fatty acids and heart disease. The next parts of the book discuss key topics in seafood safety, advances in processing technologies and methods to improve the quality of farmed fish. Improving farmed fish quality and safety (ISBN 978-1-84569-299-5) Fish farming enables greater control of product quality, but there have been concerns about the levels of contaminants found in farmed products. Their sensory and nutritional quality can also not equal that of wild-caught fish. This important collection reviews potential negative safety and quality issues in farmed fish and presents methods to improve product characteristics. The first part of the book discusses contaminants, such as persistent organic pollutants and veterinary drug residues and methods for their reduction and control. The second part addresses important quality issues, such as genetic control of flesh characteristics and the effects of feed on product nutritional and sensory quality. Foodborne pathogens: Hazards, risk analysis and control (ISBN 978-1-85573-454-8) As trends in foodborne disease continue to rise, the effective identification and control of pathogens becomes ever more important for the food industry. With its distinguished international team of contributors, Foodborne pathogens provides an authoritative and practical guide to effective control measures and how they can be applied in practice to individual pathogens. Details of these books and a complete list of Woodhead titles can be obtained by: · visiting our web site at www.woodheadpublishing.com · contacting Customer Services (email:
[email protected]; fax: +44 (0) 1223 893694; tel.: +44 (0) 1223 891358 ext. 130; address: Woodhead Publishing Limited, Abington Hall, Granta Park, Great Abington, Cambridge CB21 6AH, England)
Shellfish safety and quality Edited by Sandra E. Shumway and Gary E. Rodrick
Published by Woodhead Publishing Limited, Abington Hall, Granta Park, Great Abington, Cambridge CB21 6AH, England www.woodheadpublishing.com Woodhead Publishing India Pvt Ltd, G-2, Vardaan House, 7/28 Ansari Road, Daryaganj, New Delhi ± 110002, India Published in North America by CRC Press LLC, 6000 Broken Sound Parkway, NW, Suite 300, Boca Raton, FL 33487, USA First published 2009, Woodhead Publishing Limited and CRC Press LLC ß 2009, Woodhead Publishing Limited The authors have asserted their moral rights. This book contains information obtained from authentic and highly regarded sources. Reprinted material is quoted with permission, and sources are indicated. Reasonable efforts have been made to publish reliable data and information, but the authors and the publishers cannot assume responsibility for the validity of all materials. Neither the authors nor the publishers, nor anyone else associated with this publication, shall be liable for any loss, damage or liability directly or indirectly caused or alleged to be caused by this book. Neither this book nor any part may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, microfilming and recording, or by any information storage or retrieval system, without permission in writing from Woodhead Publishing Limited. The consent of Woodhead Publishing Limited does not extend to copying for general distribution, for promotion, for creating new works, or for resale. Specific permission must be obtained in writing from Woodhead Publishing Limited for such copying. Trademark notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation, without intent to infringe. British Library Cataloguing in Publication Data A catalogue record for this book is available from the British Library. Library of Congress Cataloging in Publication Data A catalog record for this book is available from the Library of Congress. Woodhead Publishing Limited ISBN 978-1-84569-152-3 (book) Woodhead Publishing Limited ISBN 978-1-84569-557-6 (e-book) CRC Press ISBN 978-1-4200-7792-6 CRC Press order number: WP7792 The publishers' policy is to use permanent paper from mills that operate a sustainable forestry policy, and which has been manufactured from pulp which is processed using acid-free and elemental chlorine-free practices. Furthermore, the publishers ensure that the text paper and cover board used have met acceptable environmental accreditation standards. Project managed by Macfarlane Book Production Services, Dunstable, Bedfordshire, England (e-mail:
[email protected]) Typeset by Godiva Publishing Services Limited, Coventry, West Midlands, England Printed by TJ International Limited, Padstow, Cornwall, England
Contributor contact details
(* = main contact)
Editors Sandra E. Shumway* Department of Marine Sciences University of Connecticut 1080 Shennecossett Road Groton, CT 06340 USA E-mail:
[email protected] Gary E. Rodrick Food Science and Human Nutrition University of Florida Building 461 Rm 215 Newell Drive PO Box 110370 Gainesville, FL 32611-0370 USA E-mail:
[email protected]
Chapter 1 Steve Jones Department of Natural Resources and the Environment University of New Hampshire Jackson Estuarine Laboratory 85 Adams Point Road Durham, NH 03824 USA E-mail:
[email protected]
Chapter 2 HeÂleÂne HeÂgaret* and Sandra E. Shumway Department of Marine Sciences University of Connecticut 1080 Shennecossett Road Groton, CT 06340 USA E-mail:
[email protected] [email protected]
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Contributors
Gary H. Wikfors NOAA Northeast Fisheries Science Center Milford Laboratory 212 Rogers Avenue Milford Connecticut, CT 06460-6499 USA
Chapter 5 Gregory L. Boyer Department of Chemistry State University of New York College of Environmental Science and Forestry Syracuse, NY 13210 USA E-mail:
[email protected]
Chapter 3 Albert Bosch* and Rosa M. Pinto Enteric Virus Laboratory Department of Microbiology University of Barcelona Spain E-mail:
[email protected] FrancËoise S. Le Guyader Laboratoire de Microbiologie IFREMER Nantes France
Chapter 4 Monique Pommepuy*, J. C. Le Saux, D. Hervio-Heath and S. F. Le Guyader IFREMER French Research Institute for Exploitation of the Sea Centre de Brest BP 70, 29280 Plouzane France E-mail:
[email protected]
Chapter 6 Per Andersen Orbicon A/S Jens Juuls Vej 16 8260 Viby J. Denmark E-mail:
[email protected]
Chapter 7 Mario Sengco Smithsonian Environmental Research Center PO Box 28 647 Contees Wharf Road Edgewater, MD 21037-0028 USA E-mail:
[email protected]
Chapter 8 Juan Blanco Centro de InvestigacioÂns MarinÄas Apdo 13 Pedras de CoroÂn s/n 36620 Vilanova de Arousa Spain E-mail:
[email protected]
Contributors
xv
Chapter 9
Chapter 12
Wen-Xiong Wang Department of Biology HKUST Clear Water Bay Kowloon Hong Kong E-mail:
[email protected]
Victor Garrido* and Steve Otwell Aquatic Food Products Laboratory Food Science and Human Nutrition Department, University of Florida, Gainesville, FL 32611 USA E-mail:
[email protected];
[email protected]
Chapter 10 Thomas M. Soniat Department of Biological Sciences University of New Orleans Lakefront New Orleans, LA 70148 USA E-mail:
[email protected]
Chapter 11 Serge Corbeil* Commonwealth Scientific and Industrial Research Organisation (CSIRO) AAHL Private Bag 24 Geelong Victoria 3220 Australia E-mail:
[email protected] Franck C. J. Berthe Animal Health and Welfare Unit EFSA Italy
Chapter 13 Douglas I. Watson* Aquaculture and Fisheries Development Centre Department of Zoology, Ecology & Plant Science University College Cork Cooperage Building, Distillery Fields North Mall Cork Ireland E-mail:
[email protected] Sandra E. Shumway and R. B. Whitlach Department of Marine Sciences University of Connecticut 1080 Shennecossett Road Groton, CT 06340 USA E-mail:
[email protected]
Chapter 14 Louis R. D'Abramo* Department of Wildlife and Fisheries Mississippi State University Box 9690 Mississippi State, MS 39759 USA E-mail:
[email protected]
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Contributors
Juan L. Silva and Taejo Kim Department of Food Science, Nutrition and Health Promotion Mississippi State University MS 39762 USA E-mail:
[email protected]
Shaun M. Moss and S. Arce Oceanic Institute 41-202 Kalanianaole Highway Waimanalo, HI 96795 USA E-mail:
[email protected]
Chapter 17 Chapter 15 Marielle C. W. van Hulten* Intervet International BV Wim de KoÈrverstraat 35 5831 AN Boxmeer The Netherlands E-mail:
[email protected] Andrew C. Barnes Centre for Marine Studies Queensland University Australia Karyn N. Johnson School of Integrative Biology Queensland University Australia
Chapter 16 Donald V. Lightner* and R. M. Redman Department of Veterinary Science and Microbiology University of Arizona Tucson, AZ 85721 USA E-mail:
[email protected]
Shaun M. Moss* and Dustin R. Moss Oceanic Institute 41-202 Kalanianaole Highway Waimanalo, HI 96795 USA E-mail:
[email protected]
Chapter 18 Clive Askew Shellfish Association of Great Britain Fishmongers' Hall London Bridge London EC4R 9EL UK E-mail:
[email protected];
[email protected]
Chapter 19 Lorna H. Murray Local Authority Food Law Enforcement Branch Food Standards Agency Scotland St Magnus House 25 Guild Street Aberdeen AB11 6NJ UK E-mail: lorna.murray@ foodstandards.gsi.gov.uk
Contributors Ron J. Lee* Cefas Weymouth Laboratory Barrack Road The Nothe Weymouth DT4 8UB UK E-mail:
[email protected]
Chapter 20 Gary E. Rodrick*, Keith R. Schneider and J. Cevallos Department of Food Science and Human Nutrition University of Florida Building 461, Room 215 Newell Drive PO Box 110370 Gainesville, FL 32611-0370 USA E-mail:
[email protected]
Chapter 21 George J. Flick Jr*, Linda A. Granata, Lori S. Marsh
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Food Science & Technology Virginia Tech (0418) Blacksburg, VA 24061 USA E-mail:
[email protected]
Chapter 22 Gary E. Rodrick* Department of Food Science and Human Nutrition University of Florida Building 461, Room 215 Newell Drive PO Box 110370 Gainesville, FL 32611-0370 USA E-mail:
[email protected] Victor Garrido Aquatic Food Products Laboratory Food Science and Human Nutrition University of Florida Gainesville, FL 32611 USA E-mail:
[email protected]
We dedicate this volume to John W. Hurst, Jr. in recognition of over a half century of dedication to ensuring shellfish quality and safety.
Preface
It is estimated that by the year 2050 the world's population will reach 10 billion people, and seafood, especially cultured shellfish, will play a major role in feeding these populations. Shellfish are a very popular and nutritious food source worldwide and their consumption continues to rise globally. Because of their unique nature as compared with beef and poultry, shellfish have their own distinct aspects of harvest, processing and handling. Guaranteeing shellfish quality and safety is critical for protecting public health as well as for marketing seafood products. This collection of review papers discusses issues of current interest and reviews steps that can be taken by the shellfish industry to maintain and improve shellfish safety and eating quality. The United States Senate recently introduced the Commercial Seafood Consumer Protection Act, legislation to improve the safety of seafood products imported into the United States, and two recent FAO publications, Huss et al. (2004) and Ababouch et al. (2005) provide excellent overviews of international management issues associated with seafood safety and international trade. We believe that Shellfish safety and quality takes these documents a step further, specifically detailing issues related to shellfish. The opening chapters provide an overview of the key issues associated with microbial and biotoxin contamination. Parts II and III then address in more detail methods to improve molluscan shellfish and crustacean quality and safety. Chapters focus on detection of algal toxins, monitoring and mitigation of the effects of harmful algal blooms, metals and organic contaminants, biofouling, disease control and selective breeding. Part IV reviews legislation, regulation, public confidence in shellfish and risk management. Chapters on post-harvest issues, such as depuration, storage and packaging complete the volume.
xx
Preface
Several individuals have helped to make this volume possible. Lynsey Gathercole, Sarah Whitworth and Woodhead Publishing recognized the gap in the available literature and persuaded us to take on the challenge of filling it. A great debt is owed to the authors for providing their time and expert contributions and especially for their patience during the inevitable delays. Max provided countless hours of support and, sadly, will not get to see the final volume. Shellfish safety and quality will be an essential reference for those in the shellfish industry, managers, policymakers and academics in the field.
References and J. RYDER (2005) Causes of Detentions and Rejections in International Fish Trade. Food and Agriculture Organization of the United Nations Fisheries Technical Paper 473, Rome. H.H., L. ABABOUCH and L. GRAM (2004) Assessment and Management of Seafood Safety and Quality. Food and Agriculture Organization of the United Nations Fisheries Technical Paper 444, Rome.
ABABOUCH, L., G. GANDINI
HUSS,
Contents
Contributor contact details . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xiii
Preface . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xix
Part I 1
2
Shellfish safety: an introduction
Microbial contamination and shellfish safety . . . . . . . . . . . . . . . . . . . S. Jones, University of New Hampshire, USA 1.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2 Major microbial contaminants of shellfish . . . . . . . . . . . . . . . . . . . 1.3 Impacts of microbial contamination of shellfish on human health . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.4 Effects of microbial contamination on the international shellfish industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.5 Incidence of microbial contamination in shellfish waters . . . . 1.6 Contamination sources and their identification . . . . . . . . . . . . . . . 1.7 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.8 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 1.9 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biotoxin contamination and shellfish safety . . . . . . . . . . . . . . . . . . . . . H. HeÂgaret, University of Connecticut, USA, G. H. Wikfors, NOAA Northeast Fisheries Science Center, USA and S. E. Shumway, University of Connecticut, USA 2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Origins of phycotoxins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3 3 4 7 10 15 20 27 28 28 43
43 47
vi
Contents 2.3 2.4 2.5 2.6 2.7 2.8 2.9
Part II
Trophic dynamics of phycotoxins in molluscan shellfish . . . . . Human health impacts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Management responses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Economic impacts of harmful algal blooms (HABs) . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Improving molluscan shellfish safety and quality
3
Viral contaminants of molluscan shellfish: detection and characterisation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . A. Bosch and R. M. PintoÂ, University of Barcelona, Spain and F. S. Le Guyader, Laboratoire de Microbiologie, France 3.1 Introduction: human enteric viruses and their fate in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2 Shellfish-borne transmission of virus infections . . . . . . . . . . . . . . 3.3 Effects of viral contamination of molluscs on the international shellfish industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4 Methods for detecting viruses in molluscan shellfish and associated problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.5 Improving detection of molluscan shellfish virus contamination using new molecular-based methods . . . . . . . . . . . . . . . . . . . . . . . . . 3.6 Depuration of viral contaminants in molluscan shellfish . . . . . 3.7 Future trends in virus studies in shellfish . . . . . . . . . . . . . . . . . . . . 3.8 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4
52 57 59 66 67 67 68
Monitoring viral contamination of molluscan shellfish . . . . . . . . . M. Pommepuy, J. C. Le Saux, D. Hervio-Heath and S. F. Le Guyader, IFREMER, France 4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2 Identifying sources of pollution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 Identifying the conditions responsible for microbial contamination of shellfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4 Potential strategies for reducing microbial contamination in shellfish harvesting areas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5 Improving risk management strategies for shellfish harvesting areas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.6 Conclusions and future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.7 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5 Algal toxins and their detection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . G. Boyer, State University of New York, USA 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 Major algal toxins found in shellfish and their sources . . . . . .
83
83 85 88 89 93 95 96 98 108 108 110 112 114 118 120 121 129 129 130
Contents 5.3 5.4 5.5
vii
Current methods for detection of algal toxins in shellfish . . . . New techniques and future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
142 153 154
6 Monitoring of harmful algal blooms . . . . . . . . . . . . . . . . . . . . . . . . . . . . P. Andersen, Orbicon A/S, Denmark 6.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 Action plan design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.3 Regulation of mandatory harmful algal monitoring programmes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4 Methods and techniques used to forecast and monitor harmful algal blooms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.5 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 6.7 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
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7 Mitigation of effects of harmful algal blooms . . . . . . . . . . . . . . . . . . . M. Sengco, Smithsonian Environmental Research Center, USA 7.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2 Novel techniques to mitigate the effects of harmful algal blooms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.3 Ethos of harmful algal bloom (HAB) control . . . . . . . . . . . . . . . . 7.4 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.5 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 7.6 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8
162 164 166 168 170 171 172 175 175 177 190 190 191 191
Modelling as a mitigation strategy for harmful algal blooms . . 200 J. Blanco, Centro de InvestigacioÂns MarinÄas, Spain 8.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 200 8.2 Why model the accumulation of toxins in bivalves? . . . . . . . . . 201 8.3 Historical use and development of toxin/toxicity accumulation models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 204 8.4 Models of the kinetics of accumulation and transformation of toxins in shellfish. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 206 8.5 Applications of modelling for improved shellfish safety and quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 220 8.6 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 221 8.7 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 222 8.8 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 223
9 Metals and organic contaminants in bivalve molluscs . . . . . . . . . . W.-X. Wang, HKUST, Hong Kong 9.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2 Metal concentrations in bivalve molluscs . . . . . . . . . . . . . . . . . . . . 9.3 Internal speciation of metals in bivalve molluscs . . . . . . . . . . . .
228 228 229 233
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Contents 9.4 9.5 9.6 9.7 9.8 9.9 9.10
10
Exposure routes and application of the kinetic model . . . . . . . . Uptake and transfer of metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Safety standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Detection, management, and risk assessment . . . . . . . . . . . . . . . . . Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
234 236 240 241 242 243 243
Managing molluscan shellfish-borne microbial diseases . . . . . . . . T. Soniat, University of New Orleans, USA (formerly of Nicholls State University, USA) 10.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.2 Microbial indicators and pollution-associated pathogens . . . . . 10.3 Enteric viruses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.4 Naturally occurring pathogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.5 Pathogens associated with handling, processing, and distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.6 Management of pollution-associated pathogens . . . . . . . . . . . . . . 10.7 Management of naturally occurring pathogens . . . . . . . . . . . . . . . 10.8 Management of pathogens associated with handling, processing, and distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.9 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.10 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 10.11 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10.12 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
248
11 Disease and mollusc quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . S. Corbeil, Commonwealth Scientific and Industrial Research Organisation (CSIRO), Australia and F. C. J. Berthe, Animal Health and Welfare Unit, Italy 11.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.2 Major pathogens and diseases of molluscs causing significant economic losses in molluscan aquaculture . . . . . . . . . . . . . . . . . . . 11.3 Diagnostic methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.4 Effects of shellfish disease on the international shellfish industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.5 Reducing disease in molluscan aquaculture . . . . . . . . . . . . . . . . . . 11.6 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11.7 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 11.8 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12
Hazard analysis and critical control point programs for raw oyster processing and handling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . V. Garrido and S. Otwell, University of Florida, USA 12.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
248 249 252 254 257 258 259 261 262 263 263 263 270
270 271 279 279 281 283 284 285 295 295
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ix
12.2 HACCP for oyster production and safety . . . . . . . . . . . . . . . . . . . . 12.3 HACCP plan for processing of frozen raw oysters . . . . . . . . . . . 12.4 Hazard analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.5 Identify the critical control points (CCP) . . . . . . . . . . . . . . . . . . . . 12.6 Definition of critical limits (CL) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.7 Designate monitoring procedures . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.8 Corrective action (CA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.9 Specify verification (and validation) procedures . . . . . . . . . . . . . 12.10 Specified records . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12.11 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Appendix: Annex 1 ± examples of HACCP and sanitation records . .
298 300 301 301 306 306 309 309 310 310 311
13 Biofouling and the shellfish industry . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. I. Watson, University College Cork, Ireland and S. E. Shumway and R. B. Whitlatch, University of Connecticut, USA 13.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.2 Biofouling and shellfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.3 Problems and benefits of biofouling . . . . . . . . . . . . . . . . . . . . . . . . . 13.4 Current removal/treatment methods . . . . . . . . . . . . . . . . . . . . . . . . . . 13.5 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.6 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 13.7 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13.8 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
317
Part III 14
317 318 320 325 331 331 332 332
Improving crustacean safety and quality
Optimization of crustacean quality through husbandry and adherence to post-harvest standards for processing . . . . . . . . . . . . L. R. D'Abramo, J. L. Silva and T. Kim, Mississippi State University, USA 14.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.2 Land (site) selection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.3 Water: source, conservation, and preservation of quality . . . . . 14.4 Fertilization and semi-intensive systems . . . . . . . . . . . . . . . . . . . . . 14.5 Formulated feeds, bio-flocs, and intensive pond culture systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.6 Substrate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.7 Water quality management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.8 Collection during harvest . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.9 Harvest and post-harvest treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.10 Safety and quality standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.11 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14.12 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
339 339 340 341 342 344 345 346 348 349 350 357 357
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15
Development of vaccines and management of viral diseases of crustaceans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . M. C. W. van Hulten, Intervet International BV, The Netherlands and A. C. Barnes and K. N. Johnson, Queensland University, Australia 15.1 Introduction: disease and the foundations for preventative healthcare in aquaculture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.2 Using the RNA interface to target shrimp viruses . . . . . . . . . . . . 15.3 Developing vaccines to manage viral disease in shrimp . . . . . 15.4 Using vaccines as part of health management strategies . . . . . 15.5 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15.6 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 15.7 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
16
Specific pathogen-free shrimp stocks in shrimp farming facilities as a novel method for disease control in crustaceans . . . . . . . . . . . D. V. Lightner and R. M. Redman, University of Arizona, USA and S. Arce and S. M. Moss, The Oceanic Institute, USA 16.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.2 The concept of domesticated SPF shrimp: a historical perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.3 Events leading to development of Litopenaeus vannamei as the dominant species in the Americas . . . . . . . . . . . . . . . . . . . . . . . . 16.4 Adaptation of the SPF concept to domesticated shrimp stocks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.5 International Principles for Responsible Shrimp Farming . . . . 16.6 Biosecurity and the culture of wild seed/broodstock . . . . . . . . . 16.7 Biosecurity through environmental control and best management practices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.8 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.9 Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16.10 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
17 Selective breeding of penaeid shrimp . . . . . . . . . . . . . . . . . . . . . . . . . . . . S. M. Moss and D. R. Moss, Oceanic Institute, USA 17.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.2 Selective breeding . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.3 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17.4 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Part IV 18
359
359 365 369 372 375 376 376 384 384 386 388 392 397 413 414 415 415 416 425 426 427 444 445
Regulation and management of shellfish safety
Legislation, regulation and public confidence in shellfish . . . . . . C. Askew, Shellfish Association of Great Britain, UK 18.1 Introduction: public confidence in shellfish . . . . . . . . . . . . . . . . . .
455 455
Contents 18.2 18.3
Hygiene legislation and public confidence . . . . . . . . . . . . . . . . . . . Environmental legislation for the quality of shellfish growing waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.4 Limitations of the regulatory approach . . . . . . . . . . . . . . . . . . . . . . . 18.5 Self-regulation and good management practice (GMP) . . . . . . 18.6 Dietary and health advisories . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.7 Public perception of health benefits and risks associated with shellfish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.8 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.9 The risk-averse marketplace . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18.10 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19 Risk management of shellfisheries . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . L. H. Murray, Food Standards Agency, UK and R. J. Lee, Cefas Weymouth Laboratory, UK 19.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.2 Interaction between public health controls and industry . . . . . . 19.3 Identification of need for improved bases for, and application of, risk management in practice . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.4 Optimising risk management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.5 Improved application of risk management to microbiological and biotoxin problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.6 Official and industry roles in risk management . . . . . . . . . . . . . . 19.7 Future trends . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.8 Interaction of research, legislation and risk management . . . . . 19.9 Shared resources and working together . . . . . . . . . . . . . . . . . . . . . . 19.10 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19.11 Sources of further information and advice . . . . . . . . . . . . . . . . . . . 19.12 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
xi 461 462 464 464 465 469 470 471 472 474 474 476 476 476 479 485 486 497 500 502 502 503
Part V Post-harvest issues 20 Molluscan shellfish depuration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . K. R. Schneider, J. Cevallos and G. E. Rodrick, University of Florida, USA 20.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.2 Types of depuration plant . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.3 Importance of seawater quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.4 Types of seawater treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.5 Rules and guidelines for controlled purification . . . . . . . . . . . . . . 20.6 Depuration plant location, design, and construction . . . . . . . . . . 20.7 Source of shellfish to be depurated . . . . . . . . . . . . . . . . . . . . . . . . . . 20.8 Equipment construction and depuration facility design . . . . . . . 20.9 International depuration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
509 509 510 511 515 519 524 526 529 535
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21
Contents 20.10 Shellfish relaying . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20.11 References and further reading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
537 538
Slaughter, storage, transport, and packaging of crustaceans . . . G. J. Flick, L. A. Granata and L. S. Marsh, Virginia Tech, USA 21.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.2 Slaughter/cooking . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.3 Packaging and preservation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.4 Contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.5 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21.6 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
542 542 543 549 552 560 561
22 Packaging, storage and transport of molluscan shellfish . . . . . . . V. Garrido, Institute of Food and Agricultural Sciences, USA and G. E. Rodrick, University of Florida, USA 22.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.2 Product specification . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.3 Packaging formats and materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.4 Product labeling and tagging . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.5 Product size standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.6 Accepting shellfish shipments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.7 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22.8 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
568 568 568 569 571 574 574 574 575
Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
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Part I Shellfish safety: an introduction
1 Microbial contamination and shellfish safety S. Jones, University of New Hampshire, USA
Abstract: Microbial contamination is a challenging and significant issue for the shellfish industry. It is the main public health concern associated with consuming shellfish, and it often limits shellfish harvesting throughout the world. Enteric viruses, pathogenic Vibrio species, and fecal-borne bacterial pathogens are the main causes of shellfish-borne disease. These microorganisms have widely different properties, sources, virulence factors, and fate in the environment, and the current indicators used to classify harvest waters have significant limitations. A great deal of progress is currently being made in the detection of pathogenic microorganisms and in understanding their fate in the environment. With increasing human development in coastal areas, emerging diseases, habitat destruction, and global climate changes, the challenges associated with managing microbial contamination and shellfish safety continues to evolve. Key words: microbial contamination, shellfish safety, enteric viruses, vibrios, harvest water classification, fecal pollution indicators, microbial survival, pollution sources.
1.1
Introduction
The quality of coastal and estuarine waters throughout the world has become adversely impacted by a variety of contaminants, including microorganisms. In many areas where sewage treatment is inadequate, microbial contamination is by far the most important contaminant affecting shellfish safety. Even in welldeveloped areas, however, microbial contamination from nonpoint pollution remains a critical problem. The studies and results presented here reflect published findings that pertain to the study-specific geographical areas. Different
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Shellfish safety and quality
findings and phenomena could be expected in other areas of the world because of differences, however, in environmental conditions and pollution characteristics. The main themes in this chapter include an initial definition of microbial contaminants and a discussion of those that are the major contaminants in shellfish. Human diseases are then presented to emphasize their significance to the shellfish industry and consuming public. Specific impacts of microbial contamination on the shellfish industry are then discussed, followed by an overview of their incidence in the natural environment and their sources. The final sections include probable future issues and trends in research, and a summary of useful sources for further information.
1.2
Major microbial contaminants of shellfish
Microbial contaminants as defined here are pathogenic microorganisms that cause disease in shellfish-consuming humans. Included are fecal-borne viral, bacterial and protozoan pathogens, and naturally occurring bacterial pathogens, but biotoxin-producing algae are excluded as they are covered in other chapters. It is also necessary to include fecal indicator and organisms used in microbial source tracking (MST). The indicator bacteria used within the shellfish industry are generally fecal (and total) coliforms and Escherichia coli. Enterococci, Clostridium perfringens and other indicators are also useful in helping to elucidate the sources and fate of fecal-borne contamination in shellfish waters (Watkins and Burkhardt, 1996). MST methods have effectively used other organisms such as Bacteroides spp. (Field et al., 2003), Bifidobacterium spp. (Bernhard and Field, 2000) and F-RNA coliphage (Vinje et al., 2004), among others, to track sources of fecal contamination. Though indicator organisms are not necessarily pathogens, their universal use for assessing shellfish safety necessitates their inclusion in this chapter. It is also useful to note that many new pathogen detection methods do not require culturing the target organism and focus instead on the direct detection of species, strain or virulence-specific genes from environmental samples. The issue of microbial contamination and shellfish safety can thus be addressed using methods that involve detection of microbial cells, genetic material, or both. The most frequent causes of shellfish-borne disease (CDC, 2006; Cato, 1998; Rippey, 1994) are of greatest concern. Most shellfish-borne diseases are likely caused by enteric viruses, though pathogenic vibrios are emerging as an increasing threat, and infections from Vibrio vulnificus have the highest fatality rate of any foodborne infectious agent. Fecal-borne bacterial pathogens have become less prevalent worldwide (Rippey, 1994) as a result of successes with their management, though they still account for a significant fraction of shellfish-borne disease in some areas (Cato, 1998). Thus, the major microbial contaminants include viruses, vibrios, and to a lesser extent fecal-borne bacteria.
Microbial contamination and shellfish safety
5
1.2.1 Human pathogenic viruses Viruses are species-specific intracellular parasites and are the leading cause of shellfish-borne disease in humans. Although human pathogenic viruses are readily taken up and accumulated, they do not infect or grow in molluscan shellfish. They can persist for extensive periods in the marine environment (Gantzer et al., 1998; Callahan et al., 1995) and in shellfish (Formiga-Cruz et al., 2002; Lees, 2000). The main source of human viruses is sewage from septic systems, wastewater treatment facilities and direct discharges, all human sources. Relatively high viral concentrations can exist in wastewater treatment facilities (Katayama et al., 2008; Gantzer et al., 1998), and standard treatments such as chlorination are only partially effective in inactivating or removing viruses from effluent (Tree et al., 2003; Tyrrell et al., 1995). Thus, varying yet potentially significant loading of viruses is discharged to shellfish harvest waters. Levels of viruses vary considerably over predictable cycles, with highest levels observed in winter months (Katayama et al., 2008; Formiga-Cruz et al., 2002; Burkhardt et al., 2000; Dore et al., 2000). This seasonal variation in environmental incidence relates well with the incidence of viral infections in humans, and is also influenced by environmental factors and the presence of individuals carrying pathogenic viruses within the human population. Lees (2000) summarized the many common types of human viruses that have been associated with contaminated shellfish, including the rotaviruses, astroviruses, enteroviruses, adenoviruses, hepatitis A, and the calciviruses. Small round structure viruses are a subset of calciviruses that include Norwalk-like viruses (NLV). Viruses most commonly associated with infectious disease incidents through shellfish consumption are NLV and hepatitis A, though they produce different disease symptoms. Viruses are comparatively more difficult to remove from shellfish than most bacteria (Schwab et al., 1998). In addition, considerable effort has therefore been focused on their detection in shellfish because they are also more difficult to detect (Myrmel et al., 2004; Le Guyader et al., 2003; Lees, 2000; Green et al., 1998; Henshilwood et al., 1998; Dore and Lees, 1995). 1.2.2 Pathogenic vibrios The genus Vibrio is comprised of bacteria found free-living in marine environments and at elevated levels in association with a variety of eukaryotic hosts, including shellfish (Thompson et al., 2004) and seaweeds (Mahmud et al., 2007). They play key roles in ecosystem carbon cycling, as an important food source for copepods and as light organ symbionts. In other cases vibrios are pathogens, primarily of shellfish and fish, though the human pathogens are notable. Of particular concern are Vibrio cholerae (Colwell, 2004), Vibrio vulnificus (Gulig et al., 2005; Linkous and Oliver, 1999) and Vibrio parahaemolyticus (Yeung and Boor, 2004), which cause severe diarrheal disease, gastroenteritis, wound infections, and septicemia (Thompson et al., 2004). Their normal association with shellfish presents a common mechanism
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Shellfish safety and quality
for human infection (Morris, 2003; Potasman et al., 2002). V. hollisae and other potentially pathogenic species are of less concern at present. The widespread and consistent detection of V. vulnificus and V. parahaemolyticus in bivalve shellfish suggests shellfish as an important ecological niche for these vibrios. Persistence appears to involve interactions between the vibrios and the shellfish hemolymph that result in low levels of elimination (Pruzzo et al., 2005). The presence of a protozoan pathogen of oysters, Perkinsus marinus, can lead to greater inefficiency of oyster hemolymph for eliminating V. vulnificus (Tall et al., 1999). Levels in oysters may also depend on oyster genotype (Sokolova et al., 2006). The lack of elimination of V. vulnificus may lead to significant growth that may be shed to overlying waters (Tamplin and Capers, 1992). Vibriophage can control pathogenic vibrios in shellfish and can be both numerous and diverse in mollusks (DePaola et al., 1997, 1998; Baross et al., 1978). In fact, oysters may be one of the main reservoirs of bacteriophagecontrolling vibrios, and this phage population may change seasonally (Comeau et al., 2005). Vibriophage density and diversity tracks Vibrio abundance in shellfish but not in sediments, where the bacteria are numerous but the phage are not (Comeau et al., 2005). The presence of pathogenic vibrios in shellfish and overlying waters is most pronounced in the warmer waters of tropical and sub-tropical ecosystems (Martinez-Urtaza et al., 2008; Zimmerman et al., 2007; Thompson et al., 2004; DePaola et al., 2003) while they are typically rarely detected (Bauer et al., 2006) or absent (Wilson and Moore, 1996) in northern temperate and colder climates. LaValley (2005) found varying detection of V. vulnificus occurrence in both freshly harvested and depurated oysters. Recent outbreaks in Alaska (McLaughlin et al., 2005) and the northeast and northwest coasts of the US (DePaola et al., 2000) have, however, heightened concerns of the spread of these organisms to colder waters. 1.2.3 Fecal-borne bacteria Fecal-borne bacteria are present in nearly all coastal waters where human activities and animals contribute fecal contamination. Fecal-borne bacterial species have been the cause of significant outbreaks of shellfish-borne diseases (Rippey, 1994) and are used as indicators for classifying shellfish harvesting waters almost exclusively worldwide. Fecal-borne bacteria are found in the gastrointestinal tracts of a wide range of organisms, including humans, all livestock and poultry, a wide range of rodents, waterfowl, other birds and other wild animals, and marine fish (Anderson et al., 1997; Silva and Hofer, 1993; Pourcher et al., 1991). Many of these species can also persist and even grow outside of the host in the natural environment, including coastal waters, sediments, and seaweed wrack (Jones et al., 2006; Brands et al., 2005; Anderson et al., 1997; Weiskel et al., 1996; Wilson and Moore, 1996; Gonzales et al., 1992). Other species, such as Aeromonas and Plesiomonas spp. can occur either in sewage or in estuarine environments (Rippey, 1994).
Microbial contamination and shellfish safety
7
There have been a variety of fecal-borne bacteria used as indicators of fecal contamination, with total and fecal coliforms, enterococci, E. coli, fecal streptococci, and Clostridium perfringens being the most commonly used. Some of these taxa may include virulent strains and species. For example, many strains of E. coli are non-virulent, but H7:O157 is a common pathogenic strain. Fecal coliforms include several pathogenic bacterial species, including Salmonella and Shigella spp., E. coli, Klebsiella spp. and Aeromonas hydrophila (APHA, 1995), which are all species that have been causes of shellfish-borne diseases in humans. 1.2.4 Protozoa Giardia lamblia and Cryptosporidium parvum are relatively common sources of waterborne disease outbreaks worldwide. Both have been detected in shellfish from a wide array of areas worldwide (Graczyk et al., 2007; Schets et al., 2007; Graczyk, 2003; Potasman et al., 2002; Stowell, 2001). Neither has been implicated in any major reported shellfish-related outbreak (Butt et al., 2004; Potasman et al., 2002).
1.3 Impacts of microbial contamination of shellfish on human health The relationship between sewage pollution and shellfish-borne disease incidence has been recognized for well over a century, though the actual microbial vectors were largely unknown until their successful identification during the last few decades. The incidence of shellfish-related infectious diseases is known to be associated with a few major types of pathogenic microorganisms, norovirus, Vibrio spp., Salmonella spp., Shigella spp., Hepatitis A virus (CDC, 2006; Cato, 1998; Rippey, 1994) and to a lesser extent other bacterial species and viruses. Although many studies have reported the presence of Cryptosporidium and Giardia in shellfish, there have been no reported outbreaks of cryptosporidiosis or giardiasis from human consumption of shellfish (Potasman et al., 2002). The availability of summarized data for shellfish-borne disease outbreaks is spotty and inconsistent, yet useful for general comparisons. Data covering outbreaks associated with shellfish or seafood in general from as far back as 1898 to as recent as 2002 have been summarized by CDC (2006), Cato (1998), and Rippey (1994). Large differences can be seen in the total number of the different bacterial causes in different areas (Table 1.1). Japan had a much great incidence of vibrio-borne outbreaks, especially compared with Canada and the EU, probably as a result of environmental conditions and types of food and cooking, but as well diagnosis and reporting. The US reported fewer outbreaks associated with fecal-borne bacteria than all other areas. Sumner and Ross (2002) calculated risk rankings for different hazard/product pairings as part of a seafood safety risk assessment for Australia. Viruses in oysters from
8
Shellfish safety and quality
Table 1.1 Most recent summaries of seafood or shellfish-borne disease outbreaks reported in different areas worldwide Causative agent V. parahaemolyticus V. cholerae Other vibrios Salmonella spp. Shigella/E. coli/ fecal coliforms Total bacteria Norovirus Hepatitis A Total viruses
Japan* Canada*
EU**
USy
USz
USx
Australia{
680 3 0 57
2 0 0 3
7 4 0 27
20 3 1 2
ND ND ND 3
14 16 10 ND
ND ND ND ND
18 815
17 26
0 ND
2 30
5 ND
ND ND
ND 6
15 1 16
7 42 53
ND ND 3
* Cato (1998) seafood-borne outbreaks: Japan: 1987±96; Canada: 1991±97. ** Cato (1998) fish and shellfish-number times reported: 1983±92. y CDC (2006) US shellfish-borne outbreaks: 1998±2002. z Rippey (1994) US shellfish-borne outbreaks: 1898±1990; sewage and/or wastewater. x Rippey (1994) US shellfish-borne outbreaks: 1967±90; associated with Vibrio genus. { Sumner and Ross (2002) seafood-borne outbreaks: 1990±2000.
uncontaminated waters had a relatively low risk ranking, V. vulnificus in oysters had an intermediate ranking, and viruses in oysters from contaminated waters had the highest ranking of shellfish-related hazard/product pairs. In the US, there was an average of 32 annual cases of reported V. vulnificus septicemia from eating raw Gulf of Mexico oysters from 1995 to 2001 (WHO/FAO, 2005). The relationship between the presence of pathogenic microbial species and disease is complex; species that are pathogenic often have strains that vary from virulent to avirulent. Using pathogenic vibrios as an example, the infective dose for humans is dependent in part on the environmental conditions at harvest, the presence and concentration of pathogenic strains, the physiological state of cells in the shellfish tissue (Smith and Oliver, 2006), host susceptibility (Gulig et al., 2005; Hlady and Klontz, 1996), the degree to which shellfish are cooked before being consumed, and a variety of other factors (FDA, 2005; WHO/FAO, 2005). Many shellfish-borne infectious diseases are caused by bacterial species that are for the most part non-pathogenic, including of Escherichia/Shigella, Salmonella, Campylobacter, and Vibrio (Hofreuter et al., 2006; Schaechter et al., 2001; Chatzidaki-Livanis et al., 2006). Most of the types of human viruses that have been detected in contaminated shellfish, i.e., the rotaviruses, astroviruses, enteroviruses, adenoviruses, and the calciviruses cause viral gastroenteritis, but infectious hepatitis from hepatitis A virus is also prevalent (Lees, 2000). The Norwalk-like viruses (NLV) are the most common, and significant recent progress has been made in detecting NLV in shellfish (Jothikumar et al., 2005; Loisy et al., 2005; Kageyama et al., 2003). Most fecal-borne bacteria also cause gastroenteritis as shellfish-borne disease. Improved wastewater treatment in many areas of the world has diminished their
Microbial contamination and shellfish safety
9
significance as disease agents in shellfish, but they remain problems (Cato, 1998) and pathogenic species are commonly detected in shellfish in many parts of the world (Brands et al., 2005; Wilson and Moore, 1996). 1.3.1 Pathogenic vibrios The pathogenic vibrio species of most concern are Vibrio parahaemolyticus (Yeung and Boor, 2004), Vibrio vulnificus (Gulig et al., 2005; Linkous and Oliver, 1999), and Vibrio cholerae (Colwell, 2004). All cause diarrhea and gastroenteritis, but V. vulnificus, and more rarely V. parahaemolyticus, may also cause septicemia through wound and intestinal infections with tissue damage occurring with rapid replication (Gulig et al., 2005). Death can occur within 24 hours and its case/fataility rate (~50%) is the highest of all foodborne pathogens (WHO/FAO, 2005; Hlady and Klontz, 1996). V. parahaemolyticus is one of the leading causes of foodborne illness in countries such as Japan and Taiwan (Pan et al., 1997; Wong et al., 1999) and is the most frequent cause of foodborne vibrio-associated gastroenteritis in the US (Daniels et al., 2000), with an estimated 2800 cases each year from the consumption of raw oysters (FDA, 2005). V. cholerae still infects millions annually throughout the world, though the incidence of shellfish-borne disease is relatively low (CDC, 2006; Rippey, 1994). Some climate change scenarios suggest significant trends of warming sea temperatures, in which case pathogenic vibrios may soon be found in coastal waters previously considered to be too cool. Not all environmental strains or vibrios have equal potential to cause disease, and pathogenic and non-pathogenic strains often coexist (Hurley et al., 2006; Deepanjali et al., 2005; Rosche et al., 2005; DePaola et al., 2003, 2000; Louis et al., 2003). Fortunately, pathogenic and non-pathogenic strains can be discriminated based on the presence or absence of known virulence genes (Nordstrom et al., 2007; Harwood et al., 2004; Panicker et al., 2004). There has been limited research on the heterogeneity of estuarine populations (Zimmerman et al., 2007; Lin et al., 2003; Jiang et al., 2000). Estuarine V. parahaemolyticus populations are thought to contain relatively low levels of pathogenic strains and these are a small fraction of total V. parahaemolyticus populations (FDA, 2005). Horizontal transfer and acquisition of genetic material, including virulence factors, between different bacterial species is likely to occur in shellfish waters and shellfish tissue where bacteria live in large populations as consortia. V. vulnificus, V. parahaemolyticus, and V. cholerae can co-exist within shellfish, and growth of V. cholera (Meibom et al., 2005) on chitin oligomers induces a state of natural competence allowing them to uptake foreign DNA. Other studies have shown evidence of horizontal gene acquisition between pathogenic vibrio species (Gonzalez-Escalona et al., 2006; Nishibuchi et al., 1996). There have been increases in the number of pathogenic vibrio outbreaks in the US in recent decades associated with the consumption of raw shellfish, including oysters and clams, and these outbreaks correlate with environmental conditions that enhance microbial populations (Balter et al., 2006; Underwood
10
Shellfish safety and quality
et al., 2006; McLaughlin et al., 2005; Collins, 2003; CDC, 1998). Vibrio disease outbreaks are generally associated with harvested shellfish from areas with warm water, like the Gulf of Mexico and the waters of Southeast Asia. Less is known about the incidence of virulent strains of V. vulnificus and V. parahaemolyticus in north temperate ecosystems. One general assumption is that strains from these areas are not virulent. Bauer et al. (2006) detected V. vulnificus, V. parahaemolyticus, and V. cholerae in Norwegian shellfish, but the isolates were negative for virulence factors. In the US, however, several recent outbreak of V. parahaemolyticus-induced gastroenteritis involving consumption of raw Alaskan oysters (McLaughlin et al., 2005), oysters and clams from Washington State and British Columbia (Balter et al., 2006), and several cases of V. parahaemolyticus infections in Maine can be traced to shellfish from state waters (Koufopolous, 2007; Maine CDC, 2005). These suggest virulent strains can be present in colder waters.
1.4 Effects of microbial contamination on the international shellfish industry The success of the international shellfish industry is significantly affected with microbial contamination. To protect public health, shellfish harvesting in most areas of the world is limited as a result of careful monitoring, especially under conditions where shellfish may be exposed to microbial contaminants. Assuming humans consume the shellfish raw or partially cooked, microbial contaminants taken up by shellfish and accumulated in tissue can remain viable and potentially virulent, and thus cause disease. The most effective strategy for limiting the harvest of contaminated shellfish is to harvest from areas with good water quality, i.e., microbial contamination is absent or at minimal levels. Comprehensive risk assessments that take into consideration a balance between shellfish harvesting to support the industry and protecting public health are essential tools to help guide how shellfish harvesting relative to microbial contamination is managed. 1.4.1 Major issues The many issues that are associated with microbial contamination for the international shellfish industry can be categorized into three main issues. The first is that microbial contamination limits shellfish harvesting. The concern by both the public health community and the industry to prevent disease outbreaks forces reduced harvesting efforts in areas that are contaminated. The changing landscape of scientific study results for effects of pathogens on human health, detection methods for specific pathogens, and emerging infectious diseases poses a challenge for the management community and industry to improve their abilities to prevent disease outbreaks and to adjust to new or re-emergent threats. Conflicting uses of coastal resources and other human activities is another issue. Other uses of coastal resources may conflict with shellfish harvesting if
Microbial contamination and shellfish safety
11
they result in increased microbial contamination. Expanding human development, especially in critical shoreline and wetland areas, can contribute to degradation of water quality directly through increased pollution, or via loss of ecosystem processes with habitat destruction that serve to mitigate upstream pollution. Climate change and the potential for sudden emergence of infectious diseases (Hunter, 2003) via insects, birds and other vectors are also human activity induced concerns that influence marine microbial communities and pathogen levels (Harvell et al., 1999). The popular press and other socioeconomic factors greatly influence the attitudes of potential shellfish consumers. Throughout the world, the relatively infrequent occurrence of shellfish-borne disease outbreaks often are accompanied by stories in the press that portray all shellfish consumption as dangerous without a balanced effort to put the cause of the outbreak into a more general context of safety for most shellfish consumption. At the crux of the overall problem of microbial contamination for the international shellfish industry is how to regulate harvesting, in relation to microbial contamination that successfully strikes a balance between the economic interests of the industry and the consumer, and the public health risk. Basic questions relate to whether microbial contamination is present, and whether contamination that is present is a threat to public health, either by its composition or its strength. Countries and their states and regions have adopted different standards for classifying shellfish harvesting waters, in most cases based on the likelihood of these areas becoming contaminated by human sewage. These standards, based on microbial indicators, have for the most part served both sides of the issue, though outbreaks have occurred in areas thought to be safe for consumption and there is a wide range of problems with current indicators. This topic will be covered in more detail in the next section. Beyond these issues, the variety of standards used in different areas of the world is in itself an issue for the international shellfish industry. Countries seeking to export shellfish are required to abide by these different standards, and choices are thus required which classification strategy should be adopted, thus limiting the number of countries that can import product. Different standards also confuse importation as the safety of imported product is more difficult to ascertain if harvesting standards are different from the importing countries food safety standards. There are efforts underway through the International Conference for Molluscan Shellfish Safety to determine how standards may be harmonized and to provide the means for sharing of new detection methods and related research between laboratories. 1.4.2 Existing and potential alternative standards for fecal-borne microbial contamination of shellfish waters Water quality (US) and shellfish tissue (EU) standards have long been established as a means of determining the level of sewage and microbial
12
Shellfish safety and quality
contamination likely to be in harvested shellfish. Simultaneous with the longterm use of these accepted indicators, many studies have identified a wide array of limitations that inform their application and have stimulated further studies on more effective, alternative indicators. There are many different fecal-borne human pathogens that have been detected and may be present in shellfish harvesting waters, as well as indigenous microorganisms that are also human pathogens. Direct monitoring of all potential pathogens is prohibitive for many reasons (Watkins and Burkhardt, 1996), including the cost for routine monitoring at multiple sites, many pathogens are difficult to detect, some are present rarely and at low concentrations, pathogens are not indexed to each other and infectious doses are not always known, avirulent strains exist, new pathogens are always emerging, and pathogen data are not predictive. Thus, an alternative strategy was obviously needed to address this issue. Most countries now have a classification system for shellfish harvesting areas that is based on the likelihood of these areas becoming contaminated by human sewage, largely because the main risk from shellfish consumption at the beginning of the twentieth century was sewage contamination. The concept of using fecal-borne indicator bacteria as surrogates for all fecal-borne pathogens logically emerged and has served as a relatively effective tool for managing public health risks associated with shellfish consumption. Fecal coliforms, a subset of the first indicators (total coliforms), are still widely used today and are made up of many different bacterial species with phenotypic properties that matched the defined assay conditions (APHA, 1995). This property was beneficial because it captured the most ubiquitous bacteria such as E. coli, and all of the targeted pathogens met the conditions. Thus, if Salmonella or Shigella spp. strains were present in higher concentrations than the more common E. coli, the assay would still be positive. Problems also existed as a result of the wide array of non-intestinal and naturally occurring species that met the assay conditions. Thus, measured fecal coliform levels could exceed standards even though no fecal pollution may be present. Other ideal indicator criteria that are violated by fecal coliforms, E. coli, enterococci, and other microbial indicators of fecal pollution include: growth in the environment; occasional lack of statistical relationship to pathogen concentrations; die off sooner than pathogens in unfavorable environmental conditions and under disinfection; and a tendency to become viable but non-culturable (Griffin et al., 2001; Watkins and Burkhardt, 1996). With noroviruses and vibrio bacteria being the main concerns with shellfish-borne microbial pathogens, alternative indicators are needed to address the full range of potential public health threats. Waters contaminated by sewage may also contain a great variety of human pathogenic viruses that may be taken up and accumulated by molluscan shellfish. The survival of traditional indicator bacteria in environmental waters, and their residence times and depuration kinetics in shellfish are significantly different from those for viruses. Consequently, standard bacterial indicator monitoring does not accurately reflect the presence of these human pathogenic
Microbial contamination and shellfish safety
13
viruses, especially in shellfish (Formiga-Cruz et al., 2003; Dore et al., 2003, 2000; Dore and Lees, 1995; Richards, 1988; Metcalf et al., 1979). In addition, direct enumeration of noroviruses, hepatitis A viruses, and other human enteroviruses in water and shellfish is again not practical. Thus, a reliable indicator that represents the viral quality of waters and shellfish is needed. The standards by which different countries classify shellfish harvesting areas are based on both water quality and the sanitary quality of shellfish tissue. Despite this basic difference, both systems use fecal-borne bacterial indicators as standards, fecal coliforms for the water quality-based system, and E. coli for the tissue-based standard. These are closely related because E. coli is a fecal coliform species. Lees (2000) summarized the different legislative standards for harvesting live shellfish in the US and the EU. Each classification system has two related classifications that require the same treatment of harvested shellfish: approved and restricted classifications in the US correspond to Categories A and B in the EU, requiring no treatment and depuration or relaying, respectively. The EU includes a Category C that requires protected relaying for >2 months. When microbiological standards in each system are exceeded, then harvesting is prohibited. Fecal coliforms and E. coli have been shown to be poor indicators of virus levels in shellfish (Formiga-Cruz et al., 2003; Dore et al., 2003, 2000; Dore and Lees, 1995; Richards, 1988; Metcalf et al., 1979). There are also reports on the presence of enteric viruses in shellfish otherwise meeting standards (Romalde et al., 2002), and sporadic outbreaks of shellfish-borne illnesses caused by viral pathogens continue to occur periodically around the world (Ng et al., 2006; Gallimore et al., 2005; Cheng et al., 2005). Thus, existing bacterial indicators of sewage are not reliable indices of risk from sewage-borne viral pathogens. As a result, considerable effort over the past two decades has been focused on identifying a more reliable assay for virus levels (Lees, 2000). Several viral indicators have been proposed (Muniain-Mujika et al., 2003), but several recent studies in Europe and the US have identified male specific coliphage as a potentially reliable viral indicator (Dore et al., 2003, 2000; Formiga-Cruz et al., 2003; Muniain-Mujika et al., 2003; Chung et al., 1998; Dore and Lees, 1995; Havelaar et al., 1993). Male-specific coliphage (MSC) are bacterial viruses (bacteriophage) that infect and replicate in E. coli cells that have F-pili. MSC are RNA or DNA viruses that do not replicate in E. coli below about 28 ëC (Dore et al., 1995; Woody and Cliver, 1995), so replication in most shellfish harvesting environments is limited. Relatively low levels of the MSC are found in fresh human and animal feces, and high levels occur in sewage (Calci et al., 1998). Enumeration methods for MSC are relatively inexpensive, easy to perform, and rapid, providing results within 24 hours. As such, these bacterial viruses are potentially important as indicators of sewage contamination monitoring the viral quality of shellfish harvesting waters and shellfish tissue. In the US, the Interstate Shellfish Sanitation Conference (ISSC) is in the process of evaluating MSC as a viral indicator (ISSC, 2007). A quantitative relationship between
14
Shellfish safety and quality
measurable levels of MSC and the absence of viral pathogens may exist (Dore et al., 2000), though further studies would be needed to verify this in a wider geographical context and under a range of environmental conditions. 1.4.3 Other limitations Neither fecal-borne bacteria nor viral indicators are useful in managing shellfish harvesting for control of diseases from naturally occurring vibrios (Marino et al., 2005; Koh et al., 1994; Shiaris et al., 1992), though Robertson and Tobin (1983) found V. parahaemolyticus concentrations increased with increases in fecal indicator bacteria concentrations in Nova Scotia. In the US, the ISSC has formed a Vibrio Management Committee with subcommittees to develop illness control measures for both V. vulnificus and V. parahaemolyticus. The main strategy recommended for control of V. vulnificus has been to require development and implementation of a Vibrio vulnificus Management Plan in any state with two or more V. vulnificus illnesses confirmed from shellfish originating from that state. The plan consists of increased educational efforts toward individuals with underlying health conditions that increase their risk for disease, processes to track each individual illness and products implicated in illnesses, limited harvest restriction, reduction in time from harvest to refrigeration, and post-harvest treatments. For V. parahaemolyticus, the main pre-harvest strategy to control illnesses is an interim control plan. The plan involves reacting to illnesses from an identified shellfish growing area, notify the shellfish industry and public health agencies of the potential for illness at historical times of onset/time of year, conduct an oyster meat sampling and assay program, recommend post-harvest temperature control, and, in more intensive cases, issue a consumption warning or ban on harvesting (ISSC, 2005; Daniels et al., 2000). Monitoring of shellfish for V. parahaemolyticus would be required in all areas where V. parahaemolyticus has occurred, the extent to which is dependent on the frequency of outbreaks in the previous 3 years. Growing areas would be closed for harvest if five or more tdh+ V. parahaemolyticus colony-forming units were detected per 0.1 g tissue in replicate samples. Reopening these areas would require confirmation through further testing of shellfish containing fewer than this threshold concentration in two consecutive samples. The Food and Drug Administration (FDA) has recently proposed the MSC assay for use in the National Shellfish Sanitation Program (NSSP) to monitor viral levels in quahogs and oysters as a means to re-opening harvest areas impacted by emergency closures due to wastewater treatment failure (ISSC, 2007). Shellfish that have been subjected to sewage contamination are assumed to contain viral pathogens and the NSSP presently requires a 21-day closure. Detection of MSC as indicators of pathogenic viruses in shellfish could be used as the basis for re-opening shellfish harvest areas in as few as 8 days after a sewage contamination event. Beyond harvest limitations resulting from emergency wastewater treatment failure events, significant areas of shellfish resources
Microbial contamination and shellfish safety
15
lie within the direct outfall zones of wastewater treatment facilities where shellfish harvesting is prohibited. Partially because of the lack of data concerning pathogen levels in shellfish the outfall zones, the FDA has adopted the relatively conservative 1 : 1000 dilution line for bed closure. Research is needed to determine how restrictive such closures should be, either to support this or a different strategy. Several other issues remain as significant impacts of microbial contamination on the international shellfish industry. The sources for most microbial contamination remain unknown as a result of a lack of programs to incorporate microbial source tracking methods into harvest area classification strategies. Under these conditions, general closures result even though the actual sources of elevated levels of indicator bacteria may have little or no public health significance. Outbreaks caused by shellfish from countries with inadequate control programs in place, or from illegal product from areas where harvesting is prohibited can be devastating influences on consumer attitudes to shellfish consumption. Public education efforts are key strategies for allaying the effects and to aid the press in accurate reporting on these incidences.
1.5
Incidence of microbial contamination in shellfish waters
The incidence of microbial contamination is a concern at critical steps in the process from when shellfish are harvested from coastal waters to the consumer, including post-harvest conditions that may exacerbate contamination risks. Besides post-harvest processes, which are covered in a different chapter, elevated temperatures during handling of shellfish immediately after their harvest can stimulate growth of pathogenic bacteria, especially vibrios, in shellfish tissue that increases their potential for infection. Temperature conditions are thus critical during harvest and initial processing, transportation and retailing. One management strategy currently used for controlling vibrios in US shellfish stock is to minimize temperature of freshly harvested oysters to reduce proliferation (ISSC, 2005). DePaola et al. (2000) state that any tdh and/or trhpositive strains of V. parahaemolyticus in oysters from the environment at concentrations >10/g tissue would be extraordinary. In addition, the concentrations of total V. parahaemolyticus rarely exceed 104/g. This may at first suggest a minimal public health threat. Vibrios are, however, known to grow rapidly to much higher concentrations, frequently >104/g, under inadequately managed post-harvest conditions, i.e., elevated temperatures from lack of refrigeration (Cook et al., 2002). The needle in the haystack can thus become a highly significant strain within the oyster microbial community. 1.5.1 Environmental effects on microbial contamination in shellfish waters Pathogenic vibrio species The abundance of vibrios in natural habitats, including bivalve shellfish (Pruzzo et al., 2005) and seaweeds (Mahmud et al., 2007) is intimately tied to abiotic
16
Shellfish safety and quality
factors (Martinez-Urtaza et al., 2008; Randa et al., 2004; Louis et al., 2003; Pfeffer et al., 2003; Lipp et al., 2002; Motes et al., 1998; Jones and SummerBrason, 1998). Temperature and nutrients strongly correlate with populations of all Vibrio spp; however, several studies indicate that distinct Vibrio populations are also associated with specific environmental conditions (Sousa et al., 2006; Eiler et al., 2006; Thompson et al., 2004). An important predictor of both general and specific populations is salinity. Vibrios tolerate a wide range of salinity (Farmer et al., 2005), with high levels detected at moderate salinities (5± 25 ppt) and an inverse correlation between salinity and abundance at elevated (>25 ppt) salinities (Martinez-Urtaza et al., 2008; Huq et al., 2005; Castaneda et al., 2005; Parvathi et al., 2004; Randa et al., 2004; Louis et al., 2003; DePaola et al., 2000; Motes et al., 1998; Chowdhury et al., 1992; O'Neill et al. 1992; Miller et al., 1982; Singelton et al. 1982; Kaper et al., 1979). Other factors such as dissolved organic carbon (DOC), suspended solids and plankton and copepods may also play a role (Gugliandolo et al., 2005; Jones and Summer Brason, 1998; Watkins and Cabelli, 1985). These correlations suggest direct effects of salt and other environmental factors on Vibrio growth and or survival in estuarine ecosystems. The potential for growth of pathogenic vibrios in harvested oysters is a highly significant factor relative to the public health risk for these bacteria (FDA, 2005; FAO/WHO, 2005). Surprisingly, despite these correlations between salinity and vibrio abundance or virulence, the direct mechanistic effect of salt on vibrios has been investigated to only a limited extent. For V. cholerae, high salinity induces conversion to the viable but non-culturable (VBNC) state (Thomas et al., 2006), whereas for V. vulnificus higher salt concentrations in sterile water reduced viability (Kaspar and Tamplin, 1993). Increases in total Vibrio populations seasonally or through time strongly correlate with increased temperature (Colwell, 2004; Thompson et al., 2004; DePaola et al., 2000), and nutrients (DOC) (Jones and Summer-Brason 1998). The climatic variables of higher temperature and increased flow/lower salinity are predicted conditions associated with global climate change (Louis et al., 2003). Acute rainfall can also result in the discharge of untreated waste into estuaries leading to increases in nitrogen and DOC. There have been many studies on the effects of environmental factors on pathogenic vibrio species throughout the world (Martinez-Urtaza et al., 2008; Lhafi and Kuhne, 2007; Zimmerman et al., 2007; Normanno et al., 2006; Eiler et al., 2006; Huq et al., 2005; Binsztein et al., 2004; Parvathi et al., 2004). All of these have served to provide useful information to shellfish growers, regulators, and researchers. Much less is known about the incidence of pathogenic strains of V. vulnificus and V. parahaemolyticus, especially in colder waters like those of the northeast US. Perhaps the key study on the occurrence of pathogenic V. parahaemolyticus strains in the US was by DePaola et al. (2003), in which 21.8% of samples of Alabama oysters had detectable pathogenic strains. Bauer et al. (2006) found a low frequency of trh+ but no tdh+ V. parahaemolyticus strains, while V. vulnificus was detected in only 0.1% of 885 blue mussels samples from the cold waters of Norway. DePaola et al. (2000) found tlh
Microbial contamination and shellfish safety
17
positive V. parahaemolyticus in oysters collected from Oyster Bay, NY, in October following the outbreak from that area in 1998, but no tdh+ strains. The water temperature in October was low enough to decrease total V. parahaemolyticus levels to 50% of fecal coliforms, Klebsiella pneumoniae, and Gram-negative bacteria were not adsorbed to particles > 5.0 m in diameter in natural stormwater in Michigan. This showed that most of the microorganisms of concern were not physically removed from the effluent, but the fact that a significant fraction of the microorganisms were attached suggests that bacterial attachment is a significant factor in the fate in stormwater runoff. Milne et al. (1986) studied the adsorption of fecal coliforms to both estuarine and sewage effluent suspended solids in laboratory experiments. They found almost immediate adsorption of fecal coliforms (20%) to estuarine suspended solids and this increased with time, indicating that fecal coliforms are deposited to estuarine sediment beds downstream of effluent discharges. Once the bacteria enter the water column, they may be more susceptible to environmental conditions. Solar radiation has been shown to be a major factor in the die-off of fecal coliforms and enterococci (Kay et al., 2005; Solic and Krstulovic, 1992). Shiaris et al. (1992) found tidal exposure to be a significant factor associated with disappearance of fecal coliforms, and probably enterococci, in sediments below a sewage outfall in Massachusetts, probably as a function of solar radiation. Pommepuy et al. (1992) showed that Salmonella sp. survived longer in turbid rather than clear marine waters because the suspended particles helped to protect bacterial cells from sunlight. Sorensen (1991) and Gonzalez et al. (1992) showed that predation by eucaryotic microorganisms was a very significant factor controlling bacteria survival in marine waters. De Vicente et al. (1988) reported greater die off of P. aeruginosa in laboratory seawater than in freshwater, but P. aeruginosa die off was slower than that of fecal and total coliforms in seawater. Studies by Pettibone et al. (1987) and Evison (1988) both found enterococci better able to survive in the environment compared with E. coli. Growth of fecalborne bacteria may also be possible in marine fish that live near sewage outfalls and polluted beaches (Silva and Hofer, 1993). Temperature by itself typically increases die off of bacteria and protozoa (Medema et al., 1997), but other factors can contribute to die off at higher temperatures. There is a balance in some areas as increased temperature also is accompanied by increases in re-growth of some organisms. Anderson et al. (1997) showed summer levels of enterococci in seaweed were elevated compared with winter levels. They concluded this reflects either re-growth or increased contamination by animals and insects. Weiskel et al. (1996) also found evidence of increases in fecal coliforms during summer in shoreline deposits of decaying vegetation/wrack in Buttermilk Bay, MA. There are many studies that have shown E. coli to be present in shellfish and in overlying waters, and they come from a variety of sources (McLellan, 2004). Other fecal-borne bacteria also occur in shellfish, including Salmonella and
Microbial contamination and shellfish safety
19
Campylobacter spp. (Brands et al., 2005; Teunis et al., 1997; Wilson and Moore, 1996). Fecal-borne bacteria typically depurate rapidly from shellfish tissue following contamination events (Marino et al., 2005; Croci et al., 2002; Jones et al., 1991), and may have short residence times in shellfish in harvest waters. The situation is different for viruses. Enteric viruses and MSC always remain in shellfish for longer periods of time and at higher levels than fecalborne bacteria (Lees, 2000). Enteric viruses and MSC have been detected in shellfish from many areas in the world (Umesha et al., 2008; Dore et al., 2003) and can persist for extensive periods in the marine environment (Gantzer et al., 1998; Callahan et al., 1995) and in shellfish (Myrmel et al., 2004; Formiga-Cruz et al., 2002; Lees, 2000). An interesting and consistently observed phenomenon is the higher levels of MSC and enteric virus incidence during winter that accompanies an increase in disease incidence compared with other months. Several research groups have reported this in many areas of the world, including Norway (Myrmel et al., 2004), England and Wales (Dore et al., 2003), France, and the US (Burkhardt and Calci, 2000). Burkhardt and Calci (2000) suggested this may be associated with selective accumulation of viruses by oysters, and potentially other shellfish, during the winter. Catastrophic natural events Some disasters that result from catastrophic natural events may be unusual yet significant sources of microbial contamination to coastal waters, including shellfish harvesting areas. Hurricane Katrina that impacted the US Gulf of Mexico in August 2005 not only resulted in the closure of shellfish harvesting areas well into September 2005, but also caused loss of product that spoiled owing to lack of refrigeration or exposure to flood waters following the hurricane (FDA/CFSAN, 2005). Elevated levels of pathogenic vibrio species were detected in post-hurricane floodwaters, (Demcheck et al., 2005). Five deaths involving V. parahaemolyticus and V. vulnificus as wound infections occurred as a result of exposure to contaminated flood waters (CDC, 2005). In a similar fashion, after 42 consecutive days of rain during March 2006 in the Honolulu area of Hawaii, a sewage main ruptured and untreated wastewater spilled into a canal (UH-WRRC, 2006). The canal contained ~8 105 enterococci/100 ml, and there were three related wound infections reported, two not serious, with one of those related to fecal-borne bacteria. The other involved a V. vulnificus infection where a man died 2 days after falling into a polluted canal. In many other areas of the world, disasters may displace populations who then lack sources of clean water and sanitation, creating conditions where cholera and other diseases may thrive. 1.5.2 Microbial physiology and survival capacity There seems to be an obvious ecological purpose for the many bacteria that enter a VBNC physiological state. A variety of conditions can induce this state in
20
Shellfish safety and quality
bacteria, including low temperatures and carbon/energy source concentrations, both of which are key ecological conditions that could affect survival in marine ecosystems. This condition has more significant public health implications because of the potential for pathogens to be present yet not be detected using traditional culture-based methods, and may affect molecular genetic methods (Vora et al., 2005). The pathogenic vibrio species of concern to shellfish consumers are enteropathogenic and also retain pathogenicity in the VBNC state (Vora et al., 2005; Binsztein et al., 2004; Wong et al., 2004; Nishino et al., 2003; Mizunoe et al., 2000; Whitesides and Oliver, 1997). Many fecal-borne bacteria also exhibit the VBNC state. Pommepuy et al. (1996) reported VBNC E. coli retained enteropathogenicity even when exposed to seawater and sunlight. Enterococcus faecalis can survive for long periods of time without any nutrients (Hartke et al., 1998) and can be detected as VNBC cells using competitive polymerase chain reaction (PCR) (Lleo et al., 1998). The traditional concept for fecal indicator bacteria is that they are adapted for existence in gastrointestinal tracts and cannot survive for long in the environment, especially in seawater. Not only can they survive, but some have been shown to grow under favorable conditions (Ishii et al., 2007; Kon et al., 2007; Jones et al., 2006; Hartel et al., 2005; Solo-Gabriele et al., 2000). In addition, wastewater disinfection is intended to be lethal for microorganisms, and its effectiveness is essential for preventing shellfish exposure to potentially abundant bacterial and viral pathogens in wastewater. In many comparative studies, E. coli, Ent. faecalis and other fecal-borne bacteria have been shown to be inactivated by chlorine to a greater degree and more rapidly than enteroviruses and MSC (Tree et al., 2003, 1997; Tyrrell et al., 1995). Bolster et al. (2005), however, found E. coli could not only survive chlorination, but could also grow in some estuarine water, though they used a recovery method for E. coli detection that is not commonly used. Similar findings were reported by Blatchley et al. (2007) and Knorr and Torella (1995) showed evidence of Salmonella multiplication in a wastewater depuration pond in Spain. UV disinfection (Blatchley et al., 2007) and ozone (Xu et al., 2002) have been effective at inactivating viruses in wastewater, with some preconditions to ensure effectiveness.
1.6
Contamination sources and their identification
There are several ways to define what may be considered fecal-borne sources of microbial contaminants in shellfish waters. One way is to consider marinas, dysfunctional wastewater treatment facilities, septic systems, animal feedlots, and various types of runoff as sources. Another way is to consider actual source species. The latter fall into categories that include human versus non-human, or at a more detailed level, human, pet, livestock, wild birds, and other wild animals. Still another approach would be to consider sources as being related to transport mechanism, i.e., direct deposition or transported via groundwater,
Microbial contamination and shellfish safety
21
stormwater or surface water (Weiskel et al., 1996). All approaches for defining sources are useful in directing different management actions to eliminate contamination of shellfish harvesting waters. The following section discusses different sources of microbial contamination and strategies used to identify sources. Pollution source identification is a critical step in the management of shellfish harvesting waters. Strategies that result in accurate identification of the most significant source(s) of pollution are invaluable for focusing the most effective allocation of what are often scarce resources to improve water quality. Though naturally occurring pathogens such as Vibrio spp. are significant public health concerns for the shellfish industry, their source is for the most part the natural environment; their incidence relative to fecal contamination was discussed in the previous discussion. It is the potentially manageable fecal-borne contamination, from sewage and other sources, that is the focus of this section. The categorization of sources of microbial contamination as they relate to actual source species is inclusive of other ways of categorizing sources. At the simplest level of discrimination, sources can be considered to be of either human or non-human origin. Examples of what constitutes human and non-human sources in shellfish areas in the Northeast US are presented in Table 1.2. The delineations could be rearranged depending on perspective, but this scheme suggests different strategies for managing different types of contamination Table 1.2 Different sources of fecal contamination in shellfish areas of the northeast US Human sources Point sources Inadequate wastewater treatment facility sewage treatment Illegal sanitary connections to storm drains Illegal disposal to storm drains Combined sewer overflows Sanitary sewer overflows Non-point sources Failing septic systems Landfills Marinas, pump-out facilities & boaters Direct deposition Non-human sources Domestic animals and urban wildlife Dogs, cats Foxes, raccoons, skunks Livestock and rural wildlife Cattle, horses, poultry, other livestock Otters, muskrats, deer, bear Waterfowl
22
Shellfish safety and quality
sources. Both types can be significant sources of contamination. Accurate identification of which type of source may be significant is obviously critical for assessing public health risks and resource allocation for eliminating sources. Human pathogenic viruses are from human sources. Thus, the prevailing perception is that there is a greater public health risk associated with human sources compared with non-human sources (Sinton et al., 1998; Wiggins, 1996; O'Shea and Field, 1992). The public health significance of non-human sources of fecal pollution may be less than for human-borne sources, but pathogens may also occur in non-human sources (Arnone and Walling, 2007) and can be a concern when exposure to high concentrations occurs. The US Environment Protection Agency (EPA) has recently acknowledged the potential public health risks associated with non-human sources of pollution (US EPA, 2007). It is accepted that pathogens in humans are only present in significant numbers in diseased individuals, while some animal hosts may act as more stable reservoirs of pathogens. There are many reports of human disease occurrence associated with livestock animals, but few reported cases associated with wild animals (Field and Samadpour, 2007; Craun et al., 2004). This difference between domestic and wild animals is probably a function of both the frequency of exposure and the number of studies conducted on the different animal groups. Diseases associated with exposure to pet feces are also well documented, so domestic animal sources are of concern at a level that falls between that for human sources and wild animal sources. There are a variety of transport mechanisms by which these types of source may contaminate shellfish waters, but little information on their significance is available at present. For shellfish growing waters, non-human sources of microbial contamination have always been a management dilemma because of the competing issues of public health protection and opening areas for harvest. 1.6.1 Microbial source tracking Recent adoption of biotechnological techniques for application to water quality issues has spawned a number of approaches to address identification of sources of fecal-borne contamination. Many methods use non-microbiological approaches to identify sewage contamination, while other approaches that utilize microorganisms are often called `microbial source tracking' (MST) methods. MST methods have been used successfully for nearly 15 years throughout the US. The various methods range from those that determine phenotypic aspects of bacteria to cultivation and `library'-dependent methods to direct detection of sourcespecies specific genetic markers that allow for source tracking. There has been a series of reviews that have compared the advantages and disadvantages of a comprehensive list of MST methods. The US EPA published the `Microbial Source Tracking Guideline Document' in 2005 (US EPA, 2005) as a comprehensive review of all aspects of MST up to that time period. The general consensus conveyed in that and other independent reviews has been that most MST methods have merit, no one method has been deemed superior to the others
Microbial contamination and shellfish safety
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for all conditions, all methods are useful and amenable to improved application with modification and optimization of the methods, and that use of a combination of methods may be useful in complex MST studies or in large watersheds (Fields and Samadpour, 2007; Meays et al., 2004; Stoeckel et al., 2004; Stewart et al., 2003; Myoda et al., 2003; Simpson et al., 2002; Scott et al., 2002). Progress is being made in the development of new methods that may overcome observed limitations with existing MST methods for identifying source species in all environments. For example, Walters et al. (2007) reported the detection of E. coli O157:H7 to be highly related to detection of ruminantspecific markers in Bacteroidales markers, and not with those from humans and swine in river water from Alberta, CA, suggesting the source of this pathogen is probably cattle feces. They also reported an increased likelihood of the presence of Salmonella spp. when ruminant-specific Bacteroidales markers were detected. This study is significant for public health risk assessments because it demonstrates a relationship between the indicator and actual human pathogens. Efforts are underway to improve identification of bird and other wild animal sources of fecal contamination, including ducks (Devane et al., 2007), elk and gulls (Dick et al., 2005), and raccoons (Ram et al., 2007). Several published MST studies, using a range of methods, have been conducted in shellfish and other coastal waters. Parveen et al. (1999) used ribotyping to differentiate E. coli isolates collected from human and non-human sources of contamination in Apalachicola Bay, Florida. The isolates from nonhuman sources were from drainage from protected marshlands, and the human source isolates were from wastewater treatment facility effluent. Simmons et al. (1995) reported on the use of pulse field gel electrophoresis (PFGE) to identify fecal coliform sources in Virginia. In an earlier study, they found that raccoons were the main source species of pollution that limited harvest of shellfish (clams) in the Cherrystone inlet of Virginia. Fecal coliform levels declined after raccoons were trapped and moved from the area. Some recent studies have employed human-specific and ruminant-specific primers with Bacteroidales PCR to identify areas of elevated fecal contamination and sources in Tillamock Bay, Oregon (Shanks et al., 2006; Bernhard and Fields, 2000). Several recent studies in the US have used E. coli ribotyping as an MST approach for identifying source species of fecal pollution in coastal waters. Clayton (2006) reported birds as the most prevalent source of bacterial pollution in shellfish (manila clams, oysters) harvesting waters on Vancouver Island, British Columbia, Canada, based on E. coli ribotyping analysis. Jones (2008a) reported the results of three small studies conducted in Massachusetts (Martha's Vineyard) and coastal New Hampshire. Each area had different types of source species that were the most significant sources of fecal pollution. In the salt ponds of Martha's Vineyard, birds, especially Canada geese, were the most prevalent source, constituting 86% of identified sources, and confirming local suspicions. In the New Hampshire Seacoast, wild animals (66% of identified sources) were the most significant type of source in Berry Brook in Rye, NH, while humans (41%) and pets (29%) were by far the most prevalent source types in Garrison
24
Shellfish safety and quality
Table 1.3 Identified types of sources for all sites on three sampling dates in Stockton Springs Harbor, Maine Dry weather (19 August 2007)
Wet weather (16 November 2007)
Wet weather (18 January 2008)
Source type
Isolates
Total (%)
Isolates
Total (%)
Isolates
Total (%)
Human Pet Bird Wild animal Livestock
0 0 2 2 0
0 0 15 15 0
4 0 13 5 1
10 0 31 12 2
14 2 10 9 0
25 4 18 16 0
4 9 13
31 69 100
23 19 42
55 45 100
35 22 57
61 39 100
Identified Unknown Total
Brook in Dover, NH. In these areas and in South Carolina (Kelsey et al., 2008), local, temporally specific source species databases were most appropriate for source species identification in small watersheds. In another recent study conducted in Stockton Springs, Maine (Jones, 2008b), fecal coliform/E. coli concentrations in the harbor and tributaries were higher and more widespread during wet than during dry weather, and during winter compared with autumn. In all, 55% of the isolated E. coli strains were identified to source species. Birds, particularly herring gulls, were the most prevalent (22% of all isolates) type of source species, while human (16%) and wild animal (fox, raccoon; 14%) sources were also significant. The pattern of which types of source were most significant changed seasonally, with the fraction identified as coming from human sources increased from 0% of the isolates during summer to 25% of the isolates in winter (Table 1.3). The fraction of isolates identified from wild animals remained relatively constant (12±16%) and the fraction from birds (15±31%) remained high over the 7-month study period. These findings emphasize the changing nature of environmental, ecological and human activity conditions that that determine the types of source that may pollute shellfish waters through harvest seasons. The common focus on bacteria may be a limitation for assessing sources of viral contamination. A variety of MST approaches have been developed that address viruses as indicators or that involve direct detection of specific viruses, including human-specific adenoviruses and enteroviruses (US EPA, 2005). Other methods have focused on bovine enteroviruses, porcine adenoviruses and Teschoviruses to identify animal sources. Great progress has been made in developing procedures for detecting the four subgroups of F+-specific RNA coliphages for MST. Kirs and Smith (2007) developed a Q-PCR MST assay for these viral indicators, while Love and Sobsey (2007) developed a group-specific antibody-based particle agglutination technique that is rapid and simple for detecting F+ coliphage in shellfish waters.
Microbial contamination and shellfish safety
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1.6.2 Integrated studies There have been studies published that address fecal contamination sources using an integrated approach and a variety of tools. The relationship between urbanization and pollution of shellfish waters has been addressed by Lipp et al. (2001), Mallin et al. (2001), Weiskel et al. (1996), and Maiolo and Tschetter (1981). More recently, Kelsey et al. (2003) used both geographical information system (GIS) and MST, specifically analysis of multiple antibiotic resistance in E. coli, to determine sources of fecal pollution in South Carolina. The results suggested that non-human sources may constitute a majority of the fecal pollution in the study area, with a few human sources near boat moorings and sewage collection lift stations. Hartel et al. (2007) included use of a humanspecific marker (esp) in Enterococcus faecium (Scott et al., 2005) to complement an in-depth assessment of the use of fluorometric detection of optical brighteners as an indicator of fecal contamination. Inclusion of organic matter analysis and an additional filter for fluorometry were required for accurate source identification in coastal waters. Several other studies have performed simulation modeling to predict the influence of coastal lagoons on water quality in shellfish areas. The Thau lagoon in southern France is an important shellfish growing area. A simulation approach was developed by Fiandrino et al. (2003) that involved bacterial transport and survival to predict the timing, distribution, and relative loading of bacterial contamination from the two main rivers that flow into the lagoon. Sanders et al. (2005) considered bacterial transport and survival to predict the cycling of bacteria from runoff, shore birds and re-suspended sediments during dry weather in a California coastal lagoon. Both modeling approaches successfully predicted concentrations of targeted fecal indicator bacteria, and serve as useful management tools for those areas. Pommepuy et al. (2004) used a hydrodynamic model that also took into account microbial behavior to predict E. coli, F-RNA phage and norovirus levels in shellfish at several sites within the Golfe du Morbihan in France. 1.6.3 Other sources Adaptation of wild animals to urban settings, global animal trade (Marano et al., 2007), encroachment of human settlement into wilderness areas, and a variety of related factors (Harrus and Baneth, 2005) along with improved detection and identification of well-known and emerging pathogens may result in more frequent wild animal-borne infections in the future, and thus a greater significance given to them as public health threats. A variety of human pathogens have been isolated from wild birds and other animals (Waldenstrom et al. 2003; Moore et al. 2002; Ching et al., 2000), including Campylobacter spp., Cryptosporidium spp., Heliobacter canadensis, E. coli, Salmonella spp., and a variety of parasites. Wild bird species (e.g., crows and gulls) are attracted to sewage treatment lagoons, garbage dumps, manure-applied agriculture fields, and other potential sources of fecal-borne pathogens. Little information at present is
26
Shellfish safety and quality
available on the significance of wild birds as vectors in the transmission of human diseases, though fecal-borne pathogenic bacteria have been isolated from gulls (Dixon, 2007). Recent studies suggest that herring gulls (Larus argentatus), and to a much lesser extent great black-backed gulls (L. marinus), may carry E. coli from human sources to remote islands off the coast of Maine and New Hampshire (Nelson et al., 2008). DePaola et al. (1994) confirmed earlier reports of the incidence of V. vulnificus in the gastrointestinal tract of finfish in the Gulf of Mexico at levels for some of the 17 species greater than in oysters, sediment and water. They found levels to be high in the summer and much lower in most finfish during winter. Certain finfish species, especially sheepshead, may serve as reservoirs for V. vulnificus, which would be especially significant during winter, when environmental conditions are less favorable for growth. Ballast water is another unique yet potentially significant source of pathogens in shellfish waters. Several studies have surveyed ships entering US coastal waters from other countries and have reported detection of bacterial pathogens. Ruiz et al. (2000) detected V. cholerae 01 and 0139 in all tested commercial ships entering Chesapeake Bay. In contrast, Burkholder et al. (2007) detected no V. cholerae in Department of Defense vessels, yet found other bacterial pathogens. It is unknown if migratory fish carry pathogenic vibrios to areas where they are less prevalent. Drake et al. (2007) estimate 1020 bacteria and viruses, including pathogenic forms, are delivered to Chesapeake Bay each year in ballast water. 1.6.4 Related issues It is clear that both human and non-human sources of microbial contaminants can be significant and contain potentially harmful pathogens. A complicating factor for managing shellfish waters is the potential presence of indicator bacteria that persist and sometimes grow in favorable environments. Coastal sediments and beach sands can be significant reservoirs of fecal-borne bacteria (Hartel et al., 2005; Solo-Gabriele et al., 2000). Recent studies have shown beach sand and sediments can act as sources and sinks for E. coli from human and waterfowl sources (Ishii et al., 2007) as well as being a favorable environment for the establishment of distinct E. coli populations (Kon et al., 2007). In most cases, these bacteria do not include pathogenic strains, but seaweed wrack and macrophytic algae can harbor both fecal-borne indicator bacteria and pathogens (Ishii et al., 2006; Grant et al., 2001; Weiskel et al., 1996). Jones et al. (2006) reported high levels (1±3 105 most probable number (MPN)/g dry weight (DW)) of enterococci were harbored in seaweed wrack and underlying sand (1±6 103 MPN/g DW), whereas beach sand had low (30 ëC) for enterococci growth during warm summer months, and can be a significant contamination source (6% annual input; Weiskel et al.,
Microbial contamination and shellfish safety
27
1996) in coastal areas in the northeast US. Growth of fecal-borne bacteria in coastal soils in tropical areas has also been reported (Byappanahalli and Fujioka, 1998). As previously mentioned, the varying effects of different disinfection strategies on microorganisms during sewage treatment are a concern to shellfish resource and water quality managers. Bolster et al. (2005) showed that pure cultures of chlorine-exposed E. coli not only recovered from the disinfection process, but also grew in low to moderately saline (0 to 10.5 ppt) estuarine water microcosms. The most extensive growth was observed in water containing elevated levels of nutrients, i.e., ammonium, dissolved organic carbon, at sites in close proximity to wastewater treatment facility effluent discharges. Thus, discharge of what appears to be effluent with few culturable indicator bacteria cells may actually contain indicators and pathogens that may recover and even grow in shellfish-harvesting waters under favorable conditions. Further work is needed to enable more accurate assessment of the public health significance of this wide variety of different sources of fecal-borne indicator bacteria in shellfish-harvesting waters.
1.7
Future trends
There are many critical issues that warrant attention by the research community to gain a better understanding of the critical issues in the near future. Perhaps the most critical issue is the largely unknown impacts of climate change and global warming on infectious diseases, both existing, emerging and new. Shellfish management strategies and indicators used to open and close shellfish harvesting will need to take into account the potentially changing landscape of what diseases pose the greatest risk. This also implies a continuous effort to adapt to changing problems as well as improvements in technologies and methods. New indicators and direct detection of pathogens should emerge on a relatively frequent basis given the capabilities of modern-day molecular-based methodologies. Perhaps the easiest, yet most challenging issue is how to communicate better around the planet Earth. New advances are being discovered on a daily basis in the different research and government laboratories around the world in all aspects of microbial contamination of shellfish and growing waters. It remains difficult to stay current with new developments and to identify an effective means of sharing this information. The International Conference on Molluscan Shellfish Safety has been discussing how to accomplish this goal, and has made some recent efforts to put the solutions in place. For many areas, microbial contamination has exceeded standards, and the nature of the most significant sources is such that their elimination would be quite difficult. Despite this condition, shellfish can be harvested from areas where standards are only slightly exceeded if they are processed to remove pathogens following harvest and prior to being marketed. This has traditionally
28
Shellfish safety and quality
involved depuration and relaying, but these strategies are not always effective for all the different types of microbial contaminants. New post-harvest treatment technologies are under development and can be adopted as further strategies to protect the shellfish-consuming public. Along with considerations for postharvest processing is the need for a more integrated approach to assessing the actual risk posed by detected microbial contamination. The potential for regrowth in natural environments and the significance of non-human sources of fecal contamination suggests that modeling risk is more complicated than in previous conceptual frameworks. More accurate modeling that integrate environmental, meteorological, hydrodynamic and microbial factors would be useful for predicting, or preventing possible outbreaks.
1.8
Sources of further information and advice
US · Center for Disease Control and Prevention: http://www.cdc.gov/ · Interstate Shellfish Sanitation Conference (ISSC): http://www.issc.org/ · Dr James Oliver's Vibrio vulnificus home page: http://www.vibrio.com/ · US FDA Foodborne Pathogenic Microorganisms and Natural Toxins Handbook The Bad Bug Book: http://vm.cfsan.fda.gov/~MOW/intro.html EU · Centre for Environment, Fisheries & Aquaculture Science, UK: http:// www.cefas.co.uk/ · French Research Institute for Exploitation of the Sea: http://www.ifremer.fr · AEPM-EMPA, European Mollusc Producers Association: www.euraquaculture.info/index.php?option=com_content&task=view&id= 26&Itemid=43 NZ · New Zealand Food Safety Authority: http://www.nzfsa.govt.nz/
1.9
References and further reading
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BALTER S, HANSON H, KORNSTEIN L, LEE L, REDDY V, SAHL S, STAVINSKY F, FAGE M, JOHNSON G, BANCROFT J, KEENE W, KOEPSELL J, WILLIAMS M, MACDONALD K, NAPOLILLI N, HOFMANN J, BOPP C, LYNCH M, MOORE K, PAINTER J, PUHR N, YU P
(2006), `Vibrio
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parahaemolyticus infections associated with consumption of raw shellfish ± three states, 2006', Morb Mortal Wkly Rep, 55 (31), 854±856. BARBE D E, FRANCIS J C (1995), `An analysis of seasonal fecal coliform levels in the Tchefuncte River', Wat Res Bull, 31, 141±146. BAROSS J A, LISTON J, MORITA R Y (1978), `Incidence of Vibrio parahaemolyticus bacteriophages and other Vibrio bacteriophages in marine samples', Appl Environ Microbiol, 36, 492±499. BAUER A, OSTENSVIK O, FLORVAG M, RORVIK LM (2006), `Vibrio parahaemolyticus, V. cholerae and V. vulnificus in Norwegian shellfish', in Molluscan Shellfish Safety. Proceedings of the 5th International Conference on Molluscan Shellfish Safety, 295±300. Henshilwood K, Deegan B, McMahon T, Cusack C, Keaveney S, Sike J, O'Cinneide M, Lyons D, Hess P (eds). Galway, Ireland, 4±18 June 2004. The Marine Institute, Rinville, Oranmore, Galway, Ireland. BERNHARD A E, FIELD K G (2000), `A PCR assay to discriminate human and ruminant feces on the basis of host differences in Bacteroides-prevotella genes encoding 16S rRNA', Appl Environ Microbiol, 66, 4571±4574. BINSZTEIN N, COSTAGLIOLA M C, PICHEL M, JURQUIZA V, RAMIÂREZ F, AKSELMAN R, VACCHINO
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DJ
2 Biotoxin contamination and shellfish safety H. HeÂgaret, University of Connecticut, USA, G. H. Wikfors, NOAA Northeast Fisheries Science Center, USA and S. E. Shumway, University of Connecticut, USA
Abstract: Some phytoplankton species can produce biotoxins. As molluscan shellfish filter-feed on these toxic species, they bioaccumulate the phycotoxins, which can be transferred to higher trophic levels, resulting in several types of poisoning syndromes. The main shellfish-mediated, phycotoxin-associated, human poisoning syndromes are paralytic, amnesic, neurotoxic, diarrheic and azaspirazid shellfish poisonings; a few other phytoplankton species can also cause harmful effects in humans. This chapter lists the major phycotoxins described, their origins, occurrences, and impact on human health as well as management procedures developed to address them. The trophic dynamics of the phycotoxins are also discussed, as is their bioaccumulation, biotransformation, and detoxification processes. Key words: shellfish, toxins, harmful algal blooms.
2.1
Introduction
The potential for molluscan shellfish to contain toxins affecting human consumers has been recognized for all of human history. Native American taboos against eating shellfish from waters affected by toxic phytoplankton blooms presaged the often-recounted story of Captain Vancouver's crew's 1793 fatal experience with biotoxin-contaminated shellfish (Dale and Yentsch, 1978). Seasonality in toxicity led to recognition that specific microalgae eaten by molluscs are the origins of many such toxins (Shumway, 1990). Further, monitoring and surveillance programs implemented to protect human consumers of shellfish have led to the recognition of additional toxins not previously known
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(e.g., azospiracid) and provided convincing evidence that microalgal blooms leading to biotoxin-contaminated seafood are increasing in geographic distribution and severity (Hallegraeff, 2004). The presence of toxicity in shellfish is the ultimate result of a complex cascade of environmental, biological, physiological, and biochemical interactions. Protection of human consumers of shellfish has focused research and management on all aspects of this cascade of interactions. Research has presented the management community with options for consumer protection, and these options often are context-specific; i.e., specific for a given shellfish species or a specific biotoxin. Thus, a review of biotoxin contamination and shellfish safety requires description of general principles and then application of these principles to specific instances. There are approximately 4000 described species of marine microalgae, or phytoplankton, but only a few (70±80 species, about 2%) produce toxic compounds (Scoging, 1998). These harmful algae are natural and have been part of the ecosystem for millennia. As human activities have increased, incidence and severity of harmful algal blooms (HABs) have also increased and so have risks for human exposure (Hallegraeff, 2004). Generically, a phytoplankton `bloom' refers to the development of a microalgal population in a given water mass (Smayda, 1997). The maximal population density of a toxic bloom can vary from 200 cells/ml to more than 50 000 cells/ ml, depending on the species of toxic algae and the conditions of the bloom. Certain species of toxic phytoplankton contain red or brown pigmentation; if the concentration of the bloom is high enough, a bloom can become visible and thus described as a `red tide' or a `brown tide' (Shumway, 1990). Proliferation of harmful algae does not necessarily impart color to the water and, conversely, harmless algae can also discolor water masses. Chemical compounds produced by phytoplankton, collectively termed phycotoxins, potentially are responsible for poisoning incidents worldwide. These incidents can trigger disease and death of marine life (Landsberg, 2002) and are also a threat for human intoxication or death if contaminated seafoods are ingested or aerosols are inhaled. The types of intoxication or disease associated with specific harmful algae are diverse, and several seafoodpoisoning syndromes have been described, depending upon the type of phycotoxin to which humans are exposed (Table 2.1). Although these toxic syndromes can result from the consumption of finfish or shellfish, this review will focus only on the phycotoxins present in shellfish. The main shellfish-mediated, phycotoxin-associated, seafood poisoning syndromes are: paralytic shellfish poisoning (PSP), amnesic shellfish poisoning (ASP), neurotoxic shellfish poisoning (NSP), diarrheic shellfish poisoning (DSP) and azaspirazid shellfish poisoning (AZP). A few other phytoplankton species can cause harmful effects in humans, but are not included in the aforementioned, symptom-based syndromes, e.g. phytoplankton species including Pfiesteria spp. and Alexandrium monilatum. Bivalves are filter feeders that, as they consume phytoplankton, can
Table 2.1
Origins of phycotoxins
Disease
Causative organisms
Major toxins
Bioactive mechanism
Acute symptoms
Chronic symptoms
Diagnostic
Paralytic shellfish poisoning (PSP)
Pelagic dinoflagellates Alexandrium spp. Gymnodinium catenatum Pyrodinium bahamense
Saxitoxins Gonyautoxins C-Toxins
Na+ Channel Blocker: neurotoxins bind to site 1 on the voltage-dependent sodium channel, blocking the influx of sodium into excitable cells and restricting signal transmission between neurons
Tingling and numbness in mouth, ataxia, dizziness, headache, respiratory distress and muscular paralysis Nausea, vomiting, diarrhea, paraesthesia, respiratory depression
None known
Clinical, mouse bioassay of food, HPLC
Neurotoxic shellfish poisoning (NSP)
Pelagic Brevetoxins dinoflagellates and and raphidophytes brevetoxinKarenia spp. like compounds
Na+ channel blocker: binds to site 5 on the sodium voltage-sensitive channels and alters properties of excitable cells by shifting activation to more negative potentials, triggering membrane depolarization
Nausea, vomiting, diarrhea, bronchoconstriction, reversal of temperature sensations, paraesthesia
None known
Clinical, mouse bioassay, HPLC, ELISA
Amnesic shellfish poisoning (ASP)
Pelagic diatoms Pseudo-nitzschia spp.
Glutamate receptor agonist: binds to kainite-type glutamate receptors in the brain, functioning as an excitatory neurotransmitter causing depolarisation of neurons, followed by calcium ion influx, neuronal swelling, and cell death
Nausea, vomiting, diarrhea, amnesia, paraesthesia, respiratory depression
Amnesia
Clinical, mouse bioassay of food, HPLC
Domoic acid
Table 2.1
Origins of phycotoxins
Disease
Causative organisms
Major toxins
Bioactive mechanism
Acute symptoms
Chronic symptoms
Diagnostic
Diarrheic shellfish poisoning (DSP)
Pelagic or benthic dinoflagellates Dinophysis spp., Prorocentrum lima
Okadaic acid Dinophysistoxins Pectenotoxins and yessotoxins
Protein phosphatase type 1 and 2A inhibitor: increases protein phosphorylation Unknown mechanisms for PTX and YTX
Gastrointestinal distress, nausea, vomiting, diarrhea
None known
Clinical, mouse bioassay, HPLC, ELISA
Azaspirazid shellfish poisoning (AZP)
Pelagic dinoflagellate Protoperidinium sp.
Azaspiracid
Unknown mechanism
Nausea, vomiting, severe diarrhea (i.e., similar symptoms to DSP) and stomach cramps
Additional taxa implicated in human poisoning
Pfiesteria spp.
Radical forming toxic organicligated metal complex Uncharacterized Uncharacterized Goniodomin A
Newly documented carbonsulfur-metal-based radical production
Prorocentrum minimum Prorocentrum micans Alexandrium monilatum
Human poisoning not confirmed Human poisoning not confirmed Can change its conformation which might alter the actomyosin ATPase activity
HPLC, high-performance liquid chromatography. ELISA, enzyme-linked immunosorbent assay. Details compiled from Shumway (1990), Landsberg (2002), and Hallegraeff (2004).
Biotoxin contamination and shellfish safety
47
accumulate toxins within tissues, along with nutritional compounds associated with the phytoplankton. As little as 6 hours of filtration of toxic microalgae by bivalves can be enough for the shellfish to become toxic to human consumers (Scoging, 1998). Shellfish accumulate microalgal biotoxins generally in the digestive gland (hepatopancreas), but also in other tissues. Toxin retention time varies according to the group of toxins, the tissues in which the toxins are located, and also the shellfish species (Schantz, 1984). Very effective and protective monitoring and regulatory mechanisms have been established to limit human risk of exposure to toxic shellfish in many countries (Shumway et al., 1988; Smayda, 2004; Andersen et al., 2004; Todd, 2004; FernaÂndez and Shumway, 2004; Backer et al., 2004). Thus, public health relies on biotoxin monitoring programs, which generally close shellfish-harvesting areas when shellfish are likely to be contaminated. In non-industrialized countries, the monitoring programs are usually not as well developed, and more cases of intoxication can or do occur. Monitoring programs can be based upon one of two fundamental approaches: either by monitoring the water for presence of harmful algal taxa, or monitoring shellfish for toxicity. (See Chapter 5, this volume for a more detailed discussion on monitoring programs and regulations.)
2.2
Origins of phycotoxins
There are several types of biotoxins produced by various phytoplankton species. Phycotoxins are accumulated and can sometimes be metabolized by shellfish. These compounds can be noxious or lethal to humans. The toxins are listed in Table 2.1; included are compounds in the saxitoxin (STX) family (20 different toxins, including STX, NeoSTX, GTX) responsible for PSPs. Okadaic acid (OA), the dinophysistoxins (DTX), the pectenotoxins (PTX), and yessotoxins (YTX) are responsible for DSP; NSP is attributed to the exposure of shellfish to a group of polyethers called brevetoxins. ASP is caused by an amino acid, domoic acid, as the contaminant in shellfish. Finally, azaspiracid toxins are responsible for AZA. Some algal species also produce toxins that have not yet been identified or fully characterized chemically, such as Prorocentrum minimum, Prorocentrum micans, Pfiesteria spp., and Alexandrium monilatum. Human sensitivity to microalgal biotoxins varies among individuals; recorded intoxication levels vary between 144 and 1660 g per person, and lethal quantities can reach 300 to 12 400 g pr person (Van Egmond et al., 1993). PSTs are produced by dinoflagellate species, such as Alexandrium species, Gymnodinium catenatum and Pyrodinium bahamense var. compressum, and by freshwater cyanobacteria such as Anabeana circinalis, Aphanizomenon flosaquae, Cylindrospermopsis raciborskii, Lymgbya wollei, and Planktothrix sp. These species produce a group of toxins referred to as saxitoxins and saxitoxin derivatives. Saxitoxin was first isolated from toxic Washington butterclams, Saxidomus gigantea (Schantz et al., 1957; Schantz, 1960) harvested in the state of Washington (western North America), hence the name. Alexandrium
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(formerly Gonyaulax and Protogonyaulax) catenella (Whedon et Foid) Balech, was the first dinoflagellate to be associated with PSP toxins. Indeed, 102 people were sickened and six people were killed after ingestion of shellfish exposed to A. catenella near San Francisco in 1927 (Sommer and Meyer, 1937). Since then, Scoging (1998) reported about 1600 annual cases of PSP worldwide, of which about 300 can be attributed to Alexandrium. The two major species of dinoflagellates responsible for PSP exposure in humans are Alexandrium catenella and A. tamarense (Shumway et al., 1988; Landsberg, 2002). These species are present in different regions and in different concentrations. Additionally, toxin composition can vary drastically from one species or strain to the next, according to geographic origin, environmental conditions, or culture conditions (Cembella et al., 1988; Anderson, 1990; Anderson et al., 1990, 1994). As toxin compositions and concentrations in the various toxin-producing dinoflagellates differ according to all these criteria, the amount of toxin to which any animal is exposed will differ as well (Landsberg, 1996). Blooms of toxic algae producing PSTs occur worldwide; they have been recorded throughout Europe, on coasts of South and North American continents, and in coastal Africa, as well as throughout the Asian continent and in the Pacific Ocean. The different types of PSTs are STXs and derivatives, carbamate toxins (C toxins), gonyautoxins (GTX), neosaxitoxins (neoSTX), deoxydecarbamoylsaxitoxins (doSTX), and decarbomoylsaxitoxins (dcSTX), associated with several strains of dinoflagellates. Twenty-six derivatives of saxitoxins have been identified in shellfish or in the dinoflagellates responsible for PSP (Lagos and Andrinolo, 2000). PSTs are potent neurotoxins that bind to site 1 on voltagedependent sodium channels, blocking the influx of sodium into excitable cells (Kao, 1966), thus limiting signal transmission between neurons. PSPs in humans have thus far been caused exclusively by toxic dinoflagellates (Shumway, 1990, Hallegraeff, 2004); human PSP intoxication linked to saxitoxins from cyanobacteria has not been reported. Indeed, most cases of PSP in humans have been observed following ingestion of bivalves (Shumway, 1990), but also a few gastropods and crustaceans (Shumway, 1995), and, very rarely, toxic fish (Maclean, 1979; Adnan, 1984). PSTs can accumulate throughout the food chain, and experiments with copepods have highlighted the mechanisms controlling the availability of toxins to higher trophic levels (White, 1979, 1981; Turriff et al., 1995; Teegarden and Cembella, 1996; Turner et al., 2000). NSP results from consumption of shellfish contaminated with brevetoxins. Neurotoxic shellfish poisoning following bivalve consumption has been known in Florida since the late 1800s (Walker, 1884), but the cause of these poisonings was not identified until the 1960s (McFarren et al., 1965; Steidinger, 1993). Cases of NSP in the Gulf of Mexico have been mostly associated with the consumption of filter feeders, such as eastern oysters, Crassostrea virginica, quahogs, Mercenaria mercenaria, and M. campechiensis, surfclams, Spisula solidissima raveneli; sunray venus, Macrocallista maculata; coquinas, Donax variabilis; cross-barred venus, Chione cancellata; and a few other species.
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Brevetoxins are produced by the dinoflagellate Karenia brevis, and by the raphidophyte Chattonella cf. verruculosa. Brevetoxin-like compounds have also been identified in other algal species, such as Karenia spp. and several raphidophytes, Chattonella antiqua, C. marina, Heterosigma akashiwo, and Fibrocapsa japonica (reviewed in Landsberg, 2002). Nine brevetoxins have been isolated from K. brevis (Baden, 1989; Schulman et al., 1990): PbTx-1 to PbTx-3 and PbTx-5 to PbTx-10. Karenia spp. blooms generally produce more toxins than raphidophyte blooms and are therefore much more toxic to aquatic organisms; they are the only blooms reported to have effects on human health. Blooms of Karenia sp. triggering NSP originally occurred in the Gulf of Mexico and were observed only there until 1987 (Steidinger, 1993). In 1987 to 1988, the Atlantic coast of Florida was closed for shellfish harvesting because of NSP, and the Gulf Stream transported the Florida Atlantic coast Karenia brevis red tide to the coast of North Carolina. NSP was also observed in New Zealand for the first time in 1992 (Bates et al., 1993; Chang et al., 1995; Satake et al., 1996). Brevetoxins can cause massive fish kills, but thusfar, no human deaths have been associated with NSP. DSP cases have been observed worldwide; however, most cases have been reported in Europe, North and South America, Japan, and Southeast Asia (Sechet et al., 1990). Diarrheic shellfish toxins (DSTs) are lipophilic toxins produced by various species of phytoplankton within the genera Dinophysis and Prorocentrum. Toxin production varies appreciably between the different species of dinoflagellates, but also according to regional and seasonal parameters (Vale and Sampayo, 2003). The first group of toxin-producing species, common in Europe, D. acuta and D. acuminate, produce primarily OA; whereas, in Japan, Dinophysis fortii is the principal source of DTX-1 (Landsberg, 2002). OA and DTX mainly accumulate in the hepatopancreas of shellfish (Edebo et al., 1988; Alvito et al., 1990; Aune and Yndestad, 1993). The second group of toxins responsible for DSP are PTXs, neutral toxins consisting of polyether-lactones. PTXS are produced by various Dinophysis species. Ten PTXs have been isolated thus far, six of which have been chemically identified: PTX1, -2, -3, -4, -6 and -7. The third group of toxins associated with DSP events includes a sulfated polyether, YTX, and its derivatives. YTX was first isolated from the digestive organs of scallops (Patinopecten yessoensis) in Japan (Ciminiello et al., 1999). DSP in humans has exclusively been observed following consumption of shellfish. ASP was first observed in Prince Edward Island, Canada, in 1987, after 107 people became ill from eating contaminated blue mussels (Todd, 1993). ASP cases have been mainly observed in Canada and in the USA. A few occurrences of ASP have been recorded in Europe, Australia, New Zealand, and Japan (reviewed in Hallegraeff, 2004). The toxin responsible for ASP, domoic acid (DA), was originally discovered as a product of a red macroalga, Chondria armata, and was later isolated from several other red macroalgae. Subsequently, DA was demonstrated to be produced by diatoms as well: Pseudo-nitzschia spp. and Nitzschia spp., but not by dinoflagellates. The ASP observed in humans
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Shellfish safety and quality
following the consumption of shellfish is exclusively attributable to DA produced by diatoms ± nine species of Pseudo-nitzschia and one species of Nitzschia (reviewed in Landsberg, 2002). AZP was first observed in November 1995 when several people from the Netherlands became ill after eating mussels (Mytilus edulis) cultivated in Killary Harbour, Ireland (McMahon and Silke, 1996; Satake et al., 1998a). A few differences between symptoms of AZA poisoning and DSP, such as a slowly progressing paralysis, were observed in the mouse assay using mussel extracts. Moreover, no trace of any DSP-producing, toxic organism was detected in the waters. Thus, azaspiracid (formerly called Killary Toxin-3 or KT3) was identified as a new toxic syndrome. Azaspirazids (AZA) represent a new group of phytotoxins, first identified on the Irish coast, but now known to occur over the entire western coast of Europe. Several analogues of AZAs have been identified: AZ-1-11. AZAs are polyether toxins (Furey et al., 2003); they accumulate in tissues of bivalves, which were previously exposed to harmful algae in the genus Protoperidinium, considered non-toxic previously. AZA has been associated specifically with the microalga Protoperidinium crassipes (Gribble, 2002). AZAs can be accumulated in several bivalve species, not only mussels. The presence of AZA has been reported in many different species of bivalves, including mussels (M. edulis), oysters (C. gigas), scallops (P. maximus), cockles (C. edule), and clams (T. philippinarium), and can reach levels that exceed the EU regulatory limit (0.16 g total AZAs/g). Tetrodotoxin (TTX) represents another type of potential human poisoning. TTX poisoning has been observed in humans following consumption of finfish but it has also been observed in bivalve molluscs, which can represent a source of human shellfish poisoning. TTX has been associated with bacteria, Vibrio, and Aeromonas (Noguchi et al., 1986; Tamplin, 1990), and more recently with Alexandrium tamarense or the bacteria associated with it (Kodama et al., 1993, 1996). Direct association between dinoflagellates and human cases of TTX poisoning has not been demonstrated (Landsberg, 2002). TTX produces similar symptoms in humans to PSP, as it blocks site 1 of the voltage-dependent sodium channel in nerve and muscle membranes (Kao, 1993). Spirolides have been observed in shellfish from Nova Scotia, Canada, Northern Europe: Norway, Denmark, and from the Gulf of Maine, USA (Aasen et al., 2005; Gribble et al., 2005; MacKinnon et al., 2006). These neurotoxins are produced by Alexandrium ostenfeldii (Cembella et al., 2000; Hu et al., 2001) and can trigger rapid death of mice injected with extracted spirulides (Landsberg, 2002). Thus far, even though it remains a potential toxic agent, no instance of human poisoning has been recorded. Other algae Alexandrium monilatum produces hemolysins, which have demonstrated neurotoxic effects on mice experimentally (Clemons et al., 1980; Erker et al., 1985). These hemolysins are different from STX or GTX. Moreover, recent analyses from Hsia et al. (2006) reported that A. monilatum also produces the toxin they
Biotoxin contamination and shellfish safety
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termed goniodomin A, a toxin also produced by the Japanese rockpool dinoflagellate, Alexandrium pseudogoniaulax. Alexandrium monilatum causes sedation, abdominal constriction, fecal clumping in the perianal area, ataxia, tremors, cyanosis, loss of reflexes, convulsions, and death of adult mice. This dinoflagellate has been associated with fish kills in the last few years. Thus, entire cells as well as extracts of A. monilatum have been tested on mice, rats and fishes and triggered death of these species (Gates and Wilson, 1960; Aldrich et al., 1967; Clemons, et al., 1980; Erker et al., 1985). So far, no poisoning has been observed in humans attributable to A. monilatum. Prorocentrum micans has sometimes been suspected to be associated with shellfish poisonings, even though most studies have shown P. micans to be a non-toxic algal species (Landsberg, 2002). In Portugal, in 1955, several human poisonings and one death followed the consumption of toxic cockles, exposed to a bloom of P. micans (Pinto and Silva, 1956). The toxins of P. micans have not been identified, but the poisoning involved various neurological symptoms, including loss of sensitivity in the lips and chin, numbness and paraplegia, tremors, ataxia, and a feeling of floating. As these observations have not been repeated at other times in other places experiencing P. micans blooms, it appears possible that the one reported incident of poisoning may have been caused by a co-occurring, undetected microalgal species. Prorocentrum minimum has also been suspected to be responsible of venerupin shellfish poisoning (VSP) (Denardou-Queneherve et al., 1999). Putative toxin from this species is uncharacterized so far, but a B-diketone, or some uncharacterized venerupin or prorocentrin have been suspected to be responsible for the toxicity (see tables in Lansberg, 2002). Venerupin, which has been reported several times as the causative agent of hepatotoxicity following ingestion of poisonous clams and oysters in Japan, however, also has never been characterized (Hashimoto, 1979). Williams et al. (1997) observed tumorpromoting and hepatotoxic microcystin in mussels. These results, associated with the observations that liver damage found in all venerupin-related events, suggest that `venerupin' toxicity could also have been caused by microcystins. This suggests that microcystins could potentially be responsible for a new type of shellfish intoxication, hepatotoxic shellfish poisoning (HSP). Microcystins are produced by cyanobacteria, which occur mainly in fresh or brackish waters, but have also been observed in marine waters. The effects of Pfiesteria spp. have been observed on human health. A hydrophilic Pfiesteria toxin (PfTx) was isolated from Pfiesteria piscicida in 1997 (Fairey et al., 1999), consisting of a metal±organic complex. This toxin was shown to affect fish and mammals; indeed, PfTx was demonstrated to be lethal to fish and toxic to mammals (Fairey et al., 1999; Burkholder et al., 2001, 2005; Moeller et al., 2001; Levin et al., 2003). Melo et al. (2001) purified PfTx and described a pharmacological mode of action. According to Moeller et al. (2007) the toxicity of Pfiesteria spp. is mediated by metal-containing, organic compounds which produce toxic free radicals based on carbon and sulfur. These
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radicals have a very short lifetime, which explains the difficulty scientists encountered to isolate and characterize the toxins from Pfiesteria spp. Pfiesteria spp. also affects bivalve shellfish health (Springer et al., 2002; Shumway et al., 2006); therefore, the consumption of bivalves exposed to Pfiesteria spp. by humans could trigger a risk of poisoning.
2.3
Trophic dynamics of phycotoxins in molluscan shellfish
The intensity and the occurrence of phytoplankton blooms leading to toxic shellfish have increased drastically in the past few decades. Hallegraeff (2004) proposed four major reasons to explain this rise in outbreaks of toxic phytoplankton: the increase of scientific awareness of toxic species; enhanced utilization of coastal waters for aquaculture; stimulation of harmful algal blooms attributed to eutrophication or unusual climatic conditions; and the transportation of cells or cysts of harmful algae through ballast water or via transportation of shellfish from one body of water to another. Toxin content in shellfish is a function of uptake, metabolism, and depuration. Rates of these processes are different for various toxins and shellfish species, complicating prediction of exposure risk for human consumers of the shellfish. Indeed, toxin accumulation and retention times vary among different bivalve species and depend upon intrinsic and extrinsic factors, such as the concentration of the bloom as well the toxicity of the algal cells and by the filtration and elimination rates of the shellfish (Shumway, 1990). 2.3.1 Toxin uptake There are different vectors by which shellfish are exposed to microalgal toxins. The exposure can be direct, with toxin transfer directly from the algal cells, whether they are intact or lysed, or indirectly, by consumption of other grazers already contaminated, as toxins accumulate through the food chain. Harmful microalgae are generally suspended in the water column and accordingly are encountered by filter feeders. Toxic cells, or particle-associated toxins, can sometimes sink to the bottom as well. Some harmful microalgae can also be benthic organisms, and some, especially dinoflagellates, have sedimentary cysts or resting stages as part of their life cycles. Sometimes these stages are even more toxic than the free-swimming stages; cysts of Alexandrium can for example be 100 times more toxic that the vegetative cells after they have undergone several months of dormancy (Dale and Yentsch, 1978). These toxins, cells, or cysts present in the sediment are consumed by benthic organisms, thereby exposing the deposit-feeders to biotoxins. Nevertheless, filtration of planktonic, vegetative cells represents the most direct means of exposure in bivalve molluscs. Shellfish can also be affected by the extracellular toxins (exotoxins) exuded into the surrounding water by many microalgae. Another direct exposure of
Biotoxin contamination and shellfish safety
53
shellfish to microalgal toxins consists of direct contact with the cells, either because the microalgal toxins are present at the surface of the cells or through mechanical damage caused by the microalgae interacting with the gills or epithelial tissues of the bivalve (Landsberg, 2002). These two processes do not generally result in accumulation of toxins within the shellfish and therefore are not often responsible for human intoxication. Some shellfish, such as some gastropods and crabs, feed on bivalves or other organisms and can thereby be exposed to toxins indirectly from prey that have already accumulated phycotoxins. Toxins bioaccumulated can, in many cases, have been bioconverted and biomagnified and are now available to animals not feeding on microalgae. Humans are examples of predators only exposed to biotoxins following indirect exposures, as we may eat shellfish previously exposed to microalgal toxins. 2.3.2 Biotoxin accumulation As described in the previous section, bivalve shellfish can be filter feeders or benthic, deposit feeders. Thus, if any harmful algae are present in the water or in the sediment, the bivalves encounter them. The algae then can be ingested and pass through the stomach and digestive system, or be rejected in the form of pseudofeces. When the cells are ingested, they can be assimilated or rejected as intact cells in the feces. As the harmful algal cells pass through the gut and the digestive tract, some are digested and the toxins are accumulated in the tissues of the bivalves, which affects the accumulation rate of toxins in the shellfish (Morono et al., 2001) as well as the gut passage time. The ability of shellfish to retain and accumulate toxins varies greatly according to the shellfish species (Bricelj and Shumway, 1998; Fernandez and Shumway, 2004). In Maine coastal waters, PSTs were detected 12 days earlier in mussels Mytilus edulis than in clams Mya arenaria; M. edulis accumulate two to four times more PST than the clams (Hurst and Gilfillan, 1977; White, 1982; Larocque and Cembella, 1991; Bricelj and Shumway, 1998). Hence, mussels appear more appropriate for monitoring purposes as they accumulate more toxin, and more rapidly, than most of the other bivalve species, thereby allowing earlier detection. Thus, the toxicity of shellfish species can vary according to intrinsic parameters, such as previous exposure to HABs, uptake dynamics and detoxification capabilities, tissues affected, feeding rate, and food retention. Extrinsic factors associated with the phytoplankton species and the bloom characteristics, amounts of toxins present in the algal species, or the environmental conditions also affect toxin accumulation in shellfish. Some species of shellfish, when exposed to a harmful alga, reduce or suppress completely filtration, moderating or eliminating the accumulation of toxins internally (Fernandez and Shumway, 2004). Filtration rate can be reduced as a shellfish species avoids the specific harmful alga, or because the harmful algae affect the mechanism for food capture. Fernandez and Shumway (2004) reported the retraction of the siphon of quahogs, Mercenaria mercenaria, when exposed
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Shellfish safety and quality
to Alexandrium tamarense, followed by a complete isolation of the clams from the environment by shell-valve closure. Efforts to induce toxicity in these clams appeared to be unsuccessful. In contrast, blue mussels, Mytilus edulis, are known to accumulate large concentrations of PST when exposed to toxic dinoflagellates (Shumway, 1990). Certain bivalve shellfish can accumulate more toxin than others. Mussels, M. edulis, feed actively on toxic cells (Bricelj et al., 1990), as they possess nerves insensitive to PSP toxins; this species therefore accumulates high toxin levels (Bricelj and Shumway, 1998). In contrast, oysters, Crassostrea virginica, reach only a very low level of toxicity; they are very sensitive to PSP toxins (Bricelj and Shumway, 1998) and display behavioral and physiological mechanisms to eliminate or diminish exposure to the toxic algae. DSP-producing species have, to date, not been demonstrated to affect the filtration rate of shellfish, neither in the field, nor in laboratory experiments (Bauder et al., 2001). Similarly, Whyte al al. (1995) showed no effect of the ASP-producing species Pseudo-nitzshia multiseries, on feeding by mussels, Mytilus edulis. Conversely, DSP-containing dinoflagellates seemed to induce valve closure in Crassostrea gigas (Jones et al., 1995). Karenia brevis, responsible for the production of the brevetoxins triggering NSP, affected the valve closure of oysters C. gigas and mussels Brachiodontes recurvis (Sievers, 1969). Shellfish toxicity also depends upon the biomass of the organisms coupled with the amount of biomass into which the toxins are distributed. Thus, as the toxin content remains constant, if an organism, after toxin contamination, changes its weight, toxicity will be adjusted as well. Moreover, bivalves accumulate different concentrations of toxins in different tissues. Overall, the visceral mass of bivalves seems to concentrate toxins more than the rest of the body (review in Bricelj and Shumway, 1998). Clams also accumulate most of the toxins primarily in their siphon, as it is the first organ in contact with the algal cells. Similarly, brevetoxins accumulate in the gut and hepatopancreas in most species (McFarren et al., 1965; Steidinger et al., 1973, 1993; Hemmert, 1975; Roberts et al., 1979; Baden et al., 1982; Tester and Fowler, 1990; Steidinger et al., 1998). The accumulation of PSP toxins has been demonstrated to vary among individuals in a bivalve population according to extrinsic, environmental factors, including: temperature, microhabitat differences, especially availability of the toxic dinoflagellates, concentration of the toxic dinoflagellates, or tidal immersion time of the shellfish population (Quayle, 1969; Bricelj and Shumway, 1998). The difference can also be the result of intrinsic factors, such as feeding rate of the shellfish or variation in body mass (Bricelj and Shumway, 1998). Bricelj et al. (2005) demonstrated also a genetic basis for differences in biotoxin accumulation in individual shellfish. They suggested the existence of genetic selection for resistance to PST in areas subjected to PSP outbreaks of sufficient intensity and occurrence. Interspecific differences in the accumulation of PST have been associated with differences in nerve resistance to toxins, as measured by the concentration of saxitoxin necessary to block the conduction of the nerve action potential in in vitro trials (Twarog et al., 1972).
Biotoxin contamination and shellfish safety
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Bivalves can also accumulate different concentrations of DST depending on many factors (Alvito et al., 1990; Suzuki and Mitsuya, 2001), and can remain toxic for variable periods of time. OA and DTX mainly accumulate in the hepatopancreas of shellfish (Edebo et al., 1988; Alvito et al., 1990; Aune and Yndestad, 1993). Domoic acid (DA) has also been observed in various species of bivalves (reviewed in Landsberg, 2002). The responses of each bivalve to the diatoms and to the DA itself vary according to the bivalve and algal species and are very species-specific. Moreover, DA accumulates in various tissues in different concentrations. In mussels collected from a naturally contaminated site, for example, 93:4 1:9% of the DA contained in the whole animal was found in the hepatopancreas, which represents only 30% of the total biomass of the animal (Grimmelt et al., 1990). Thus, the amounts of toxins in the body of animals depend upon the nature of the tissues. Additionally, detoxification rates are different in each tissue. As DA is a hydrophilic toxin, a large portion of it is excreted instead of being bioaccumulated (Novaczek et al., 1992). When accumulated, DA can also be biotransformed; for example, DA present in the digestive gland of mussels can be converted into isodomoic acid isomers (Wright et al., 1990). The balance in the organisms of the accumulation of toxins is regulated by the intake, mostly from direct filtration or consumption of organisms already toxic, the loss to the environment, and by biotransformation into other toxins (Fernandez and Shumway, 2004). 2.3.3 Biotransformation Toxins accumulated within the tissues of bivalve molluscs may undergo metabolic reactions by which they are changed chemically, sometimes into more or less toxic forms. These biotransformations may be a consequence of digestive or active detoxification processes, or may simply result from participation of toxins in general metabolic processes intra- or extracellularly. Metabolic reactions involving phycotoxins can be quite important in determining the safety of human consumers of shellfish, particularly in terms of estimating risk of illness when toxin-producing phytoplankton are observed in the water and determining a safe moratorium on harvest after toxin accumulation by shellfish. Specific instances of biotransformation of phycotoxins within bivalve tissues are mainly apparent from mis-matches between toxin profiles of toxic phytoplankton ingested and toxin profiles within the animal over time. For example, some PTXs, such as PTX-2, have been observed exclusively in dinoflagellates and not in shellfish, suggesting that an oxidation reaction takes place in the hepatopancreas of shellfish producing other PTXs. Indeed, oxidation of PTX2 to PTX6 in scallops (Patinopecten yessoensis) was demonstrated by Suzuki et al. (1998). Evidence of bioconversion of brevetoxin PbTx-2 in PbTx-3 in various shellfish species has also been demonstrated in Florida following a NSP outbreak, as PbTx-2 and 3 were present in the water, but only PbTx-3 was found in shellfish (Poli et al., 2000).
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Issues related to biotransformation of microalgal biotoxins within the tissues of edible bivalves may become more important to seafood-safety monitoring if and when chemical toxin-detection technologies replace bioassays currently employed. The chemical specificity inherent in chromatographic or immunologically based (e.g., ELISA) methods carries the risk that small, structural changes in toxin molecules resulting from biotransformation may render stilltoxic metabolites undetectable. Clearly, chemical toxin-detection methods will need to be informed by knowledge of biotransformation of microalgal toxins within the tissues of bivalves sampled for testing. 2.3.4 Natural shellfish detoxification The rate of detoxification in shellfish depends upon toxin and shellfish species (Shumway, 1990; Bricelj and Shumway, 1998; Fernandez and Shumway, 2004). Data are mostly available for commercially important shellfish species. Fernandez and Shumway (2004) reviewed the approximate retention times of DA for various species of bivalve molluscs, by recording the time it took for the bivalve species to reach a level of toxin below quarantine or detection level. The retention time varies considerably according to the species, from a few hours to a few days for DA in the mussels Mytilus edulis and M. galloprovincialis (Novaczek et al., 1992, Blanco et al., 2002a) to several months or years for DA in the scallop Pecten maximus (Blanco et al., 2002b). Toxin elimination can occur fairly rapidly in some cases. Fletcher et al. (1998) reported that oysters, Crassostrea gigas, which reached a NSP level of 25 to 100 mouse units (MU) per 100 g of drained oyster meat, following a 24 hour exposure to K. brevis cells at a concentration of 10 to 25 millions per oyster, reduced their level of toxicity considerably and were almost at the acceptable regulatory limit for human consumption, after only 3 days of depuration. Toxin depuration is also highly dependent upon the tissues in which the toxins are accumulated. Toxins present in the digestive glands are usually eliminated more rapidly than toxins present in the other organs. DSP toxins in Argopecten irradians (Bauder et al., 2001) and DA in Pecten maximus (Blanco et al., 2002b) appear to be exceptions. The factors regulating depuration are not fully understood; for example, relative importance of season (Prakash et al., 1971), or water temperature (Shumway and Cembella, 1993) are difficult to differentiate. Elevated temperature has been reported to retard DSP toxin loss in Mytilus galloprovincialis (Blanco et al., 1999), to advance ASP toxin loss in Mytilus edulis (Silvert and Subba Rao, 1992; Novaczek et al., 1992) and Pecten maximus (Blanco et al., 2006), and to have no effect on the PSP toxin loss of Saxidomus giganteus (Madenwald, 1985). The amount of non-toxic phytoplankton ingested appears to be another factor enhancing slightly the detoxification rate in shellfish exposed to PSP and DSP toxins (Sampayo et al., 1990; Marcaillou-Le Baut et al., 1993; Blanco et al., 1997, 1999). Metabolic processes have also been cited as affecting detoxification rate of DA in Pecten maximus (Fernandez and Shumway, 2004).
Biotoxin contamination and shellfish safety
2.4
57
Human health impacts
As humans consume shellfish, they can be exposed to several types of poisoning syndromes from the phycotoxins accumulated in shellfish. The five major types of shellfish poisoning defined above are, after all, based upon clinical symptoms in human victims: PSP, NSP, ASP, DSP, and AZA. The symptoms associated with each type of poisoning present general and specific characteristics. 2.4.1 Effect on humans, toxins involved and their biochemical mechanisms PSP begins with paresthesia and a sensation of tingling and numbness around the lips and the mouth, then the face and the neck. A sensation of muscular weakness, of lightness and feeling of floating develops, followed by uncoordinated gestures, dizziness, ataxia, incoherence, and progressive respiratory depression. In high concentration, PSP can lead to respiratory paralysis and can be fatal to humans (Catteral, 1985; Kao, 1993). NSP in humans is characterized by paresthesia, reversal of temperature sensations, fever, myalgia, dizziness, vertigo, ataxia, muscle or abdominal pain, chills, nausea, diarrhea, burning pain in the rectum, headache, bradycardia, and dilated pupils (Steidinger et al., 1973; Hemmert, 1975; Baden, 1983, 1988; Morris et al., 1991). Despite the intense symptoms, to date, no human deaths have been reported following NSP exposure. NSP cases have been recorded in the west coast of Florida in the US, on the southern Atlantic coast, in the Caribbean Sea and in New Zealand. Toxins responsible for NSP are potent neurotoxins and hemolysins called brevetoxins. Brevetoxins are complex, polycyclic ethers, which bind to sodium channel site 5 on the voltage-sensitive channels on neurons and alter the properties of the excitable cells, shifting activation to more negative potentials, thereby triggering membrane depolarization (Huang et al., 1984; Catterall, 1985; Poli et al., 1986; Lombet et al., 1987). Brevetoxin pathogenic dose for human is very low, about 42±72 MU, and oral LD50 value in rats varies between 520 to 6600 g/kg (Llewellyn, 2001). Acute symptoms associated with DSP in humans involve gastrointestinal pain, diarrhea, nausea, and vomiting (Aune and Yndestad, 1993; Quilliam and Wright, 1995). Symptoms can appear as soon as 30 minutes after ingestion of the shellfish and last for 3 or 4 days; however, DSP has never been reported to generate human mortalities. DSP in humans is attributable to consumption of shellfish containing diarrheic shellfish toxins (DSTs), including three groups of toxins: (1) OA and DTX1-4, which are OA derivatives (Aune and Yndestad, 1993); (2) PTX (Yasumoto et al., 1985; Murata et al., 1982, 1986); and (3) YTX (Yasumoto and Satake, 1998). OA and DTXs are produced by Dinophysis spp. and Prorocentrum lima. Biochemical mechanisms by which these toxins affect humans exposed have been well described. OA inhibits protein phosphatases types 1 and 2A, which increases protein phosphorylation. Consequently, intracellular processes, such as metabolism, cellular division, contractility, gene transcription, maintenance of cytoskeletal structure, membrane transport and secretion, and receptor-mediated
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Shellfish safety and quality
signal transduction are affected. Moreover, OA stimulates expression of certain proto-oncogenes and activation of H1 kinase in vitro. Finally, OA induces several mitosis-specific events (Bialojan and Takai, 1988; Fujiki et al., 1989; Haystead et al., 1989; Herschmann et al., 1989; Yamashita et al., 1990; Sakai and Fujiki, 1991; Fujiki and Suganuma, 1993; Rossini, 2000). Mice were exposed experimentally to OA, DTX-1, and DTX-3 to assess short-term effects and estimate exposure levels. After less than 15 minutes, severe diarrhea associated with destruction of the absorptive epithelium of the ileum villi occurred (Terao et al., 1993). OA triggered rapid changes the large and small intestines of rats, leading to hypertension, and accumulation of goblet cells (Edebo et al., 1988; Lange et al., 1990). DTX-1 was responsible for excessive fluid accumulation in the intestines of suckling mice (Hamano et al., 1985). In adult mice, the level of toxins triggering diarrhoea is equal to or above 40 g for OA and equal to or above 35 g for DTX1±4 (Scoging, 1998). PTXs are neutral toxins, consisting of polyether-lactones, which originate from Dinophysis species also (reviewed in Landsberg, 2002) and are present in various species of bivalves. Possible impacts of PTXs on human health are not yet clearly defined, and neither is the mechanism of action. Similarly, the mechanisms of action, as well as the impact on human health, of YTX and its derivatives are poorly understood. These toxins are known to be sulfated polyethers, but molecular structures have not been clearly defined. YTXs appear to interact with calcium channels (de la Rosa et al., 2001). YTXs are produced by the dinoflagellate Gonyaulax grindleyi ( Protoceratium reticulatum) (Satake et al., 1997; Draisci et al., 1999). ASP is the result of DA intoxication. The acute symptoms are nausea, vomiting, abdominal pain, diarrhea, and neurological effects such as memory and consciousness loss, seizures, dizziness, disorientation, and confusion (Debonnel et al., 1989; Wright et al., 1989; Perl et al., 1990; Teitelbaum et al., 1990; Todd, 1990, 1993; Nijjar and Nijjar, 2000). In cases of severe intoxication, these symptoms may be followed by coma or death in human victims. Chronic symptoms of permanent amnesia have been reported, including loss of short-term memory (Todd, 1993), but the extent of the chronic effects of ASP are still unclear. DA is part of the group of amino acids called kainoids, which are exocitotoxins or neuroexcitants that affect the mechanisms of neurotransmission in the brain (Quilliam, 2004). It is a crystalline, water-soluble, heat-stable toxin with properties of a typical amino acid; it is an analogue of glutamate, and acts as a glutamate receptor agonist. DA binds to the kainite type of glutamate receptor in the brain, especially in the hippocampus, acting as an excitatory neurotransmitter (Debonnel et al., 1989; Sutherland et al., 1990). DA causes depolarization of neurons, followed by an influx of cellular calcium ions, neuronal swelling, and cell death (Novelli et al., 1990; Bates, 1998). The neurons located in the hippocampus are responsible for memory retention, which explains the loss of memory associated with ASP (Bates, 1998). AZP has been recognized recently as a serious risk for human health (Ofuji et
Biotoxin contamination and shellfish safety
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al., 1999a) and is attributed to the toxin azaspiracid (Satake et al., 1998a,b; Ofuji et al., 1999a,b; Draisci et al., 2000; James et al., 2000). The symptoms ± nausea, vomiting, severe diarrhea, and stomach cramps ± are very similar to those of DSP (Satake et al., 1998a). Azaspiracid-2 (AZA-2) and azaspiracid-3 (AZA-3) have been identified specifically as 8-methylazaspiracid and 22-demethylazaspiracid, respectively (Ofuji et al., 1999a). Two analogs, azaspiracid-4 (AZA-4) and azaspiracid-5 (AZA-5), were identified as 3-hydroxy-22demethylazaspiracid and 23-hydroxy-22-demethylazaspiracid (thus hydroxylated analogs of AZA-3) (Ofuji et al., 2001). James et al. (2003) recently isolated AZA-6±11 five new hydroxyl analogs of azaspiracids. Despite importance in human health, the mechanisms of action of azaspiracid are still largely unknown. Several studies are ongoing to understand the mechanisms involved in azaspiracid poisonings. Twiner et al. (2005) highlighted cytotoxic and cytoskeletal effects of azaspiracid-1 (AZA-1) on mammalian cell lines. Alfonso et al. (2005) also demonstrated that azaspiracid-4 (AZA-4) inhibits plasma membrane Ca+2 entry by stored, operated channels in Ca signaling within human T lymphocytes. TTX can be responsible for fatal human poisoning; it has been recorded in numerous species of finfish, but has also been found in gastropods (Shumway, 1995; Lin and Hwang, 2001) and in molluscs such as the Japanese scallop, Patinopecten yessoensis (Kodama et al., 1993). TTX produces similar types of symptoms as STX; these toxins are chemically different, but their mechanisms of action are fairly similar. TTX blocks the voltage-dependent sodium channels in nerve and muscle membranes (Kao, 1993). The effects of TTX on aquatic organisms are currently unknown. Other algae Prorocentrum micans has also sporadically been associated with human intoxications, but no toxins have been identified. Indeed, in 1955, P. micans was present in Portuguese waters when humans ingested toxic cockles. One death occurred, as well as various symptoms, such as neurological abnormalities, loss of sensitivity in the lips and the chin, lack of feeling in the arms and hands, paraplegia of the legs, ataxic walking, and floating sensations were reported (review Landsberg, 2002).
2.5
Management responses
Programs and guidelines have been developed to minimize risk of human exposure to phycotoxin-contaminated shellfish. These programs are based upon intervening at several points in the sequence of activities leading to consumption of shellfish: (1) stopping harvest when potentially toxic phytoplankton species are present, (2) sampling shellfish populations before harvest and analyzing toxicity in the shellfish tissues, (3) post-harvest removal of specific tissues containing phycotoxins, and (4) issuing advisories to cook harvested shellfish
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thoroughly in cases wherein phycotoxins are heat-labile. These programs can involve, therefore, highly developed monitoring programs as well as safe depuration techniques, both discussed in the following section. 2.5.1 Monitoring programs Biotoxin accumulation in shellfish makes them toxic for human consumption. Thus, monitoring programs and regulations need to be established to limit the risk of intoxication of humans by toxic shellfish. To be efficient, the monitoring programs must allow rapid detection of any toxic algae or presence of toxic shellfish, to be able to limit the harvest and avoid contamination, as well as to prevent unnecessary disposal of toxic shellfish already harvested. Thus, many countries have monitoring programs screening the phytoplankton present in the water, but also the flesh of the shellfish on a regular basis. Moreover, a good network needs to be developed to be able to transfer the information very rapidly and efficiently to seafood harvesters, distributors, and consumers, as well as public health and medical professionals. Thus, the major points of monitoring and management programs preventing biotoxin intoxication often include the following elements (Anderson et al., 2001): (1) environmental observations of the plankton, fish kills, and abnormal animal behaviors, (2) regular sampling of plankton, and shellfish, (3) analysis of samples of water and animals for presence and quantification of harmful algae and toxicity of shellfish, (4) evaluation of the results, (5) dissemination of information and implementation of regulatory action, and finally (6) action plan or mitigation measures. Each country develops its own monitoring program, identifying one or several agencies as responsible, and may also include industry, fishermen, or private consultants. Developing efficient monitoring programs, based upon phytoplankton observations and shellfish toxicity rather than relying only upon post-harvest toxin analyses, allows a very early response to protect the public and to minimize product losses for the harvesters and costs associated with shellfish poisoning. 2.5.2 Detection methods and analysis of toxins As mentioned previously, the first method for detection of harmful algae is visual monitoring of the water, both on a macro-scale and with the microscope. Observation of the phytoplankton present in the water gives early information about the potential occurrence of HABs and subsequent contamination of shellfish with phycotoxins. The second method involved in monitoring is to quantify toxins present in marine organisms. Each toxin has a detection level at which shellfish harvesting should be closed, and different toxins may be detected with different methods. The mouse bioassay is the most classical test for analysis of most of the biotoxins transferred by shellfish (Schantz et al., 1958). This method needs to be
Biotoxin contamination and shellfish safety
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standardized, with a known strain, size and condition of the mouse. Mice are then challenged by injection of shellfish extract, and responses are compared with the responses of animals injected with different known concentrations of toxins. The mouse bioassay is a rapid, inexpensive method, but it can lack specificity. Thus, other, more accurate detection methods are also used to identify the presence and quantities of toxins in animal tissues. Antibody-based immunoassays can be used for detection of certain marine toxins (Lewis, 2001), but there are only few available because of the difficulty to obtain sufficient quantities of pure toxins. These assays can be available in rapid test kits, such as `DSP Check Elisa Kit' (Camacho et al., 2007). Other methods used for the toxin determination include HPLC, which allows definition of the toxin profile in both shellfish and toxic algae (Oshima et al., 1984; Sullivan et al., 1985). Moreover, analytical chemistry techniques that combine the physical separation capabilities of liquid chromatography (aka HPLC) with the mass analysis capabilities of mass spectrometry (LC-MS) are also used for determination of toxins, such as ASP (Holland et al., 2003) and lipophilic marine algal toxins such as OA/DTXs, PTXs, YTXs, AZAs are detected with LC/MS (Aasen et al. 2003). The mouse bioassay detects the presence of domoic acid in the tissues only when the concentration of domoic acid is higher than 40 g per gram of shellfish meat. As this concentration is much higher than the regulatory level, an alternate method was developed to monitor the concentration of toxin present in shellfish. HPLC with ultraviolet detection (HPLC-UV) was the first method used to detect the presence of domoic acid in shellfish samples (Wright et al., 1989) and is the principal analytical method used at present. Other methods have also been developed, including thin layer chromatography (TLC), for semi-quantitative detection (Quilliam et al., 1998), capillary electrophoresis (CE), a promising new technique for detection of domoic acid in bivalves (Zhao et al., 1997), and LC-MS. Recently a series of in vitro assays have been developed for rapid detection of toxins in shellfish (Cembella et al., 2004). These assays generally are expensive to develop, but economies of scale can lower the cost to where these assays can become economically competitive. A receptor-binding assay for ASP toxins has been developed to detect DA and its analogues. Alternative assays have also been developed recently, such as the evaluation of a protein-phosphatase inhibition assay for monitoring OA and derivatives (Botelho et al., 2003) or caspase-8 activation associated with OA-induced apoptosis (Cabado et al., 2003, 2004). The two screening strategies included in most monitoring programs ± water and shellfish monitoring ± are both necessary under different circumstances, but each carries advantages as well as disadvantages (Todd, 2004). Analyses of phytoplankton are usually more cost-effective than shellfish testing, and results are faster, permitting quicker decisions concerning harvest of resources in the monitored area. Moreover, phytoplankton monitoring allows observation of all species of phytoplankton present in the water, not only the species producing toxins, providing additional information on trophic conditions. Unfortunately,
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determination of species of phytoplankton in water samples requires a very highly trained phytoplankton taxonomist. Further, local hydrographic conditions must be known so that areas sampled are chosen to represent surrounding areas (Todd, 2004). Similarly, the monitoring of the shellfish flesh presents advantages as it provides a straightforward answer on the question of possible health effects of eating the shellfish. Also, monitoring shellfish tissues for toxicity gives a cumulative picture of what may have occurred in the water previously as shellfish were accumulating and possibly metabolizing the toxins. Conversely, shellfish-tissue analyses are more expensive, require specialized chemists, and take longer than microscopic observation of phytoplankton samples, which can delay decisions on the management of fisheries (Todd, 2004). After the results of the sampling have been determined and evaluated by the several institutions responsible for making decisions regarding the management of the fisheries, dissemination of information followed by regulatory actions or mitigation measures take place. For example, Todd (1993) reported the toxicity level of DA for humans to be 1 to 5 mg/kg in the Prince Edward Island outbreak in 1987. At present, shellfish contaminated with a concentration of DA higher than 20 g per gram of shellfish meat is considered to be excessive for human consumption, and this became the regulatory level set by the Canadian regulatory authorities after the outbreak of ASP in Canada in 1987. Similarly, monitoring programs in Europe and in the US regularly detect high concentrations of PSTs (Van Egmond et al., 1993). As there are no known antidotes to PSP shellfish poisoning, monitoring programs are very well developed and need to be implemented very carefully. Shellfish are monitored in many states and countries to control shellfish harvest by commercial and recreational fishermen. Recognition of the potential lethality of PSP has led to increases in monitoring programs limiting the risk of human intoxication. Most human intoxication with PSP followed consumption of toxic bivalves (Shumway, 1990), but a few reports also record intoxication attributed to consumption of toxic gastropods and crustaceans (Shumway, 1995). Saxitoxin, and its numerous derivatives, accumulated in shellfish eaten by human consumers can cause PSP, but risk of exposure can be mitigated through monitoring and control programs. Blooms of PST-producing algae occur mainly from April to October in the US and in Europe and can occur several times a year. The blooms of toxic algae are dependent on variable factors, such as light, temperature and salinity of the waters, nutrient availability, niche availability, as well as other environmental factors. Blooms producing PSTs generally occur in `cold waters', but the waters have to be above 5±8 ëC for blooms to occur. As dinoflagellates can also form cysts, if the temperature decreases drastically or if environmental conditions change and become inhospitable, the dinoflagellates may form cysts and survive in surface sediments. PSP occurrence is not predictable, and shellfish can remain toxic for various periods of time afterward. Closure of shellfisheries must be determined based upon persistence of the toxins in the shellfish; therefore, some areas can be closed for short or very long periods of time.
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Commercial harvesting is regulated and stopped for various periods of time, but as recreational fishing is also important in many regions, a very large effort is also made to advise the public with harvest warnings through the media and signs posted on public beaches. 2.5.3 Processing for detoxification Processes for active detoxification have been developed to artificially eliminate toxins from shellfish after harvest and before sale, using two different groups of techniques. The first group of techniques includes temperature or salinity stresses, ozonation, chlorination, transplantation, and other methods that accelerate the rate of detoxification of shellfish while maintaining them alive. The second group of techniques aims to select products free of toxins after applying specific processing procedures. Relay has been used as a method to facilitate detoxification with limited success, but it is effective in preventing reintoxication (Blanco et al., 1997, 1999). Methods involving temperature, salinity (Gilfillan et al., 1976, Blogoslawski et al., 1979), electric shock (Kodama et al., 1989), pH (Neal, 1967), or chlorination have shown positive results in shellfish detoxification efforts. Numerous studies have been conducted to assess the potential of ozonation to enhance detoxification of shellfish. The results appear different according to the toxins and the shellfish species; Fernandez and Shumway (2004) highlighted the cost-ineffectiveness, as well as the limited safety of using these methods in the majority of the cases. Thus, the best way to obtain shellfish free of toxin is to monitor the water and the shellfish before harvesting and to only harvest non-toxic shellfish. The second group of techniques includes a very wide range of methods, from selective evisceration or selection of certain tissues to more highly developed industrial processes. Some species accumulate toxins in particular organs and not in every tissue, thus, selective removal of the toxic tissue, if the animal is large enough, can be an effective process. For example, bay scallops accumulate most brevetoxins in the hepatopancreas, but as most people eat only the adductor muscle, the scallops are usually safe to eat. These techniques are very tightly controlled, but have the advantage of allowing the sale of shellfish harvested from HAB-affected areas, which helps mitigate the economic losses associated with a toxic outbreak. Cooking the shellfish is another method used to decrease the levels of some toxins in the meats. As many toxins are water-soluble and very heat stable; however, they generally survive regular cooking. For example, cooking food containing PSP toxin for 5 minutes will reduce the toxicity by denaturation by only 30%; 20 minutes cooking by only 40% (Scoging, 1998). Similarly DSP toxins and brevetoxins are also heat-stable and overcome standard cooking, and can survive temperatures up to 300 ëC for brevetoxins. As the denaturation of toxins with heat is only slight, consumption of shellfish is allowed only when initial levels of toxins are very low. Depuration and ozonation are not effective in reducing PSP toxin level in shellfish either, and thus are not used (Anderson et al., 2001).
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Table 2.2
Monitoring of toxins and regulatory tolerance established by FDA (1998)
Toxin
Toxicity
Regulatory tolerance
Method of analysis
PSP
PD: 0.1±2 mg; LD: 0.3±12 mg 35±40 g PD: 42±72 MU PD: 1±5 mg/kg
80 g/100 g tissue
Mouse assay
0±60 g/100 g tissue 0.8 ppm (20 MU/100 g) 20 ppm domoic acid
Mouse assay Mouse assay HPLC
DSP NSP ASP
PD Pathogenic dose for humans. LD Lethal dose for humans.
Regulations, especially for the canning industry, now accept very low legal levels of toxins in shellfish (Table 2.2), but slightly contaminated shellfish must undergo a series of sequential procedures. Shellfish must be cleaned with freshwater for a minimum of 2 minutes at 20 2 ëC, followed by pre-cooking for 3 minutes at 95 5 ëC in freshwater. The flesh and the shell are then separated, and a second cleaning with freshwater for at least 30 seconds at 20 2 ëC precedes cooking for a minimum of 9 minutes at 98 3 ëC in freshwater. The shellfish flesh then must be cooled for 90 seconds under running, cold, freshwater before mechanical separation of the edible and nonedible parts with water pressure. The shellfish meat is then conditioned in hermetic containers in a non-acidified liquid medium and sterilized in an autoclave at 116 5 ëC for more than 15 minutes. In the previous paragraph, approved detoxification methods were presented, but as it is almost impossible to reliably reduce the amounts of biotoxins present in shellfish, the best technique to avoid human consumption of toxins in molluscan shellfish is to measure the presence of toxins and limit the harvest of highly toxic shellfish by controlling and classifying shellfish harvesting areas through monitoring programs. Additionally, the development of a system verifying the type and quantity of the shellfish after harvest, the harvester, the date and location of harvest and developed by the US FDA (1998) is necessary to verify that shellfish sold for human consumption were harvested from areas not impacted by HABs. 2.5.4 Regulation of shellfish safety and quality: international policies and harmonization Harmonization of regulations is needed as shellfish can be traded internationally. The General Agreement on Tariffs and Trade (GATT), included in the World Trade Organization (WTO) composed of 125 countries, represents an agreement on sanitary and phytosanitary regulations to protect health, which does not discriminate among all the countries' members and where the decisions are based on scientific evidence and risk assessment. The EU allows free trade among the different countries, and over the past
Biotoxin contamination and shellfish safety
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decade, numerous Directives concerning food-safety legislation have been developed to ensure high-quality products, consumer protection, and fair competition. In the case of molluscan shellfish, directives are being developed to advance the process of harmonization, involving traceability of the products, and controlling shellfish production areas, harvesting methods, transport, depuration, storage, processing, and marketing, but also chemical and microbiological parameters as well as toxin content. Monitoring of water and bivalve molluscs in production areas occurs periodically in numerous areas to detect the presence of toxins in the shellfish. However, much effort still needs to be made to develop harmonization of the sampling, methods of measurement, and the levels of toxins accepted. APEC (Asia Pacific Economic Cooperation) includes 18 coastal and archipelago economies very much involved in marine resources. One of the goals is to eliminate barriers between the countries and develop a free trade area; APEC is therefore working toward standardization of sanitary controls. HABs occur on a regular basis in these countries and have resulted in major economic losses. The need for monitoring programs is recognized in these countries, but they differ according to region. Some countries have very highly developed monitoring programs and are very well equipped for analyses; whereas, others have not developed programs to control the presence of toxins in marketed shellfish. The need for harmonization in monitoring methods for toxins is urgent; therefore, APEC created a Red Tide/Toxic Algae Project, developed by the Working Group on Marine Resources Conservation, to establish standards and legislation for all countries belonging to APEC. In the US, the Food and Drug Administration (FDA) is responsible for all shipping of food items interstate. In 1925, the FDA received authorization to enlist the efforts of each individual state, which led to the creation of the National Shellfish Sanitation Program (NSSP), In 1982, the Interstate Shellfish Sanitation Conference (ISSC), composed of members of the state and federal agencies, of shellfish industry and academic institutions, was created. It provides updated regulations for sanitation processes, harvesting, processing and shipping of shellfish. After monitoring of microalgal blooms in the water and monitoring of toxin levels in shellfish exposed to toxic microalgae, the decision to close an area to harvesting is made. Indeed, based upon the level of toxin content of the shellfish, water areas are classified, and shellfish harvest can be closed. Moreover, knowledge of the provenance, quality, date of harvest, etc., of the shellfish is also required. For example, FDA (1998) developed a control program, including a requirement for each container of in-shell, molluscan shellfish to bear a tag identifying the quality and type of shellfish, as well as the location and date of the harvest. Moreover, the harvesters also must be licenced. When the molluscan shellfish is shucked, it has to be certified and has to bear a label of the processor's name, address, and certification number.
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2.6
Shellfish safety and quality
Economic impacts of harmful algal blooms (HABs)
The economic impacts of HABs have been classified under four distinct categories: (1) public health impacts, (2) commercial fisheries impacts, (3) recreation and tourism impacts, and (4) monitoring and management costs. The economic costs associated with public health impacts are represented by the costs involved in medical treatment and investigation, but also attributable to loss of income and work days (Hoagland et al., 2002). This cost can be minimized by better information-dissemination and effective shellfish closure. In countries with efficient monitoring programs, the cost of HABs on public health impacts remains fairly low. Todd (1995) estimated the cost of a PSP illness at $1400 per reported case and $1100 per unreported case, including the cost of medical treatment and transportation (for the reported cases), as well as the cost associated with investigation following illness and the costs of lost productivity during sick days. Following the guidelines developed by Todd (1989), Hoagland et al. (2002) estimated the annual cost of public health attributable to shellfish poisoning in the US at $400 000 per year. The impacts of HABs on commercial fisheries are diverse. HABs can cause direct fish or shellfish mortalities, but can also lead to loss of habitat for certain species, can generate a forced closure of the fisheries, increase the cost of processing the contaminated harvested resources, and depress consumer demand (Hoagland et al., 2002). In the US, annual impacts of HAB on commercial fisheries vary between $7 to $19 million, averaging about $12 million per year. HABs have also demonstrated economic impacts on tourism and recreational activities (Hoagland et al., 2002). Examples of the economic impacts of HABs on recreation are very diverse, including: accumulation of dead fish on the coast after a HAB event, noxious odors produced on the beach by the death of the algae or of the dead marine organisms, water discoloration, closure of recreational fishing, mortalities of protected marine species, and air-quality impacts caused by aerosols that can be produced by the harmful algae and affect human health. In the US, between 1987 and 1992, annual impacts of HABs on tourism and recreational activities varied between $0 and $29 million, averaging about $7 million per year (Hoagland et al., 2002). The last main economic impact of HABs is represented by the monitoring and management costs of HABs (Hoagland et al., 2002). The cost involved in management and monitoring of the HAB varies considerably, according to country or even areas within one country. Many countries do not have a regular monitoring program; others, such as the US, include variation among the different states. Moreover, monitoring programs and regulations are different according to the countries and regions of the world (see previous section on monitoring and management). The costs are represented by the daily, weekly, or monthly monitoring, but also surveys or investigations taking place after a onetime HAB event. The real costs associated with HABs, are, however, very difficult to estimate properly, and more effort to quantify economic effects and to standardize the collection of data will be necessary to be able to better evaluate these costs (Hoagland et al., 2002).
Biotoxin contamination and shellfish safety
2.7
67
Conclusions
2.7.1 Other marine poisonings Even though not covered in this chapter, it is important to note that other species of animals, such as horseshoe crabs, snails, puffer fish, and other fishes can also be responsible for human poisonings. Indeed, other marine toxins have been identified, such as ciguatoxin (CTX), tetrodotoxin (TTX), or scombrotoxin. Human ciguatera poisoning, caused by CTX, scombroid poisoning (also called histamine poisoning), caused by scombrotoxin, and poisoning with TTX have occurred in humans only following ingestion of finfish. To date, no occurrences of these toxins have been recorded in shellfish; therefore, these toxins and their effects are not covered in this chapter. Moreover, this chapter was limited to shellfish poisoning, but another pathway for biotoxin effects upon humans can be through aerosols produced by the phytoplankton or by means of direct contact with the skin or human tissues following drinking or swimming. Cylindrospermopsin was, for example, responsible for human hepatoenteritis after drinking of contaminated, domestic water (Bourke et al., 1983; Hawkins et al., 1985). Karenia brevis blooms producing brevetoxins can also affect human health through inhalation of toxic aerosols (Music et al., 1975). 2.7.2 Impact of HAB on bivalves This review presents the effect of the major biotoxins on human health, and does not report the effect of these toxins on marine organisms. Effects of toxic phytoplankton and biotoxins have been reported on marine organisms (reviews by Shumway, 1990, and Landsberg, 2002), but these studies are very sporadic, and the impacts of biotoxin accumulation on the shellfish themselves are very poorly documented. Indeed, as the shellfish are toxic, they are not harvested, and thus not observed. Therefore, even though harmful algae producing toxins can also have deleterious effects on the shellfish themselves, effects upon commercial quality, usually the level of toxin at which these effects occur is higher than the level required for closure of the fishery. Shellfish toxins are predominantly a human health problem rather than a shellfish quality issue.
2.8
Future trends
The research community has been pursuing innovations in toxin detection, although most new methods have not been implemented in established regulatory programs. Three basic technological approaches have been pursued: chromatographic separation and detection of toxins, molecular-biological identification of toxin-producing microalgae, and `functional assays' for detection of toxins (Rossini, 2005; Hess et al., 2006). Comparisons of traditional and experimental toxin-detection protocols have yielded enough variability between methods to indicate caution in application of new methods (Llewellyn
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Shellfish safety and quality
et al., 2001). HPLC has long been the method of choice for quantifying many microalgal toxins in research settings (Quilliam, 2004). The potential, however, for interfering compounds to result in false positive results and unnecessary closures of shellfishing, or the worse scenario of ineffective extraction or analysis allowing sale of contaminated shellfish, has limited HPLC methods to follow-up analyses after positive bioassay results. Molecular detection of toxinproducing microalgae in water samples (Tengs et al., 2001) can lessen the need for highly trained taxonomists in monitoring programs and can aid in quantification of target algal species if coupled with quantitative fluorescence detection. Again, the possibility of not detecting an unexpected, toxic species limits the practical application of this technology to routine monitoring programs. Finally, perhaps the most promising technological advance in toxin detection is the `functional assay,' or affinity binding approach to detection of biotoxins in shellfish meats. Surface-adsorbed compounds that selectively bind toxin molecules, producing a color reaction can be packaged in commercial units that resemble home pregnancy test kits (Turrell et al., 2007). Such test units have been implemented in two ways. First, fisherman can use these units to test shellfish harvested at sea; if the first catch is contaminated, harvest is voluntarily halted. Nevertheless, any shellfish that are harvested still must be cleared for sale with bioassay monitoring. Second, in some very rural areas where monitoring methods are not implemented and artisanal harvest and consumption of shellfish occur, these test units can be distributed to consumers with the provision that consumption of shellfish still is not `approved', but at least the health of the consumer is somewhat protected. In addition to investigations into new methods for toxin detection, innovations have also been pursued in satellite imaging to forecast movements of water masses containing HABs (Schofield et al., 1999). The ability to anticipate interactions between advection of blooms into areas with shellfish resources would enable more informed decisions about harvest, relay, and closure as management options. Implementation of new technologies for HAB detection and forecasting may add another level of sophistication, along with complexity, to monitoring and control programs. Nevertheless, ultimately, protection of human consumers of shellfish from illnesses caused by microalgal biotoxins remains of paramount importance. Further, changes in, or additions to, legal documents defining responsibilities and approved methods for shellfish harvest and marketing in the face of expanding HAB effects will require thorough documentation of both technical and legal considerations. Accordingly, the process moving innovations from scientific discovery to practice will be methodical and thorough.
2.9
References and further reading (2003) Application of an improved method for detection of lipophilic marine algal toxins (OA/DTXs, PTXs, YTXs, AZAs) with LC/MS. In:
AASEN J, TORGENSEN T, AUNE T
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acid, an amnesic shellfish toxin. In: Botana LM (ed) Seafood and Freshwater Toxins: Pharmacology, Physiology, and Detection. Marcel Dekker, New York, pp 325±358. NOGUCHI T, JEON J, ARAKAWA O, SUGITA H, DEGUCHI T, SHIDA T, HASHIMOTO K (1986) Occurrence of tetrodotoxin and anhydrotetrodotoxin in Vibrio sp. isolated from the intestine of a xanthid crab, Atergatis floridus. J. Biochem. 99: 311±314. NOVACZEK I, MADHYASTHA MS, ABLETT RF, DONALD A, JOHNSON G, NIJAR MS, SIMS DE (1992) Depuration of domoic acid from live blue mussels (Mytilus edulis). Can. J. Fish. Aquat. Sci. 49: 312±318. NOVELLI A, KISPERT J, REILLY A, ZITKO V (1990) Excitatory amino acids toxicity on cerebellar granule cells in primary culture. Canad. Dis. Wkly. Rep.S1E 16: 83±89. OFUJI K, SATAKE M, MCMAHON T, SILKE J, JAMES KJ, NAOKI H, OSHIMA Y, YASUMOTO T (1999a) Two analogs of azaspiracid isolated from mussels, Mytilus edulis, involved in human intoxication in Ireland. Natural Toxins 7: 99±102. OFUJI K, SATAKE M, OSHIMA Y, MCMAHON T, JAMES KJ, YASUMOTO T (1999b) A sensitive and specific determination method for azaspiracids by liquid chromatography mass spectrometry. Natural Toxins 7: 247±250. OFUJI K, SATAKE M, MCMAHON T, JAMES KJ, NAOKI H, OSHIMA Y, YASUMOTO T (2001) Structures of azaspiracid analogs, azaspiracid-4 and azaspiracid-5, causative toxins of azaspiracid poisoning in Europe. Biosci. Biotechnol. Biochem. 65: 740±742. OSHIMA Y, MACHIDA M, SASAKI K, TAMAOKI Y, YASUMOTO T (1984) Liquid chromatographic fluorometric analysis of paralytic shellfish toxins. Agric. Biol. Chem. 48: 1707± 1711. PERL TM, BEDARD L, KOSATSKY T, HOCKIN JC, TODD ECD, REMIS RS (1990) An outbreak of toxic encephalopathy caused by eating mussels contaminated with domoic acid. New Engl. J. Med. 322: 1775±1780. PINTO JS, SILVA ES (1956) The toxicity of Cardium edulum L. and its possible relation to the dinoflagellate Prorocentrum micans. Ehr. Notas Estud. Inst. Biol. Marit. 12: 22. POLI MA, MENDE TJ, BADEN DG (1986) Brevetoxin, unique activators of voltage-sensitive sodium channels, bind to specific sites in rat brain synaptosomes. Mol. Pharmacol. 30: 129±135. POLI MA, MUSSER SM, DICKEY RW, EILERS PP, HALL S (2000) Neurotoxic shellfish poisoning and brevetoxin metabolites: a case study from Florida. Toxicon 38: 981±993. PRAKASH A, MEDCOF JC, TENNANT AD (1971) Paralytic shellfish poisoning in Eastern Canada. Fish. Res. Bd. Can. Bull. 177. QUAYLE DB (1969) Paralytic shellfish poisoning in British Colombia. Bull. Fish. Res. Bd. Can. 168: 18. QUILLIAM MA (2004) Chemical methods for domoic acid, the amnesic shellfish poisoning (ASP) toxin. In: Hallegraeff GM, Anderson DM, Cembella AD (eds) Manual on Harmful Marine Microalgae. UNESCO, Paris QUILLIAM MA, WRIGHT JLC (1995) Methods for diarrhetic shellfish poisons. In: Hallegraeff GM, Anderson DM, Cembella A (eds) Manual on Harmful Marine Microalgae. UNESCO, Paris, pp 95±111. QUILLIAM MA, THOMAS K, WRIGHT JLC (1998) Analysis of domoic acid in shellfish by thinlayer chromatography. Natural Toxins 6: 147±152. ROBERTS BS, HENDERSON GE, MEDLYN RA (1979) The effect of Gymnodinium breve toxin (s) on selected mollusks and crustaceans. In: Taylor DL, Seliger HH (eds) Toxic Dinoflagellate Blooms. Elsevier, New York, pp 419±424.
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Part II Improving molluscan shellfish safety and quality
3 Viral contaminants of molluscan shellfish: detection and characterisation A. Bosch and R. M. PintoÂ, University of Barcelona, Spain and F. S. Le Guyader, Laboratoire de Microbiologie, France
Abstract: Environmental virology started with the detection of poliovirus in water. Since then other enteric viruses responsible for gastroenteritis and hepatitis have replaced enteroviruses as the main target for detection. Most shellfish-borne viral outbreaks are restricted to norovirus and hepatitis A virus, making them the main targets for bivalve virological analysis. The inclusion of virus analysis in regulatory standards for viruses in molluscan bivalve samples must overcome several shortcomings such as the technical difficulties and high costs of virus monitoring, the lack of harmonised and standardised assays and the challenge posed by the ever-changing nature of viruses. Nowadays methods are available to detect, quantify and characterise viral pathogens in molluscan shellfish to reduce the risks of shellfish-borne virus diseases. Key words: norovirus, hepatitis A virus, enteric viruses, bivalve molluscs, shellfish, virus analysis.
3.1 Introduction: human enteric viruses and their fate in the environment A wide variety of different viruses, representing the majority of the families of animal viruses, can be present in human and animal faecal wastes and urine. Especially important are a variety of non-enveloped human and animal enteric pathogenic viruses that can enter the environment through the discharge of waste materials from infected individuals. Viruses can contaminate food products, and drinking and recreational waters, and be transmitted back to susceptible individuals to continue the cycle of infection. These enteric viruses cause a wide
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spectrum of illnesses in humans including hepatitis, gastroenteritis, meningitis, fever, rash, conjunctivitis, and maybe diabetes or severe acute respiratory syndrome (SARS) (Table 3.1). It is estimated that billions of cases of gastrointestinal illness occur annually worldwide (Parashar et al., 1998; Oh et al., 2003) and a good deal of these diarrhoeal cases is to some extent the result of faecal contamination of the environment (Cabelli et al., 1983; Koopman et al., 1982; Moore et al., 1994; Pommepuy et al., 2005). In 1979, it was estimated that between 5 and 18 million people die every year from gastroenteritis. In developing countries the burden of rotavirus disease in children under 5 years of age has been estimated to be over 125 million cases annually, of which 18 million are severe cases, and nearly half a million deaths in children under the age of 4 are attributable to rotavirus diarrhoea (Parashar et al., 1998; Oh et al., 2003). In the developed world, mortality due to rotavirus infection is very low; however, it remains an important cause of morbidity and of hospitalisation in young children. Noroviruses (previously called Norwalk-like viruses), formerly included in the SRSV (small round structured viruses), account for over 90% of foodborne gastroenteritis affecting children and adults (Lopman et al., 2003; Blanton et al., 2006). Additionally, astroviruses were reported in 1996 to be second only to rotaviruses as a cause of hospitalisation for childhood viral gastroenteritis (Glass et al., 1996), while adenoviruses and sapoviruses (previously called Sapporo-like viruses) have also been recognised as significant Table 3.1
Human enteric viruses with potential environmental transmission
Genus
Popular name
Disease caused
Enterovirus
Polio Coxsackie A, B
Paralysis, meningitis, fever Herpangina, meningitis, fever, respiratory disease, hand-and-foot-and-mouth disease, myocarditis, heart anomalies, rash, pleurodynia, diabetes? Meningitis, fever, respiratory disease, rash, gastroenteritis Gastroenteritis Hepatitis Unknown Gastroenteritis Gastroenteritis Gastroenteritis Hepatitis Gastroenteritis Gastroenteritis Gastroenteritis, respiratory disease Gastroenteriris Gastroenteritis, respiratory disease, conjunctivitis Progressive multifocal leukoencephalopathy Nephropathy
Echo Kobuvirus Hepatovirus Reovirus Rotavirus Norovirus Sapovirus Hepevirus Mamastrovirus Parvovirus Coronavirus Torovirus Mastadenovirus
Aichi Hepatitis A Human reovirus Human rotavirus Norwalk-like virus Sapporo-like virus Hepatitis E Human astrovirus Human parvovirus Human coronavirus Human torovirus Human adenovirus
Polyomavirus
JCV KV
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aetiological agents of epidemic non-bacterial diarrhoea (Lopman et al., 2003). Poliomyelitis, caused by a picornavirus, was not too long ago the most feared viral disease; however, the long-pursued objective of its eradication seems presently within reach. Another picornavirus is the aetiological agent of hepatitis A which approximately accounts for half the total number of cases of hepatitis diagnosed worldwide, and some regions, as part of the Mediterranean region, are still endemic for hepatitis A (Pinto et al., 2007). Hepatitis E virus is the primary cause in tropical and subtropical developing countries of an enterically transmitted non-A non-B hepatitis, with a mortality rate of up to 20% in pregnant women (Reyes, 1993; Schlauder and Mushahwar, 2001; Lu et al. 2006). Environmental virology may be defined as the study of the extracellular behaviour of viruses which can be transmitted through various environments (water, sewage, soil, air or surfaces) or food, and persist enough in these vehicles to represent a health threat. As a scientific discipline, environmental virology was born after a large hepatitis outbreak occurred in New Delhi between December 1955 and January 1956. The origin of the outbreak, which was attributed to hepatitis A at the time but now confirmed to be hepatitis E, was the contamination by sewage, from 1 to 6 weeks prior to the epidemic, of Jumna River, the source of water for the treatment plant. Alum and chlorine treatment prevented bacterial infections, but 30 000 cases of hepatitis occurred among the population. Enteric viruses can be transmitted by a variety of routes, including direct and indirect contact, vector transmission and vehicle transmission. Viruses are shed in extremely high numbers in the faeces of infected individuals; patients suffering from diarrhoea or hepatitis may excrete from 105 to 1011 virus particles per gram of stool (Farthing, 1989; Kageyama et al., 2003). Furthermore, a single episode of vomit of a patient with norovirus gastroenteritis may contain around 107 particles (Cheesbrough et al., 1997). Ingestion of sewage-contaminated water or food is the main route of infection with human enteric viruses, although the role of inanimate surfaces serving as vehicles for virus infection must not be underestimated. Viruses with a viraemic phase, such as the hepatitis viruses, may also be parenterally transmitted, although nowadays it is considered to be a much less frequent mode of transmission.
3.2
Shellfish-borne transmission of virus infections
Despite indigenous marine virus strains outnumbering any form of life in the sea, usually occurring in billion amounts per litre (Fuhrman, 1999; Danovaro et al., 2001), the only viral agents of public health concern in the marine environment are human viruses. Pathogenic viruses are routinely introduced into marine and estuarine waters through the discharge of treated and untreated sewage, since current water treatment practices are unable to provide virus-free wastewater effluents (Rao and Melnick, 1986).
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The demands exerted by the expanding world population and industry make the marine environment increasingly susceptible to pollution from municipal sewage, industrial effluents and agricultural wastes. Seawater pollution control relies on secondary treatment of sewage and on the theoretically infinite dilution of wastes in the receiving waters; however, the marine environment has a finite ability to receive and recover from waste disposal practices, and certainly is incapable of unlimited waste assimilation. The type of treatment applied to human waste will ultimately determine the concentration of pathogens in treated sewage and sludge, and their relative risk of disposal. The maintenance and assessment of the virological quality and safety of marine water systems employed for recreating and seafood harvesting are of seminal importance in the prevention of diseases transmitted through the faecal±oral route, and may lead to significant reductions of economic losses due to the closures of tourist resorts and shellfish harvesting areas. For this reason, it is imperative to trace and characterise the type and origin of faecal contaminants in order to assess the associated health threat and the required corrective measures. There are several routes by which viruses reach the sea, including direct discharge of treated or untreated sewage effluents, unintentional discharges by urban and rural runoff, waste input from boats, and via rivers when the discharges take place in freshwater. Humankind is exposed to enteric viruses in seawater mainly through the consumption of shellfish grown in contaminated waters, or to a lesser extent through recreational activities in sewage-polluted waters. Several phenomena, such as flooding, treated and untreated polluted effluent discharges, or sewage runoff can elevate microbial contaminants in shellfish habitats, and, since bivalves are filter feeders, these molluscs can become reservoirs of human pathogens. Other types of seafood such as crabs (Goyal, 1984) or shrimps (Botero et al., 1996) can accumulate viruses on their shells and carnivorous shellfish, such as lobsters or crabs, can feed on contaminated bivalves (Hejkal and Gerba, 1981), but their role in the transmission of viral diseases is unproven. Following the culinary tradition, bivalve shellfish are often consumed raw, like oysters and sometimes clams or cockles, or just lightly cooked, like most other molluscs. This cooking habit, together with the fact that the whole animal including viscera is consumed, poses a major public health concern since shellfish serve as passive carriers of human pathogenic viruses. Generally, commercial growth of shellfish species takes place in shallow, in-shore waters, which may receive occasional sewage pollution. The consumption of shellfish is very clearly linked to the transmission of enteric infections, and epidemics have been recorded since medieval times in many countries (Lees, 2000). Although most of these outbreaks are caused by shellfish collected by unscrupulous professionals or careless private individuals from areas where harvesting is prohibited, they also occur as a result of eating shellfish from authorised shellfish-producing areas, when these areas have been temporarily polluted (Mackowiak et al., 1976) and sanitary controls fail to provide a safe indication of viral pollution.
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Table 3.2 Examples of reported large (over 100 cases) virus outbreaks linked to shellfish consumption Year
Country
Shellfish
No. of cases
Agent
Reference
1973 1976±77
US Great Britain Australia Australia
Oysters Clams
265 800
HAV* SRSVy
Oysters Oysters
2000 150
NoVz NoV
Cockles
424
NoV
Oysters Oysters
472 181
NoV SRSV
Richards (1985) Gill et al. (1983)
1983 1986
Great Britain US Great Britain Malaysia US
Mackowiak et al. (1976) Appleton and Pereira (1977) Murphy et al. (1979) Linco and Grohmann (1980) O'Mahony et al. (1983)
Goh et al. (1984) Morse et al. (1986)
Shanghai Spain
322 813 204 292 301 183
HAV NoV
1988 1999
Cockles Clams Oysters Clams Clams
HAV HAV
Halliday et al. (1991) Bosch et al. (2001)
1978 1978 1980±81 1982 1983
* HAV: hepatitis A virus. y SRSV: small round structured viruses. z NoV: norovirus.
Regardless of the variety of health-significant viruses found in shellfish, norovirus and hepatitis A virus are the most relevant viral pathogens involved in shellfish-borne diseases (Table 3.2). Norovirus infections represent the vast majority of shellfish-related outbreaks, and hepatitis A is the most serious infectious disease caused by shellfish consumption. The US Food and Drug Administration (FDA) risk assessments estimate cases of norovirus gastroenteritis related to seafood consumption at some 100 000 per year (Williams and Zorn, 1997), and epidemics of hepatitis A caused by food occur 10 times more often than those caused by water, shellfish being the cause of more than 50% of reported cases (Cliver, 1985). The first reported association of viruses with shellfish-borne gastroenteritis infection was observed in the winter of 1976±77 in the UK when cockles were epidemiologically linked to 33 incidents affecting nearly 800 people (Appleton and Pereira, 1977). SRSV particles, like those seen in outbreaks of winter vomiting disease, were observed by electron microscopy in a high proportion of patient faeces. Nevertheless, no shellfish-borne outbreak ever had the magnitude of the one reported in Shanghai in 1988, caused by hairy clams (Halliday et al., 1991). Bioaccumulation of viruses in the shellfish digestive tract is a very rapid phenomenon. Viruses are readily adsorbed to shellfish tissue within an hour of contact, and maximum virus levels may be observed after 6 hours (Abad et al., 1997). Adsorption of viruses onto substrates such as faeces, kaolinite or unicellular algae, considerably increases shellfish accumulation efficiency (Metcalf et al., 1979).
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Commercial heat treatment (cooking) is employed to reduce the levels of microbial contaminants in shellfish. Heat can render many viruses noninfectious; however, the degree of cooking required to reliably inactivate viruses would probably render oysters unpalatable to consumers (McDonnell et al., 1997). Laboratory studies show that enteric viruses and notably hepatitis A virus may be found in cooked shellfish (Abad et al., 1997). In addition, outbreaks of gastroenteritis and hepatitis have been linked to consumption of commercially cooked cockles or oysters (Appleton and Pereira, 1977; Kohn et al., 1995). High-pressure treatment was recently shown efficient at inactivating HAV within oyster tissues, suggesting that this technology may be useful for sporadically contaminated shellfish (Calci et al., 2005).
3.3 Effects of viral contamination of molluscs on the international shellfish industry Most countries have endorsed sanitary controls on live bivalve shellfish. In the EU, these are covered by Council Directive 91/492/EEC (Anonymous, 1991) and in the US, by interstate trading agreements set out in the FDA National Shellfish Sanitation Program Manual of Operations (Anonymous, 1993). These regulations cover similar ground on the requirements, among others, for harvesting area classification, depuration, relaying, analytical methods and provisions for suspension of harvesting from classified areas following a pollution or public health emergency. A major weakness of these controls is the use of traditional bacterial indicators of faecal contamination, such as the faecal coliforms or Escherichia coli, to assess contamination and hence implement the appropriate control measures. Faecal indicators are either measured in the shellfish themselves (EU perspective) or in the shellfish growing waters (US FDA perspective). Several reports, however, describe a lack of correlation between bacterial indicator microorganisms and viruses, and pathogenic viruses may be detected in shellfish from areas classified as suitable for commercial exploitation according faecal coliform criteria (Abad et al., 1997; Lees, 2000; Le Guyader et al., 2000). The guidelines establish that shellfish meeting a microbiological standard of less than 230 E. coli or 300 faecal coliforms in 100 g of shellfish flesh can be placed on the market for human consumption. Human enteric viruses, e.g. norovirus, rotavirus and hepatitis A virus, have been detected in shellfish which were adequate for public consumption according criteria based on the numbers of bacterial indicators (Jofre et al., 1993; Bosch et al., 1994; Le Guyader et al., 2000; Romalde et al., 2002). Additionally, hepatitis A and gastroenteritis outbreaks have been associated with the consumption of shellfish meeting legal standards (Bosch et al., 2001; Le Guyader et al., 1996b; 2003; Lees, 2000; Mele et al., 1989, Boxman et al., 2006). The legislation also requires that third country imports into the EU and US have to meet the same standard as domestic products. Exporting nations have therefore developed programmes for compliance with the regulations of their
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target export markets. Nevertheless, a number of examples of trans-national outbreaks have recently been reported following trade between EU Member States (Christensen et al., 1998) and importation of shellfish from third countries into the EU and the US (Bosch et al., 2001; SaÂnchez et al., 2002; Kingsley and Richards, 2003). European shellfish trade turnover represents M¨460 per year, and increases by around 7% each year. The European production represents more than a third of the worldwide shellfish production (i.e. in 1991, 180 000 tonnes of live weight: 72% of farmed bivalves, 28% wild; Eurostat data) and 8500 companies currently employ around 23 000 workers. This activity is one of the major sources of employment in coastal areas of such countries as Ireland, France, Spain or the Netherlands. Important trade exchanges of shellfish take place worldwide. For example in the EU, France, the main oyster producer (140 000 t/year) and consumer of the EU, also exports oysters mainly to Belgium and imports oysters from Ireland and the Netherlands (about 1500 t in 1994). Sixty per cent of the 1.1 million tonnes of mussels harvested in the world each year are produced by the EU, Spain being the second largest producer worldwide after China. France produces fewer mussels than it needs for its consumption: 20±40% of the mussels are imported from other countries (Spain, the Netherlands, UK). Shellfish contamination has also an impact on the quality of life. Local population and tourists appreciate the nice quality of life in coastal areas. Coastal tourism has a high economic impact and major local implications, especially for employment. This tourism, if primarily interested in recreative activities, is also aware of and concerned by the environment and landscape preservation. Shellfish harvesting activities, if well managed and regulated, could also contribute to better protecting the environment. People would practise these activities in areas where the water is of recognised quality and the environment most protected and closest to wilderness. Thus, sustainable management of aquaculture would not only have direct and indirect incidences on employment (e.g., trade activities, tourism employment, equipment, hotel, beach activities, rentals) but also on the protection of rural and coastal development.
3.4 Methods for detecting viruses in molluscan shellfish and associated problems Virus detection in shellfish has to overcome several difficulties. On the one hand, viruses are expected to be present in shellfish in very low numbers, which nevertheless are sufficient to pose a health risk. This low virus load implies the use of methodologies yielding a high efficiency of virus recovery from shellfish tissues. On the other hand, shellfish extracts are both highly cytotoxic and not adequate to be inoculated in cell cultures for the detection of culturable viruses, and not compatible either with polymerase chain reaction (PCR) based
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methodologies for the detection of non-culturable viruses, particularly if a reverse transcription must be previously performed (RT-PCR). The key objective is then to develop procedures for shellfish analysis which result in a low volume of a non-cytotoxic or, even better nowadays, highly pure nucleic acid preparation with no inhibitory effect to the PCR. As a matter of fact, in this latter case, the degree of virus detection effectiveness achieved after RT-PCR is essentially the result of two related factors: the efficiency of recovery of the extraction procedure applied to the shellfish sample and the degree of final purity of the recovered virus. Table 3.3 lists different procedures for the processing of shellfish samples prior to the detection of specific viruses by molecular procedures since the most relevant shellfish-borne viral pathogens are non-culturable. The first decision is to choose between performing virus detection in dissected shellfish tissues or in whole shellfish meats. Studies on the localisation of human enteric viruses in shellfish tissues have revealed that most of the viruses are found in the stomach and digestive diverticula (Romalde et al., 1994; Abad et al., 1997). Atmar and co-workers reasoned that removal of these organs for virus extraction might simplify and shorten the time needed to purify viral nucleic acid for RT-PCR but also improve the sensitivity by increasing the number of individuals analysed (Atmar et al., 1996). Testing the stomach and digestive gland for virus detection presented several advantages in comparison with testing whole shellfish: less time-consuming procedure, increased test sensitivity, and decrease in the sample-associated interference with RT-PCR. Following virus extraction, a variety of subsequent nucleic acid extraction and purification protocols may be employed (Table 3.3). Owing to the small size Table 3.3 PCR
Procedures for processing of shellfish samples prior to virus detection by RT-
Process
Method
Reference
Virus extraction
Chloroform-butanol/cat-floc elution Vertrel extraction Proteinase K treatment
Atmar et al. (1996) Mendez et al. (2000) Jothikumar et al. (2005)
Virus concentration
Organic flocculation Centrifugation Ultracentrifugation PEG precipitation
Sobsey et al. (1978) Sobsey et al. (1978) Loisy et al. (2000) Lewis and Metcalf (1988); Atmar et al. (1995)
RNA extraction
Guanidium thiocyanate
Boom et al. (1990); Lees et al. (1994) Atmar et al. (1995); Jaykus et al. (1994) Loisy et al. (2000); Schwab et al. (2000); Shieh et al. (1999)
Cetyltrimethyl ammonium bromide (CTAB) Commercial nucleic acid extraction kits
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of the PCR reaction volumes, a reconcentration step is incorporated prior to the molecular assay. Nucleic acid purification based on virus lysis with guanidine and recovery with a silica matrix (Boom et al., 1990; Lees et al., 1994), or, alternatively, the use of organic solvents for purification, followed by nucleic acid precipitation using cetyltrimethyl ammonium bromide (CTAB) (Atmar et al., 1995; Jaykus et al., 1994), remain the procedures of choice. A wide variety of commercial kits have been applied for nucleic acid purification, offering reliability combined with convenience (Shieh et al., 1999; Loisy et al., 2000; Schwab et al., 2000). Molecular analysis of viruses in environmental samples involves problems with inhibitors, low virus concentrations, and sequence variation. As the concentration-extraction procedure is not virus specific, the nucleic acid of several viruses can be extracted at the same time. RT-PCR must be performed under stringent conditions and confirmed by hybridisation. Sometimes it is necessary to analyse the amplified sequence in order to characterise the viral strains. This is particularly important for norovirus detection, owing to its wide strain diversity. However, sequence analysis is hampered by the sometimes scarce product obtained after PCR amplification from shellfish tissues. In addition, the high genetic diversity of norovirus makes it necessary to use broadly reactive primers. Despite several improvements in the methodology, up to now no single primer set is able to amplify all strains (Atmar and Estes, 2001; Vinje et al., 2003). In the absence of such a universal primer set, multiple sets need to be used to detect all strains (Le Guyader et al., 1996a). The use of multiple primer sets enhances the chance to detect a greater number of strains, and the homology of the primers with the norovirus strain is important in terms of sensitivity (Le Guyader et al., 1996a; Atmar and Estes, 2001). No single assay stands out as the best by all criteria such as evaluation of sensitivity, detection limit, and assay format for stool analysis, and thus it is even more difficult for shellfish sample with very low contamination (Vinje et al., 2003). For hepatitis A virus, primer selection is easier since the degree of variation, particularly in the non-coding regions, is significantly lower (Costafreda et al., 2006). Not too long ago, methods for the detection of viral pathogens were restricted to assays for culturable viruses, focused almost entirely on enteroviruses, and the BGM cell line has long been the choice for infectivity assays of enteroviruses in environmental samples (Morris and Waite, 1980; Rao et al., 1986; Bosch, 1998). Despite the fact that enteroviruses do not appear as epidemiologically relevant environmental contaminants, it will remain important to gather data on their occurrence in the environment until the global eradication of poliomyelitis becomes a reality. However, even for this latter purpose, molecular tools provide better perspectives than cell culture techniques. Wild-type rotaviruses present difficulties in their in vitro replication, although most isolates may be adapted to grow in several cell lines such as the monkey kidney cell line MA104 or the human intestinal cell line CaCo-2 (Kitamoto et al., 1991). The standard methods for the diagnosis of specific infectious rotaviruses involve immunofluorescence tests and optical microscopic counting of infected foci in the culture (Smith and
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Gerba, 1982; Hejkal et al., 1984; Bosch et al., 1988). A further refinement in this direction was the use of flow cytometry for the detection of fluorescent foci in rotavirus infected cells (Abad et al., 1998). Flow cytometry is applicable for the detection of rotaviruses in environmental samples through an automatable and standardisable procedure that is much less cumbersome than direct optical microscopy screening of cell cultures for fluorescent foci. Another approach for the recovery of viruses that replicate poorly in cell cultures is to employ an integrated cell culture-reverse transcriptase-polymerase chain reaction system (CC-RT-PCR), enabling the in vivo amplification of virus sequences in cell culture prior to their detection by PCR, thus accomplishing the dual purpose of increasing the number of copies of target nucleic acid and of incorporating an infectivity assay as well (Ma et al., 1994; Pinto et al., 1995). This approach has been reported for detection of infectious astrovirus (Abad et al., 1997) and enterovirus (Reynolds et al., 1996; Murrin and Slade, 1997). Whenever possible, the use of a combined CC-RT-PCR procedure that utilises the major advantages of the separate methodologies, while overcoming many of their disadvantages is recommended. The inclusion of an infectivity test prior to the specific detection may contribute to solve the lack of sensitivity required for some type of samples such as environmental samples. However, so far the use of cell monolayers is of little use for the primary isolation of hepatitis A virus and unavailable for norovirus detection. The requirement of sophisticated facilities and well-trained personnel to conduct studies with enteric viruses and the unreliability of bacterial model microorganisms led to the search for alternatives. Several bacteriophage groups appear as promising candidates, among them somatic coliphages (IAWPRC Study Group on Health Related Water Microbiology, 1991), F+ specific (malespecific), RNA (FRNA) bacteriophages (Havelaar, 1993) and Bacteroides fragilis bacteriophages (Tartera and Jofre, 1987), all of them with available ISO (International Standardisation Office) procedures for their detection in water. FRNA phages in particular have been described as promising candidates to evaluate the virological quality of shellfish (Lees, 2000). Several studies have shown a correlation between the elimination kinetics of F+ RNA phages and those of enteric viruses (Power and Collins, 1989, 1990; Dore and Lees, 1995). Nevertheless reports on discrepancies in the occurrences of FRNA phages and pathogenic viruses are frequent. In shellfish associated with a large outbreak of hepatitis A reported in the East of Spain in 1999, with 184 serologically confirmed cases, the discrepancy observed between hepatitis A virus and FRNA phages was 55%, while a 50% discordance was ascertained between generic enteric virus occurrence and F+ presence (Bosch et al., 2003). In another study comparing the validity of E. coli, enterovirus and FRNA bacteriophages as indicator microorganisms, the phages failed to predict the risk of viral illness (Miossec et al., 2001). Additionally, when the comparative positivity for human enteric viruses and FRNA phages was investigated in 101 randomly chosen shellfish samples from South and West coast of France, a good correlation
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between the occurrence of enteric viruses and FRNA phages was observed in only 49% of the samples (Le Guyader, unpublished results). Phages that could be reliable indicators in some cases can be used to classify areas for sanitary safety, but cannot determine if a batch of shellfish is viruscontaminated (Hernroth et al., 2002). In the same type of cold seawater, a correlation was found between noroviruses and phage contamination of mussels, but more than half of the norovirus-positive samples were negative for FRNA phages. A positive FRNA phage result was less than twice as common in samples with norovirus than in those without norovirus, raising the question of use of FRNA phages as reliable indicators (Myrmel et al., 2004). A study conducted among different European countries showed geographic variations with shellfish collected from Southern Europe negative for FRNA contained human viruses (Formiga-Cruz et al., 2002). In Italy, most of shellfish found contaminated with hepatitis A virus did not present any phage or E. coli contamination (Croci et al., 2000). Recently a 1-year study in the Netherlands showed the presence of phages in 67% of oyster samples analysed, but without the presence of pathogenic viruses such as norovirus or hepatitis A virus (Lodder-Verschoor et al., 2005). Exhaustive studies are still required to ascertain the validity of a candidate indicator in a given scenario. A `universal' indicator for viruses, applicable to all situations, is probably unrealistic and the use of particular indicator, index, and model microorganisms for specific purposes is in order.
3.5 Improving detection of molluscan shellfish virus contamination using new molecular-based methods The advent of molecular techniques for virus detection, and particularly RTPCR, provided exquisite tools for the detection of fastidious health-significant viruses in food and environmental samples. Health-significant viruses, which were previously unrecognisable because they replicate poorly or not at all in cell cultures, became detectable with nucleic acid-based techniques. Virologists initially employed hybridisation assays which have been more recently replaced by polymerase chain reaction based procedures (Lees et al., 1994; Jothikumar et al., 1995; Bosch et al., 1996; Jaykus et al., 1996; Le Guyader et al., 1996a; Pinto et al., 1996; Schwab et al., 1998; Villena et al., 2003). Many potential users may find PCR cumbersome, since a single test entails many different manual steps, and will consider the technique as suitable only for academic or reference laboratories, and inadequate for routine monitoring. However, over the last decade, PCR technology has improved on several fronts. On the one hand, commercial PCR systems significantly ameliorated convenience, and have been quickly adopted for diagnostic laboratories. Nevertheless, the most dramatic improvement comes from the emergence of combined rapid thermocycling and fluorescence monitoring of amplified product, collectively referred as `rapid-cycling real-time PCR' (Cockerill and Smith, 2002;
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Gassilloud et al., 2003; Kageyama et al., 2003; Loisy et al., 2005a; Costafreda et al., 2006), together with nucleic acid sequence-based amplification or NASBA techniques (Jean et al., 2001; Yates et al., 2001), both of which are now applicable in several commercially available systems. These procedures enable not only qualitative determination but also, and particularly, quantitative diagnostic assays. Although the generic determination of pathogens is the essence of diagnostic practices, the possibility for quantitative detection of virus agents represents a seminal refinement in routine monitoring virology. As stated above, norovirus and hepatitis A virus are the two most significant virus targets in shellfish tissues, owing to their incidence and pathogenicity. For this reason, considerable attention has been dedicated to the development of real-time procedures for the detection of these agents in bivalve molluscan shellfish (Nishida et al. 2003; Jothikumar et al. 2005; Loisy et al., 2005a; Costafreda et al., 2006). However, methods cited in the literature are diverse, complex, poorly standardised and restricted to a few specialist laboratories. It is obvious that quality control and quality assurance issues must be solved, as well as simplification and automation of molecular procedures before they could be adopted by routine monitoring laboratories. An additional difficulty to solve in the detection of viruses in molluscs is that traditional shellfish extraction procedures are not always compatible with RT-PCR detection: inhibitory substances are concentrated and recovered along with the viruses. A great variety of procedures have been developed for the removal of inhibitors, which include dialysis, solvent extraction, proteinase treatments, lyophilisation, gel or glass filtration, nucleic acid adsorption or precipitation, antibody capture, and the use of commercial kits (Tsai et al., 1993; Atmar et al., 1995; Jaykus et al., 1996; Shieh et al., 1999; Loisy et al., 2000; Schwab et al., 2000). The rule of thumb is that the degree of final purity of the assayed sample greatly determines the sensitivity of PCR, or particularly, RT-PCR virus detection. Methodologies for the accurate quantification of norovirus and hepatitis A virus in shellfish samples are being developed. The general approach is based on the use of several controls to measure the efficiency of those critical steps for the quantification: the virus and nucleic acids extractions, and the RT-PCR reactions. The first purpose involves the use of a non-pathogenic virus of similar structural characteristics to those of the target virus. In the case of hepatitis A virus, since it belongs to the Picornaviridae family, another member of the same family is used to validate the behaviour of hepatitis A virus during its extraction from the shellfish tissue as well as during the nucleic acids extraction procedures (Costafreda et al., 2006). Encephalomiocarditis virus (EMCV) has been proposed as a model for hepatitis A virus in validation studies of hepatitis A virus removal in blood products manufacturing by several agencies such as the European Agency for the Evaluation of Medicinal products (http://www.emea.eu.int/pdfs/human/bwp/026995en.pdf) or the US FDA (http:// www.fda.gov/cber/sba/igivbax042705S.pdf). However, the use of this virus is hampered by its potential pathogenicity in several animals, including primates (Citino et al., 1988) including humans (Kirkland et al., 1989). Mengo virus is
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serologically indistinguishable from EMCV, and non-pathogenic for humans, although it may infect several other animals. The removal of the poly-C tract from the 50 NCR of the wild-type Mengo virus, gives rise to a mutant strain, i.e. Mengo virus vMC0, with the same growth and structural properties but with no pathogenic capacity (Martin et al., 1996). Mengo virus vMC0 is employed as an extraction control for hepatitis A virus (Costafreda et al., 2006), since it represents a phenotypic variant of Mengo virus, avirulent in all animal species (murine and non-murine) so far tested, and used as a vaccine for a wide variety of hosts, including baboons, macaques and domestic pigs (Osorio et al., 1996). The same Mengo virus vMC0 is at the time of writing this chapter validated within the framework of an EU committee (CEN TAG4) as an extraction control, not only for hepatitis A virus but also for other viruses, such as norovirus, in shellfish, and other food matrices as fruits and salads. It is well known that one limitation of molecular techniques is that they fail to discern between infectious and non-infectious particles which may be of critical relevance in environmental virology (Abad et al., 1994; Gassilloud et al., 2003). Several issues should, however, be taken into consideration. Most enteric viruses of public health concern bear RNA genomes. In studies employing RT-PCR, it has been shown that poliovirus genomic RNA is not stable in non-sterilised seawater (Tsai et al., 1995). Although free DNA is fairly stable, it is unlikely that a free single-stranded RNA genome like those of noroviruses or hepatitis A virus would remain stable without its protein coat in the marine environment. This presumption is less clear for the double stranded RNA genome of rotaviruses. A possible approach for the molecular recovery of infectious viruses is to employ an antibody capture RT-PCR. This has been applied to the detection of hepatitis A virus in seeded shellfish samples and shown to be both sensitive and useful to remove RT-PCR inhibitors as well (Graff et al., 1993; Deng et al., 1994; LoÂpez-Sabater et al., 1997). Since recognition by a conformationally dependent monoclonal antibody is lost when the particle conformation is altered, coupling of the molecular procedure with capture with this type of antibody may enable to discern between intact and altered virions. This approach may prove useful for other enteric viruses, provided that adequate immunological reagents for the most relevant viral pathogens are available. For this purpose, recombinant virus-like particles, which can be obtained in very high numbers in in vitro expression systems (Crawford et al., 1994; Lawton et al., 1997; Caballero et al., 2004), may be employed for the production of antibodies of non-culturable viruses. Recent developments describe methods based on antigenic detection for norovirus (Tian and Mandrell, 2006; Colquhoum et al., 2006) but the high diversity of norovirus may limit their specificity and sensitivity (Zheng et al., 2006).
3.6
Depuration of viral contaminants in molluscan shellfish
Conventional commercial processes employed to purge the microbial contamination of live bivalves are depuration, performed in tanks, and relaying,
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Shellfish safety and quality
performed in the natural environment. Tank-based depuration is now widely practised in many European countries, while it is less widely used in the US (Richards, 1988; Otwell et al., 1991). Depuration periods may vary from 1 to 7 days, since minimum time periods for depuration are not stipulated in the legislation, with around 2 days being probably the most widely used. Early studies, using artificially spiked softshell clams, reported that most viruses were purged within a 24±48-hour period, and that low levels of viruses were depurated more rapidly than high levels (Metcalf et al., 1979). More recent studies show that although depuration and relaying procedures may be insufficient to completely remove viruses (Abad et al., 1997; Richards, 1988; Schwab et al., 1998; De Medici et al., 2001; Kingsley and Richards, 2003; Pommepuy et al., 2003; Loisy et al. 2005b), they do contribute to reduce virus levels and hence the risk of infection due to shellfish consumption (Bosch et al., 1994). Process temperature appears as an important factor for effective virus removal (Power and Collins, 1990; Dore et al., 1998; Jaykus et al., 1994; Pommepuy et al., 2003), although a raise in depuration temperature may result in increased shellfish mortality. Nevertheless, epidemiological evidence reveals that enteric viruses can be transmitted to humans after consuming shellfish which has been depurated (Gill et al., 1983; Heller et al., 1986; Cook and Ellender, 1986; Sockett et al., 1993). Once again compliance with bacterial endproduct standards does not provide a guarantee of virus absence, and bacterial depuration rates can not accurately predict virus removal rates. Using rotavirus virus-like particles (VLP)s the long-term persistence of the surrogate was demonstrated (Loisy et al., 2005b). After several weeks in natural conditions, the surrogate was still detected, suggesting that after contamination by human enteric viruses, shellfish may be unsafe for human consumption for quite a long period of time (Loisy et al., 2005b). The finding of specific attachment of Norwalk virus or recombinant VLPs to shellfish digestive tissues and the capture of some particles by shellfish phagocytes may explain why depuration in oysters is not an effective mechanisms for eliminating virus (Le Guyader et al., 2006).
3.7
Future trends in virus studies in shellfish
The past two decades of virological research have contributed to significant advances in the field of medical virology. These include the development of methodologies for the detection and characterisation of non-culturable waterborne and foodborne viruses, the recognition of waterborne outbreaks caused by hepatitis A and E viruses, the consideration of rotavirus as the single most important cause of severe children gastroenteritis and norovirus as the most frequent agent of foodborne diarrhoea, the characterisation of other important agents of non-bacterial gastroenteritis such as astroviruses, sapoviruses, adenoviruses, and the assessment of the zoonotical transmission of some of the aforementioned agents. A poorly understood aspect in the epidemiology of several enteric viruses that requires further attention is the role of animal viruses in human disease.
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Nucleotide sequence analysis of some human enteric viruses has indicated a high degree of sequence similarity with animal strains. Notably, hepatitis E virus-related sequences have been detected in pigs (Meng, 2000; van der Poel et al., 2001; Banks et al., 2004) and birds (Huang et al., 2002). Zoonotic infections may occur either through direct transmission, suspected for hepatitis E virus (HEV; Reyes, 1993) and caliciviruses (Humphrey et al., 1984), or through incidental co-infection of a host with animal and human viruses, resulting in the mixing of genes and generation of novel variants (recombination/reassortment; Unicomb et al., 1999). Recombination has been demonstrated as a mechanism for rapid expansion of diversity for noroviruses and rotaviruses, but it is likely to be a common feature of the RNA viruses involved (Jiang et al., 1999; Unicomb et al., 1999; Lopman et al. 2004). Viruses related to the human rotaviruses, astroviruses, noroviruses, sapoviruses, and HEV circulate in several animal species, providing a huge reservoir for virus diversity (Shirai et al., 1985; Meng et al., 1997; van der Poel et al., 2001; Huang et al., 2002; Oliver et al., 2006). To corroborate these hypotheses, animal viruses have been recently characterised in shellfish samples from the market either in Europe or in the US (Dubois et al., 2004; Costantini et al., 2006). SARS, reported in November 2002 (Ksiazek et al., 2003), is an example of an emerging disease. The primary mode of transmission of the SARS coronavirus appears to be direct mucous membrane contact with infectious respiratory droplets and/or through exposure to fomites. Several coronaviruses are known to spread by the faecal±oral route, but there is no current evidence that this mode of transmission plays a key role in the transmission of SARS, although there is a considerable shedding of the virus in stools (Tsang, 2003). Another emerging pathogen of concern is the avian influenza H5N1 virus, highly pathogenic among birds, that in some cases has been transmitted from birds to humans. Most cases of H5N1 infection in humans to date have occurred as a result of direct contact with poultry or with surfaces and objects contaminated by their faeces. However, concern has recently been expressed about the potential for transmission of the virus to humans through water and sewage, although no definitive cases have been reported to date (WHO, 2006, http://www.who.int/ water_sanitation_health/emerging/h5n1background.pdf). The availability of quantitative and standardised virus methods will enable the future setting of legislative virus standards for bathing waters, bivalve shellfish and shellfish growing waters. Achievement of this objective will also enable the identification of key environmental factors, such as rainfall and sewage discharges, responsible for viral contamination in bathing and shellfish harvesting areas. Identification, and management, of such critical control points will provide an alternative approach to containing the virus risk and would permit the development of enhanced sanitary controls. Finally, another long-time challenge in environmental virology is to conduct actual field studies to evaluate the environmental behaviour of human enteric viruses, which has to face the impossibility of introducing pathogens in the environment. As model systems, recombinant tracers could be perfectly
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adequate for field studies of microbial tracking, since they may be produced in extremely high numbers (several milligram amounts). Additionally, their noninfectious nature makes them completely harmless and suitable for use in scenarios where the use of actual viruses is hampered by the impossibility of introducing potential pathogens into, for instance, shellfish growing waters. Recombinant norovirus particles have been employed to investigate the behaviour of norovirus (Redman et al., 1997; Le Guyader et al., 2006a) and rotavirus (Caballero et al., 2004; Loisy et al., 2004, 2005a) in environmental samples. Obviously, from the strictly structural point of view, there is no better surrogate of an actual virus pathogen to track their behaviour in the environment than a non-infectious virus-like particle of the same virus. The demonstration of the capacity of Norwalk virus to bind to shellfish tissues at the same binding site as that used to human tissues suggests a possible coevolution mechanism involving the oyster as an intermediary vector (Le Guyader et al., 2006). As knowledge increase in understanding the binding of human enteric viruses to humans, more will be understood about their behaviour in shellfish.
3.8
References
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and ESTES M.K. (1996), Collaborative evaluation of a method for the detection of Norwalk virus in shellfish tissues by PCR. Appl. Environ. Microbiol. 62: 254±258. BANKS, M., BENDALL, R., GRIERSON, S., HEATH, G., MITCHELL, J. and DALTON, H. (2004), MCGOVERN, E.R., LE GUYADER, F., METCALF, T.G.
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and LAMBDEN, P.R. (2006), Genotype 1 and genotype 2 bovine noroviruses are antigenically distinct but share a cross reactive epitope with human noroviruses. J. Clin. Microbiol. 44: 992±998. O'MAHONY, M., GOOCH, C.D., SMYTH, D.A., THRUSSELL, A.J., BARTLETT, C.L.R. and NOAH, N.D. (1983), Epidemic hepatitis A from cockles. Lancet I: 518±520. OSORIO, J.E., HUBBARD, G.B., SOIKE, K.F., GIRARD, M., VAN DER WERF, S., MOULIN, J.C. and PALMENBERG, A.C. (1996), Protection of non-murine mammals against encephalomyocarditis virus using a genetically engineered Mengo virus. Vaccine 14: 155±161. OTWELL, W.S., RODRICK, G.E. and MARTIN, R.E (eds) (1991), Molluscan Shellfish Depuration. Boca Raton, FL: CRC Press. PARASHAR, U.D., HOLMAN, R.C., CLARKE, M.J., BRESEE, J.S. and GLASS, R.I. (1998), Hospitalizations associated with rotavirus diarrhea in the United States, 1993 through 1995: surveillance based on the new ICD-9-CM rotavirus-specific diagnostic code. J. Infec. Dis. 177: 13±17. Â , R.M., GAJARDO, R., ABAD, F.X. and BOSCH, A. (1995), Detection of fastidious PINTO infectious enteric viruses in water. Environ. Sci. Tech. 29: 2636±2638. Â , R.M., ABAD, F.X., GAJARDO, R. and BOSCH, A. (1996), Detection of infectious PINTO astroviruses in water. Appl. Environ. Microbiol. 62: 1811±1813. BRIDGER, J.C.
 , R.M., ALEGRE, D., DOMINGUEZ, A., EL-SENOUSY, W.M., SA  NCHEZ, G., VILLENA, C., PINTO COSTAFREDA, M.I., ARAGONEÁS, L. and BOSCH, A. (2007), Epidemiology of hepatitis A
virus in urban sewage from two Mediterranean countries. Epidemiol. Infec. 135: 270±273.
POMMEPUY, M., CAPRAIS, M.-P., LE SAUX, J.-C., LE MENNEC, C., PARNAUDEAU, S., MADEC, Y.,
and LE GUYADER, F.S. (2003), Evaluation of viral shellfish depuration in a semi-professional size tank. In A. Villalba, B. Reguera, J. L. Romalde, R. Beiras (eds), Molluscan Shellfish Safety (pp 485±499). Conseilleria de Pesca e Assuntos maritimos da Xunta de Galicia and Intergovernmental Oceanographic Commission of UNESCO, Santiago de Compostela (Spain). POMMEPUY, M., HERVIO-HEATH, D., CAPRAIS, M.-P., GOURMELON, M. and LE GUYADER, F.S. (2005), Fecal contamination in coastal area: an engineering approach. In S. Belkin (ed.), Oceans and Health: Pathogens in the Marine Environment (pp 331±360). Kluwer Academic/Plenum Publishers, New York. POWER, U.F. and COLLINS, J.K. (1989), Differential depuration of poliovirus Escherichia coli and a coliphage by the common mussel Mytilus edulis. Appl. Environ. Microbiol. 55: 1386±1390. POWER, U.F. and COLLINS, J.K. (1990), Elimination of coliphages and Escherichia coli from mussels during depuration under varying conditions of temperature salinity and food availability. J. Food Prot. 53: 208±212, 226. RAO, V.C. and MELNICK, J.L. (1986), Environmental virology. In J.A. Cole, C.J. Knowles and D. Schlessinger (eds), Aspects of Microbiology 13. American Society for Microbiology, Washington, DC. RAO, V.C., METCALF, T.G. and MELNICK, J.L. (1986), Human viruses in sediments, sludges, and soils. Bull. WHO 64: 1±13. MONNIER, M., BREST, G.
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and ESTES, M.K. (1997), Filtration of recombinant Norwalk Virus particles and bacteriophage MS2 in quartz sand:Importance of electrostatic interactions. Environ. Sci. Tech. 31: 3378±3383. REYES, G.R. (1993), Hepatitis E virus (HEV): molecular biology and emerging epidemiology. Prog. Liver Dis. 11: 203±213. REYNOLDS, K.A., GERBA, C.P. and PEPPER, I.L. (1996), Detection of infectious enteroviruses by an integrated cell culture-PCR procedure. Appl. Environ. Microbiol. 62: 1424±1427. RICHARDS, G.P. (1985), Outbreaks of shellfish-associated enteric illness in the United States: requisite for development of viral guidelines. J. Food Prot. 48: 815±823. RICHARDS, G.P. (1988), Microbial purification of shellfish: a review of depuration and relaying. J. Food Prot. 51: 218±251. ROMALDE, J.L., ESTES, M.K., SZUCS, G., ATMAR, R.L., WOODLEY C.M. and METCALF, T.G. (1994), In situ detection of hepatitis A virus in cell cultures and shellfish tissues. Appl. Environ. Microbiol. 60: 1921±1926. REDMAN, J.A., GRANT, S.B., OLSON, T.M., HARDY, M.E.
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and BOSCH, A. (2002), Prevalence of enterovirus and hepatitis A virus in bivalve mollusks from Galicia (NW Spain): inadequacy of the EU standards of microbiological quality. Int. J. Food Microbiol. 74: 119±130. Â NCHEZ, G., PINTO Â , R.M., VANACLOCHA, H. and BOSCH, A. (2002), Molecular characterization SA of hepatitis A virus isolates from a transcontinental shellfishborne outbreak. J. Clin. Microbiol. 40: 4148±4155. SCHLAUDER, G.G. and MUSHAHWAR, I.K. (2001), Genetic heterogeneity of hepatitis E virus. J. Med. Virol. 65: 282±292. SCHWAB, K.J., NEILL, F.H., ESTES, M.K., METCALF, T.G. and ATMAR, R.L. (1998), Distribution of Norwalk virus within shellfish following bioaccumulation and subsequent depuration by detection using RT-PCR. J. Food Prot. 61: 1674±1680. J.L.
SCHWAB, K.J., NEILL, F.H., FANKHAUSER, R.L., DANIELS, N.A., MONROE, S.S., BERGMIRE-SWEAT,
D.A., ESTES, M.K. and ATMAR, R.L. (2000), Development of methods to detect `Norwalk-like viruses' (NLVs) and hepatitis A virus in delicatessen foods: application to a food-borne NLV outbreak. Appl. Environ. Microbiol. 66: 213±218. SHIEH, C.Y.-S. CALCI, K.R. and BARIC, R.S. (1999), A method to detect low levels of enteric viruses in contaminated oysters. Appl. Environ. Microbiol. 65: 4709±4714. SHIRAI, J., SHIMIZU, M. and FUKUSHO, A. (1985), Coronavirus-, calicivirus-, and astroviruslike particles associated with acute porcine gastroenteritis. Nippon Juigaku Zasshi 47: 1023±1026. SMITH, E.M. and GERBA, C.P. (1982), Development of a method for detection of human rotavirus in water and sewage. Appl. Environ. Microbiol. 43: 1440±1450. SOBSEY, M.D., CARRICK, R.J. and JENSEN, H.R. (1978), Improved methods for detecting enteric viruses in oysters. Appl. Environ. Microbiol. 36: 121±128. SOCKETT, P.N., COWDEN, J.M., LEBAIGUE, S., ROSS, D., ADAK, G. and EVANS, H. (1993), Foodborne disease surveillance in England and Wales: 1989±1991. Com. Dis. Rep. 3: 159±174. TARTERA, C. and JOFRE, J. (1987), Bacteriophages active against Bacteroides fragilis in sewage-polluted waters. Appl. Environ. Microbiol. 53: 1632±1637. TIAN, P. and MANDRELL, R. (2006), Detection of norovirus capsid proteins in faecal and food samples by a real-time immuno-PCR method. J. Appl. Microbiol. 100: 564± 574. TSAI, Y.L., SOBSEY, M.D., SANGERMANO, L.R. and PALMER, C.J. (1993), Simple method of concentrating enteroviruses and hepatitis A virus from sewage and ocean water for
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rapid detection by reverse transcriptase-polymerase chain reaction. Appl. Environ. Microbiol. 59: 3488±3491. TSAI, Y.-L., TRAN, B. and PALMER, C.J. (1995), Analysis of viral RNA persistence in seawater by reverse transcriptase-PCR. Appl. Environ. Microbiol. 61: 363±366. TSANG, T. (2003), Environmental issues. WHO Global Conference on Severe Acute Respiratory Syndrome (SARS), Kuala Lumpur, Malaysia, 17±18 June 2003. UNICOMB, L.E., PODDER, G., GENTSCH, J.R., WOODS P.A., HASAN, K.Z., FARUQUE, A.S., ALBERT,
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M.J.
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and KOOPMANS, M.P.G. (2003), International collaborative study to compare reverse transcriptase PCR assays for detection and genotyping of noroviruses. J. Clin. Microbiol. 41: 4023±4033. WILLIAMS, R.A. and ZORN, D.J. (1997), Hazard analysis and critical control point systems applied to public health risks: the example of seafood. Rev. Scientif. Tech. 16: 349± 358. YATES, S., PENNING, M., GOUDSMIT, J., FRANTZEN, I., VAN DER WEIJER, B., VAN STRIJP, D. and VAN GEMEN, B. (2001), Quantitative detection of hepatitis B virus DNA by real-time nucleic acid sequence-based amplification with molecular beacon detection. J. Clin. Microbiol. 39: 3656±3665. ZHENG D-P, ANDO T., FANKHAUSER R.L., BEARD R.S., GLASS R. and MONROE S.S (2006), Norovirus classification and proposed strain nomenclature. Virology 346: 312±323.
4 Monitoring viral contamination of molluscan shellfish M. Pommepuy, J. C. Le Saux, D. Hervio-Heath and S. F. Le Guyader, IFREMER, France
Abstract: This chapter focuses on the enteric viruses, hepatitis A virus and norovirus, responsible for the main outbreaks linked to shellfish consumption. One major drawback for viral contamination estimation in shellfish is the lack of indicator or standardised method, preventing systematic control. After a review of the main sources of pollution, we identified the conditions responsible for shellfish contamination. This will help to set up potential strategies to reduce contamination in harvesting areas such as the limitation of faecal input or the improvement of tools for management. As knowledge is progressing risk management, strategies will help to protect the consumer and will also led to improve regulation. Rapid alert systems need to be set up to prevent coastal area contamination and promising examples of coastal management have proved their efficacy to reduce faecal load and thus, improve water quality. Shellfish have long been recognised as being beneficial to human health and this benefit should also be taken into consideration in managing coastal areas and preserving the water quality. Key words: norovirus, hepatitis A virus, shellfish contamination, coastal management, rapid alert system.
4.1
Introduction
Infectious diseases linked to the consumption of raw shellfish such as oysters, mussels, cockles and clams, have long been identified. Bacterial diseases such as cholera and typhoid fever were the first to be suspected of being linked to consumption of contaminated shellfish (Butt et al., 2004). During the past
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century, various strategies have been established in shellfish growing areas throughout the world to assure the sanitary quality of shellfish. More recently, despite sanitary surveys, Vibrio parahaemolyticus, Vibrio vulnificus and enteric viruses ± especially Hepatitis A virus (HAV) and norovirus (NoV) ± were found to be associated in outbreaks of human illness. The pathogens involved in shellfish foodborne diseases can be placed in two classes. The first includes environmental pathogens that normally spend a substantial part of their life cycle outside human hosts, but which when introduced to humans cause disease with a measurable frequency (Cangelosi et al., 2004). Among them, vibrios (V. cholerae, V. parahaemolyticus, V. vulnificus) are common in marine environmental infections, especially in countries where climatic conditions allow them to proliferate (southeast US coast, South America and Asian countries), while few cases are reported in Europe. In the second class, the enteric pathogens are non-autochthonous microorganisms, discharged into the sea by raw or insufficiently treated wastewaters during epidemics in the population. Most of the time they have been excreted by sick people living on coastal watersheds, but they may be present in the intestines of healthy humans or in the animal population (van der Poel et al., 2001). Among them, viruses (especially NoV and HAV), are the chief concern in shellfishborne diseases, while bacterial infections (salmonelloses, typhoid fever) have decreased thanks to sanitary control measures set up over the past century (development of detection methods and bacterial surveys and improvement of shellfish depuration technology). This review will focus on the enteric viruses which are responsible for the main outbreaks linked to shellfish consumption (Butt et al., 2004). Viral outbreaks association with contaminated shellfish consumption was first suggested more than 50 years ago. Initially, the analysis of outbreaks was mainly based on epidemiological data and symptoms in patients (Mackowiack et al., 1976; Grohmann et al., 1980; Richards, 1987). In some cases, microscopic studies identified small viruses in patients' stools or in shellfish (Appleton and Pereira, 1977; Morse et al., 1986; Pontefract et al., 1993). The development of molecular biology and thus the ability to find low levels of enteric viruses in shellfish, has provided more accurate assessment of shellfish as disease transmission vehicles (Lees 2000; Sanchez et al., 2002; Butt et al., 2004; Boxman et al., 2006; Le Guyader et al., 1996, 2003, 2006). Despite the fact that many enteric viruses can be detected in human faeces (Metcalf et al., 1995; Bosch et al., 2008), only HAV and NoV have been clearly identified as infectious agents in consumed shellfish. Linking cases of viral disease to contaminated shellfish is not that easy, particularly for NoV, owing to the multiplicity of lineages circulating at the same time (Blanton et al., 2006; Zheng et al., 2006). One major drawback in shellfish outbreaks is the lack of consistent correlation between the indicator of fecal contamination (Escherichia coli or faecal coliforms) and human enteric viruses and thus an absence of precaution (Lees, 2000; Butt et al., 2004). In many viral outbreaks related to shellfish consumption, the level of E. coli was in compliance with the regulations
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(Christensen et al., 1998; Le Guyader et al., 2003, 2006; Kohn et al., 1995; Boxman et al., 2006). Studies have been carried out in various countries to examine the prevalence of enteric viruses in shellfish. When we focus on data obtained from shellfish collected from producing areas or from the market, and showing no bacterial contamination as defined by current regulations, noroviruses were detected from 6 to 37% of the time (Henshilwood et al., 1998; Le Guyader et al., 2000; Beuret et al., 2003; Nishida et al., 2003; Cheng et al., 2005; Costantini et al., 2006). Different concentration and extraction methods and reverse-transcriptase polymerase chain reaction (RT-PCR) assays used, as well as the sampling season and the year of the study may explain the differences between the studies. It is also possible that prevalence surveys with positive findings may be overrepresented due to a publication bias. Some outbreaks have been linked to depurated oysters (Grohmann et al., 1980; Le Guyader et al., 2006). Nonetheless, it must be kept in mind that while E. coli may disappear rather quickly either by depuration practices or natural cleansing, recent data show that viral depuration is difficult. Therefore, the 2 days of depuration stipulated by EC regulations is inefficient in eliminating viral contamination (Schwab et al., 1998; De Medici et al., 2001; Loisy et al., 2005).
4.2
Identifying sources of pollution
In some outbreaks, multiple strains of a single virus such as norovirus can be detected indicating sewage or faecal contamination (Kageyama et al., 2004; Gallimore et al., 2005; Boxman et al., 2006; Le Guyader et al., 2006). Analysis of shellfish events leading to shellfish-related outbreaks has confirmed this hypothesis, and when environmental data are available, sewage-related contamination is often demonstrated (Table 4.1). Flooding has been shown to be responsible for viral contamination in other outbreaks and is congruent with the sudden introduction of multiple NoV strains into the oyster breeding site (Mackowiak et al., 1976; Le Guyader et al., 2006). In an attempt to identify the source of pollution in an oyster-producing area, Ueki et al., (2005) did a 1-year study collecting sewage, river water, oysters and clinical samples and looking for noroviruses. A clear impact of the sewage treatment plant was found on river contamination (about 75% of the sample being contaminated) and in shellfish samples (60% contaminated). Analysis of sequences obtained showed that the same diversity was observed among strains circulating in the population, as well as in sewage or in oyster samples, leading to the conclusion that improved sewage treatment is needed to guarantee the sanitary quality of these shellfish (Ueki et al., 2005). The literature review mainly highlights the human origin of viral shellfish outbreaks, i.e. the responsibility of urban wastewater. Nevertheless, in the state of current knowledge, animal sources cannot be ignored. For example, in the Netherlands, 44% of bovine fecal samples tested positive for norovirus (Van der
Table 4.1
Analysis of shellfish events leading to outbreaks of human illness
Shellfish
Country
Outbreak
Source of contamination
Comments or treatment proposed
References
Oysters
Australia
Gastroenteritis
Heavy rain 2 months before
Murphy et al. (1979)
Cockles
UK
Hepatitis A
Sewage pollution
Cockles Clams, oysters Clams
Singapore USA
Hepatitis A Gastroenteritis
Goh et al. (1984) Morse et al. (1986)
China
Hepatitis A
Vaccination of local population
Xu et al. (1992)
Oysters
USA
Gastroenteritis
No comment
Pontefract et al. (1993)
Oysters
USA
Gastroenteritis
Australia
Gastroenteritis
Oysters
UK
Gastroenteritis
Closure of shellfish area for 1 month Harvesting of oysters prohibited for several months No comment
Kohn et al. (1995)
Oysters
Oysters
France
Gastroenteritis
Sewage pollution Illegal harvesting from contaminated area, flooding Outbreak in the population 6 months ago, direct sewage input Oyster collected in various area Human sewage disposal by harvesting boat Leaking sewage, poor tidal exchange Class B area, depurated oysters Sewage contamination
Sale prohibited, 2 days of depuration, volunteer testing before marketing Pasteurisation or steam treatment of all shellfish from this area Boiling of cockles Steamed clams still infectious
Oysters Oysters
Singapore France, Italy
Gastroenteritis Gastroenteritis
Imported frozen from China Heavy rain and sewage treatment failure
Persistence of contamination for several weeks Destruction of all imported oysters Closure of shellfish area for several weeks
O'Mahony et al. (1983)
Stafford et al. (1997) Ang et al. (1998) Le Guyader et al. (2003) Ng et al. (2005) Le Guyader et al. (2006)
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Poel et al., 2001). Even though inter-species exchange is extremely rare and has never been demonstrated for NoV, the high mutation rate could contribute to a species-crossing (Woolhouse and Growtage-Sequeria, 2005). Differences between human and animal strains are very small (Dastjerdi et al., 1999) thus, an intramolecular re-combination could lead to new pathogenic species (Lopman et al., 2004).
4.3 Identifying the conditions responsible for microbial contamination of shellfish In most of the polluted areas, evidence shows that the river and sewage outfalls discharging to the estuaries or marine bay have high levels of bacteria and viruses. Different sources of contamination are currently identified on sites where shellfish are farmed: · Sewage discharges including sewage outfall, combined sewer overflows and stormwater discharges. The type of treatment applied to sewage waters plays an important role in the faecal load discharged into marine waters (physical, biological or tertiary treatments). · Sewage network failures: many stormwater events could contribute to this pollution and could trigger persistent faecal contamination even during dry weather (Armstrong et al., 1996). · River discharges and possible run-off from agricultural activities (Kashefipour et al., 2002; Crowther et al., 2002; Vinten et al., 2004) · Other specific discharges could also come from boats, wild birds, bathers, sediments or other diffuse urban sources such as pigeons, dogs or cats (Sobsey et al., 2003; Gerba 2000; O'Keefe et al., 2005). A high prevalence of human enteric viruses was reported in raw sewage with astrovirus concentrations ranging from 5 105 to 5 107 genome units (GU)/ 100 ml (Le Cann et al., 2004) or 3 103 to 1:2 108 genomes/l (Myrmel et al., 2006). For norovirus, values varying from 3 101 to 8:5 104 GU/100ml (Lodder and de Roda Husman, 2005), 1:8 104 to 9:7 105 genome/l (Myrmel et al., 2006), less than 103 to 106 PCR detectable units/l (pdu/l) (van der Berg et al., 2005) and up to 1:7 107 genomes (G)/l (Laverick et al., 2004) have been reported in different studies. In treated water, concentrations vary greatly depending on the location: 2 102 to 5 104 GU/100 ml for astrovirus (Le Cann et al., 2004), and for norovirus 4 101 to 2:4 103 GU/100 ml (Lodder and de Roda Husman, 2005), 1:6 105 G/l (Laverick et al., 2004), or 5:7 102 pdu/l (van der Berg et al., 2005). Rivers can also be contaminated by norovirus and enterovirus (Schvoerer et al., 2001; Hot et al., 2003). On the basis of recent quantitative data, it was estimated that norovirus concentrations could reach 5 102 GU/100 ml in river (Lodder and de Roda Husman, 2005), 1:6 103 G/l (Laverick et al., 2004). In floodwater sampled after a tropical storm event, Phanuwan et al. (2006) found a high prevalence of viruses with a geometric
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113
mean of 14.0 pdu/ml for adenovirus, 13.0 pdu/ml (HAV), 5.3 pdu/ml (norovirus ggII), 0.7 pdu/ml (enterovirus) and 0.003 pdu/ml (norovirus ggI). Seasonal variations in viral contamination were also reported. In the Meuse River, Westrell et al. (2006) observed a distribution of norovirus with high peaks during winter (up to 1750 pdu/l) linked to contamination events in the catchments, especially sewage treatment failures. Noble and Furman (2001) observed the presence of enteroviruses in coastal water of Santa Monica Bay, California, especially during the winter wet season and the summer dry season. An interesting study by Haramoto et al. (2005) showed a mean concentration of about 0.087 G/ml for genogroup I (ggI) norovirus and 0.61 G/ml for ggII, with seasonal variations. In winter, these concentrations reached 0.21 G/ml for ggI and 2.3 G/ml for ggII. Lower concentrations were observed in summer (0.016 G/ml for ggI and 0.026 G/ml for ggII). For HAV, concentrations varying from 90 to 3523 copies/l were detected in estuarine waters along the Mexican borders (Brooks et al., 2005). In some cases, seasonal patterns were observed reflecting the clinical epidemiology of the agent, in both river studies (Pusch et al., 2005; van der Berg et al., 2005) and a shellfish study (Le Guyader et al., 2000). The effect of rainfall on faecal water quality and on viral contamination has been also widely reported (Miossec et al., 1998; Noble et al., 2003; Haramoto et al., 2006). Storm events cause land-based runoff and raw sewage overflow. Dramatic effects can be observed in developing countries with tropical climates and poor sewer systems. In Jakarta, Phanuwan et al. (2006) reported high concentrations of enterovirus, HAV and NoV in floodwaters after a storm event, leading to a high risk of viral infection for people who drink the water or who are in contact with overflow waters. In temperate or semi-arid climates, rainy storm events also lead to deterioration of seawater quality. Statistic observations showed that over half of the beach water quality failures in Santa Monica Bay, California, were associated with rain events (Schiff et al., 2003). Regional differences due to site-specificity also affect the impact of rainfall. In dry weather, Noble and Fuhrman (2001) observed that water quality standards were exceeded five times more often on Mexican beaches than on US beaches. In a multi-country study in Europe (Spain, Greece, Sweden, UK), using modelling approaches to compare NoV shellfish contamination, the country was found to be a significant input variable, and site-specific relationships between indicators and pathogens were highlighted (Brion et al., 2004). Moreover, the increase in coastal populations contributes to physical changes in the landscape and leads to degradation of water quality. Mallin et al. (2001) demonstrated that the coastal demographic increase over a 14-year period in coastal North Carolina, was directly correlated with the upward trends in shellfish bed closures. In this region, population growth led to limiting of natural soil filtration by impermeable surfaces (e.g. roads, parking, roofs), new faecal sources and more rapid conveyance of pollutants to the sea. At the same time, in the same area, high sensitivity to forecast events was observed when compared with preserved coastal wetlands.
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4.4 Potential strategies for reducing microbial contamination in shellfish harvesting areas Over the last decade, coastal management has become increasingly important because of socioeconomic requirements (tourism, aquaculture) leading to higher water quality objectives, and to increased pollution due to the upshift in coastal demography and industrial development. In addition to the complexity created by the topography in coastal catchments for raw wastewater collection, it is also difficult to have facilities which can cope with large seasonal variations in the faecal load. In some villages or small towns devoted to tourist activities, the population increases by 10-fold or more, for a few weeks in the summer season. Under these conditions, sewage treatment plants (STP) must deal with the need for increased capacity for just a few weeks in summer. In France, in coastal areas, primary treatment plants and activated sludge systems account for 17 and 61% respectively . The majority of small towns on the coast have been equipped only recently. Cities (>5000 inhabitants) have set up tertiary treatment systems, which represent only 22% of the STPs located near the sea (IFEN, 2001). 4.4.1 Ways in which faecal input can be limited To limit pollution from urban wastewater, different types of treatment are now available. In general, processes are applied to eliminate the faecal load, but their real efficiency in eliminating viral contamination is often unknown. The very few data obtained under full-scale conditions which have been reported in the literature are presented in Table 4.2. Primary treatment, decantation and coagulation have little effect on viral elimination, and secondary treatments ± biological treatment with activated sludges, with/or without aeration or ponds ± have a limited average efficiency (2 log and less). Tertiary treatments such as chlorination could make virus removal more efficient (Rose et al., 1996). Effectiveness in removing viruses or phages can vary dramatically over time by a factor of 3 or 4 log, ranging from low values 13 cm/s), flocculation and settling were initially observed, but the aggregates were quickly sheared apart, leading to low cell removal; however, cell damage may be more significant at these higher speeds, leading to possible loss of cells even without flocculation and settling. Clays were found to be suitable for this use because cell removal occurs quickly ± within minutes to hours ± and the possibility for environmental and ecological impacts from dispersal were assumed to be minimal as clays are natural materials (Shirota, 1989; Yu et al., 1994b). Furthermore, clays were
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abundant, readily available, inexpensive and easy to handle. Hence, clays were tested and applied in several countries to treat HABs and protect fish aquaculture. Applications Japan The first reports of clay dispersal to control blooms came from Japan (Maruyama et al., 1987; Shirota, 1989). In the field, clay was dispersed in and around mariculture cages at 110±400 g/m2 where fish were dying due to a bloom of Cochlodinium sp. (Shirota, 1989). Results showed that 78% of the fishermen deemed the treatment effective, as virtually no fish mortality was observed even though a dense bloom was present. In fact, moribund fish floating near the surface soon recovered after clay was added. Aerial dispersal of dry clay and slurry was also tested (Shirota, 1989). Helicopters applied clay at 200 g/m2 over affected sites. This method of treatment was also found to be effective, although there were no further tests or cost estimates for this approach. These results seemed to support clay application as a viable control strategy in Japan; however, its use was abandoned because of the high cost of clay, storage, and dispersal. Recently, clays have been re-evaluated and applied in some regions of Japan, on a case-by-case basis (Ichiro Imai, University of Kyoto, personal communication). Unfortunately, there have been no published reports in the literature regarding this new effort. China Most of the initial reports from China regarding the use of clays to treat local HAB species focused on the theoretical mechanisms of cell removal (Yu et al., 1994a,b, 1995a,b), and the enhancement of removal efficiency through the use of chemical flocculants (Yu et al., 1999). Reports of actual application have been limited. Li et al. (1998) tested kaolin combined with several chemical additives against Phaesodactylum triconutum, Gymodinium sp., Nitzchia closterium, and Euglena sp. in prawn culture ponds. Removal rates were 80±90% within 12 h at pH 6. Sun et al. (2001) showed that kaolinite (up to 1 g/L) was not toxic to the fleshy prawn, Penaeus chinensis. In addition, chemical additives that improve the removal ability of kaolinite were also non-toxic to the prawn at low concentrations; however, no data on the actual performance of clays against algal cells were presented in the study. Yu et al. (1995c) showed that kaolinite has a high absorption capacity for phosphate relative to montmorillonite at pH < 8.5. South Korea The most significant application of clay in the field has been performed in South Korea. Except for one year, clays (referred to as loess or yellow clay) have been used for the past decade to deal with blooms of Cochlodinium polykrikoides and their impact on fish aquaculture. In 1995, an intense bloom of C. polykrikoides devastated the industry, causing over 76.4 billion won (US$100 million) in
Mitigation of effects of harmful algal blooms 179
Fig. 7.1 Dispersal of yellow clay to remove Cochlodinium polykrikoides near fish pens in South Korea (courtesy National Fisheries Research and Development Institute Pusan, South Korea).
damage (Bae et al., 1998). During the following year, yellow clay was added in and around the fish pens (Figs 7.1 and 7.2), while larger boats dispensed clay upstream from the fish cages to reduce the number of cells approaching the pens (Fig. 7.3). About 60 000 tons of yellow clay were dispersed over 260 km2 at the rate of 400 g/m2 (Na et al., 1996). Removal rates were between 90% and 99% down to 2 m depth. Turbidity improved after 4 h and lower fish mortalities were found. Fisheries losses were reduced to US$1 million following this first application. In 1997, clay treatment was repeated using 50 000 tons with similar results. Improvements in dispersal methods and the use of electrified water to induce cell death have dramatically improved the efficacy of the treatment (Kim et al., 2000).
Fig. 7.2 Dispersal of yellow clay around fish pens in South Korea to treat harmful blooms of Cochlodinium polykrikoides (courtesy National Fisheries Research and Development Institute Pusan, South Korea).
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Fig. 7.3 Large-scale dispersal of yellow clay to treat Cochlodinium polykrikoides blooms (courtesy National Fisheries Research and Development Institute Pusan, South Korea).
Despite these improvements in removal efficiency, the removal ability of yellow clay remained much lower than that of other minerals. Therefore, Sun et al. (2004a) examined the use of sophorolipid, from the fungus Candida bombicola, on the removal ability and lethality of yellow clay. Laboratory and field trials showed improvements in cell removal with 10% less clay relative to clay alone, resulting in a 60% decrease in treatment cost. Although it was considered a promising innovation in clay application, there have been no further reports on its implementation in Korean coastal waters. United States The efficacy of clay flocculation for controlling HABs in the US, and its potential impacts, have been examined for several years from bench-top experiments to mesocosms (e.g., limnocorrals and flumes) and open water trials (Sengco and Anderson, 2004). Generally, expanding clays with a high proportion of fine particles (80%) (Sengco et al., 2001; Sengco and Anderson, 2004). Clays such as bentonite and phosphatic clay ± a by-product of phosphate mining in central Florida ± displayed consistently high removal ability. Phosphatic clays can also effectively remove extracellular brevetoxins from the media, in addition to intracellular toxins within the flocculated cells (Pierce et al., 2004). The removal of Karenia brevis with phosphatic clay (0.25 g/L) remained moderately high (>80% RE) in mesoscale experiments. Experiments were
Mitigation of effects of harmful algal blooms 181 conducted in settling columns (12.8 L) (Sengco, unpublished data), fiberglass enclosures (530 L), and limnocorrals (9400 L) (Sengco, unpublished data). Later, experiments in various flumes showed high cell removal at low current velocities (e.g. 3±10 cm/s) (Archambault et al., 2003; Beaulieu et al., 2005). Cell removal decreased as current speed increased owing to little or no settling, and the shearing of aggregates starting with velocities 13 cm/s. The combination of clays and flocculants such as polyaluminum chloride (PAC) and various cationic polymers increased cell removal relative to clays alone (Sengco et al., 2001). Recent studies also showed that chemical flocculants were needed to improve the removal ability of clays against brackish water species such as Prymnesium parvum (Sengco et al., 2005). Essentially, cationic flocculants increased the adhesiveness of the clay particles for the algal cells. Flume studies revealed, however, that flocculant-treated clays produced more voluminous aggregates that settled more slowly relative to untreated clays, leading to longer residence time in the water column (i.e. prolonged elevated turbidity) (Beaulieu et al., 2005). Furthermore, the floc layer formed by flocculant-treated clays was more readily resuspended compared to a sediment layer composed of untreated clays. This is potentially important in managing the redistribution of clays along the bottom, and the possible impact of clays on benthic organisms. At low clay concentration (0.03 g/L), Karenia brevis can remain viable, escape from the floc layer with or without resuspension, and resume vegetative growth (Sengco et al., 2001). At sufficiently high concentrations (0.50 g/L), high cell mortality (up to 100%) and no recovery was observed even with frequent resuspension. At intermediate concentrations (e.g. 0.10±0.25 g/L), survival and recovery depended on the interplay of clay amount, the frequency of resuspension, or the duration of contact between the cells and clays prior to the first resuspension event. In flow, K. brevis cells lost their motility within 30 min of exposure to phosphatic clay (0.25 g/L) at low velocity (3 cm/s) relative to clay-free controls (Sengco, unpublished data). This was followed by the loss of the organisms' characteristic morphology. At 13 cm/s, the loss of motility and shape occurred within minutes of exposure to clay. Several studies have been conducted in Puget Sound (Washington, USA) to examine the potential of using phosphatic clays to control Heterosigma akashiwo that threaten salmon aquaculture. In one report (Rensel and Anderson, 2004), hydrated phosphatic clay (375 g/m2) was added to a 144 m2 net-pen assembly at a farm site containing ten adult Atlantic salmon. Clay was also added to between two and five 4.4 m2 pens at 200 g/m2. The removal efficiencies of microflagellates ranged from 86 to 99%. Diatoms at the farm site were removed at 48 to 87%. The salmon displayed coughing behavior after the clay was added, but recovered after the material settled. In a second study (Rensel et al., 2003), an actual H. akashiwo bloom was treated with phosphatic clay (300 g/m2 equivalent to 0.12 g/L) within replicated, open-ended, floating mesocosms in East Sound, Orcas Island. Cell removal was rapid and averaged 84%, which was comparable to or slightly less than in the
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previous report (Rensel and Anderson, 2004), but better than expected compared with laboratory studies by Sengco et al. (2001). Impacts of clay flocculation on benthos The addition of clay can affect water quality and the biological communities in the upper water column. Although these potential impacts are certainly important, this section focuses on the possible impacts of dispersal on the benthic organisms. Clay flocculation essentially turns the bloom from a surface problem to a potentially significant problem along the bottom. There are concerns associated with the fate of algal toxins, the decay of high biomass resulting in oxygen depletion, and the introduction of fine sediments. There have been a number of studies that focused on the impact of clay addition in the context of HAB control. Lewis et al. (2003) demonstrated that phosphatic clay combined with polyaluminum chloride (PAC), in the absence of Karenia brevis, was not lethal to juvenile fish (Cyprinodon variegates), and to epibenthic and infaunal invertebrates (Ampelisca abdita, Leptocheirus plumulosus, and Palaemonetes pugio), following acute or chronic exposures. When K. brevis was incorporated into the clay/flocculant aggregates, chronic and acute toxicities in all four organisms were observed, but the results were similar to the toxicity of settled K. brevis cells alone (control). These results suggested that the use of phosphatic clay may not be any worse (or better) than what may occur during an untreated bloom. Shumway et al. (2003) examined the impacts of yellow loess (now called yellow clay) on the clearance rates of some commercially important bivalve molluscs (Crassostrea virginica, C. gigas, Mytilus edulis, M. trossulus, Argopecten irradians), and on the slipper shell Crepidula fornicate, and hydroids. The effect of clay on the clearance rates of the food alga Rhodomonas lens was species-specific. Scallops were the most sensitive, showing a significant decrease in clearance with as little as 0.01 g/L. By contrast, the clearance rate of C. virginica was not affected until clay concentrations reached 1.0 g/L, while the clearance rate of M. edulis decreased significantly between 1 and 10 g/ L. Given the sensitivity of these and other species to even small amounts of clay, the authors recommended great caution in using clays to control HABs. Archambault et al. (2004) conducted a series of 2-week-long flume experiments to determine the impact of fully sedimented (~2 cm/s) and resuspended (~14 cm/s) clay-cell aggregates on the survival and growth of juvenile hard clams, Mercenaria mercenaria. The experiments were performed in a recirculating flume using Heterocapsa triquetra and Prorocentrum micans and phosphatic clay (0.25 g/L). No clam mortalities occurred in either treatment. The fully sedimented treatment showed no significant differences in shell or tissue growth compared with controls (no sediment layer) and the clams rapidly resumed siphon contact with the overlying water. By contrast, a significant growth effect (~90% reduction in shell and tissue growth compared with no-clay controls) occurred in trials with suspended clay. These results suggest that clay
Mitigation of effects of harmful algal blooms 183 application is potentially more detrimental to clams under flow conditions that lead to prolonged resuspension than under conditions that promote rapid sedimentation. This may become an important consideration regarding whether to use chemical flocculants with clay as flocculants can change the settling rate of aggregates and the erodibility of the floc layer (Beaulieu et al., 2005). 7.2.2 Biological control Trophic interactions play a role in the dynamics of algal blooms. This may be referred to as natural biological control. The term biological control in this section, however, refers to the proactive use of pathogens, algal competitors, and grazers to control HABs by taking advantage of these natural antagonisms. Viruses There has been a growing recognition that viruses play a significant role in the decline of blooms (Bratbak et al., 1993; Milligan and Cosper, 1994; Brussard et al., 1999; Wommack and Colwell, 2000). Brussard (2004) provides a current overview on host±virus systems, the diversity of algal viruses, and the role of viruses in bloom mortality and structuring algal populations (Suttle et al., 1991). For instance, viruses may be important in reducing biomass and in preventing the formation of high-biomass blooms. Viruses may also play an important role in energy transfer and nutrient dynamics in aquatic systems. Algal viruses have a strong potential for use as a biological control agent for their specificity (Boesch et al., 1997). Viruses also display rapid replenishment and decay (Fuhrman, 1999). Viruses have been found in a wide variety of microalgae. Suttle et al. (1991) described a method for concentrating viruses in seawater using ultrafiltration to examine their ability to kill algal hosts. Viruses were found that affected Micromonas pusilla, Navicula sp., and Synechococcus sp. Tarutani et al. (2000) examined the impacts of a virus (HaV) on the dynamics of the raphidophyte, Heterosigma akashiwo, in Hiroshima Bay (Japan). In laboratory studies, several isolates of HaV and H. akashiwo were tested, showing that HaV displayed some clonal specificity. Lawrence et al. (2001) also found a virus infecting H. akashiwo. Called Heterosigma akashiwo nuclear inclusion virus (HaNIV), the virus forms in the host nucleus 24 h after inoculation. Within 74 hrs, 98% of the cells showed signs of infection. Later, Tarutani et al. (2001) isolated viruses in Heterocapsa circularisquama, a shellfish-killing dinoflagellate from Japanese coastal waters that were capable of destroying several strains, but not other phytoplankton species. Onji et al. (2000) isolated 18 samples of virus-like agents from Funka Bay (Hokkaido, Japan) that suppressed growth of Alexandrium catenella, Gymnodinium mikimotoi (now Karenia mikimotoi), and Tetraselmis sp. Virus-like particles (VLP) have also been found in the cyanobacterium, Lynbya majuscula, in Moreton Bay (Australia). Exposure to the virus led to decrease in host fluorescence, photosynthetic efficiency and electron transport rate after 5 days.
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Virus-like particles have also been identified in the brown-tide species, Aureococcus anophagefferens (Sieburth et al., 1988). Gastrich et al. (2002) found the prevalence of the virus ranged from 0.2 to 8.1% within the population. Ultrastructural differences were observed between infected and healthy cells. Gobler et al. (2004) found higher densities of viruses during blooms compared with most estuarine environments. However, mesocosm experiments showed an increase in growth rates of A. anophagefferens with the addition of viruses relative to controls. It was suggested that the positive growth response was due to the release of organic nutrients that stimulate the bloom, or alterations in the microbial communities. Bacteria Bacteria can influence the growth and decline of blooms (Doucette, 1995; Doucette et al., 1999). They can cause algal mortalities through direct attack or release of algaecidal substances (Imai, 1997). In a review, however, Mayali and Azam (2004) concluded that the evidence for algaecidal bacteria influencing the decline of blooms may be circumstantial due to limitations of current methods. Numerous studies and observations have been made regarding the interaction between algaecidal bacteria and bloom-forming species. For instance, Furuki and Kobayashi (1991) noted that certain bacteria can both promote and inhibit the proliferation of Chatonella sp. in the Sea of Harima (Japan). Similarly, natural bacterial communities isolated from an inlet in Kochi (Japan) influenced the growth of the dinoflagellate Gymnodinium nagasakiense and the diatom Skeletonema costatum (Fukami et al., 1991). Yoshinaga et al. (1997) isolated 28 strains of bacteria from various genera that could kill Gymnodinium (now Karenia) mikimotoi after a bloom that occurred in Tanabe Bay (Japan). Most of the bacteria did not affect several diatoms and the dinoflagellate Alexandrium catenella, which suggested specificity for the target species. Imai (1997) examined several strains of Alteromonas sp. and Cytophaga sp. for their killing ability against Chatonella ovata, C. verruculosa, Alexandrium tamarense, Heterocapsa circularisquama, Eutreptiella gymnastica, and Oltmannsiellopsis viridis. The two thecate dinoflagellates (A. tamarense and H. circularisquama) were relatively unaffected compared with the other species, which suggest that the theca may offer some protection. In a recent report, Imai et al. (2002) reported finding algaecidal bacteria growing on the surface of macroalgae such as Ulva sp. (Chlorophyta) and Gelidinium sp (Rhodophyta). The number of algaecidal bacteria was estimated in the order of 105 to 106 per g (wet weight of seaweed). These bacteria were lethal against Karenia mikimotoi, Fibrocapsa japonica, and Heterosigma akashiwo. The authors proposed a novel approach of using this natural flora for preventing and controlling blooms by co-culturing macroalgae in the vicinity of aquaculture sites, or the establishment/restoration of macroalgal bed along the coast in order to seed the water with algaecidal bacteria. In Korea, the distribution and lethality of algaecidal bacteria against Cochlodinium polykrikoides were studied in a local bay. Bacterial numbers
Mitigation of effects of harmful algal blooms 185 were found in the order of 102 to 103 cells/mL (Park et al., 1998). One isolate, Micrococcus sp LG-1, displayed high lethality and selectivity for C. polykrikoides. In the US, Doucette et al. (1999) reported the isolation of two bacterial strains with the ability to kill Gymnodinium breve (now Karenia brevis). One strain, 41-DBG2, produced lethal substances that affected K. brevis and another related species, Gymnodinium mikimotoi (now Karenia mikimotoi). A broader list of bacteria and their targets is presented in Mayali and Azam (2004). Parasites Eukaryotic parasites (Sommer et al., 1984; Heaney et al., 1988; Bruning et al., 1992; Coats et al., 1996; Coats, 1999; NoreÂn et al., 1999; Erard-LeDenn et al., 2000; Park et al., 2004) have been identified as important microbial controls that can retard/inhibit bloom formation, or facilitate bloom decline. Fungal parasites can have a role in the succession of phytoplankton by preventing or delaying the occurrence of some species selectively, and can affect bloom concentration (Donk, 1989; Bruning et al., 1992). In one study, Wetsteyn and Peperzak (1991) studied infections in two diatoms, Cocinodiscus concinnus and C. granii, by the marine fungus, Lagenisma coscinodisci. The highest infection rates were 22.2± 58.3% and 7.1±41.9% for each diatom in the field, respectively. This study also showed that fungal infections were affected by water temperature. Fungal infections in freshwater species have also been described (Kudoh and Takahashi, 1990; Holfeld, 1998, 2000). Parasitic dinoflagellates have long been considered to have a significant influence on the ecology of bloom-forming dinoflagellates (Coats, 1999). Species of Amoebophrya are particularly noteworthy, as they are widely distributed in coastal environments, with infections known for numerous host taxa from Europe, North America, Asia, and Australia (Cachon, 1964; Taylor, 1968; ElbraÈchter, 1973; Nishitani et al., 1985; Cachon and Cachon, 1987; Fritz and Nass, 1992; Coats and Bockstahler, 1994; Coats et al., 1996). Amoebophrya spp. have a simple life cycle including a free-swimming infective dinospore that attaches to the host and penetrates through the host cell membrane, a vegetative trophont that grows inside the host cell, and a multinucleate, multiflagellate vermiform stage that is released upon death of the host and undergoes cytokinesis to yield hundreds to thousands of dinospores (Cachon 1964). Infections prevent reproduction of the host (ElbraÈchter, 1973; Park et al., 2002), are short in duration (2±3 days, Coats and Bockstahler, 1994; Coats and Park, 2002), and invariably result in death of the host, all of which make these parasites likely candidates for controlling host populations. Infection levels are highly variable, ranging from 80% removal) and a reduction in hemolytic activity in fish farm operation in the Chesapeake Bay (Maryland, USA). Similarly, cultures of Prorocentrium triestinum, Scrippsiella trochoidea, and Karenia digitata were killed within 15 min after exposure to 1 g O3/m3 (Ho and Wong, 2004). The treatment also reduced concentrations of ammonium and total inorganic nitrogen while dissolved oxygen levels remained within acceptable levels. Ozone treatment was also tested against Karenia brevis and its toxins (Schneider et al., 2003). Direct treatment of a K. brevis culture with 25 mg of ozone resulted in an 80% loss of cells within 10 s. All of the cells destroyed after 60 s. Similarly, free brevetoxins introduced into seawater were significantly reduced after a 10-min treatment. However, 135 mg of ozone was needed. The survival of Cyprinodon variegates in fish bioassays was inversely related to the time after ozone treatment, indicating a reduction in toxicity over time. 7.2.4 Ultrasonic devices At the 9th International Conference on Harmful Algal Blooms held in Tasmania, a device called the Aquasonic was described that reportedly kills algae through ultrasonic vibrations. These vibrations cause the cell vacuole to tear, leading to implosion of the cell. Unfortunately, there have been few data available to support these claims. Lee et al. (2001a) showed that short exposures (3 s) to ultrasonic radiation (120 W input power and 28 kHz) forced the settling of blue-green algae, owing
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to the collapse of gas vacuoles according to observation using electron microscopy. Further tests showed that photosynthetic activity decreased and growth was inhibited. At 1200 W and 28 kHz, the microcystin concentration did not increase. In a further study, Lee et al. (2001b) tested several ultrasound devices in a lake to determine whether they can limit the formation of cyanobacterial blooms. Bosma et al. (2003) demonstrated the application of ultrasound for the harvesting of microalgae by causing aggregation in a low shear stress environment. Under optimum conditions, up to 92% of the cells could be efficiently harvested with this method. Currently, there have been no reports on the use of ultrasonic devices in marine systems, although a brief mention was found in Shirota (1989) without further details.
7.3
Ethos of harmful algal bloom (HAB) control
HABs have seen a dramatic increase in frequency, magnitude, distribution, and impacts in recent years (Anderson, 1989; Hallegraeff, 1993). This has prompted considerable interest in processes that regulate formation, persistence, and decline of phytoplankton blooms, leading to research programs to understand fundamental processes underlying bloom dynamics and their impacts. New advances have also been made in mitigating blooms through new tools to aid in monitoring cells and toxins, and to provide better prediction and more rapid responses. The significant and recurring impacts of HABs would seem to justify bloom control as an additional, high-priority research topic, yet relatively little focused research on HAB control strategies has been undertaken in the US or the world (Anderson, 1997; Boesch et al., 1997; CENR, 2000). Anderson (2004) outlined many of the primary arguments and counter-arguments for and against bloom control. Among the arguments against control is the complexity and size of blooms, the remaining lack of understanding about blooms, and the potential for causing greater harm to the environment and ecosystems than the solution. Anderson (1997, 2004) argued that control would focus on regions where more is known about the dynamics of HABs ± that is the treatment approach would be informed by the available understanding of blooms in specific locations. Moreover, the author suggested that control measures would need to be more limited or targeted in space and time, instead of a broad, haphazard application (e.g., for clay dispersal). There would also be a need to evaluate the risks as well as the benefits of treatment. Clearly, bloom control will remain both an important and controversial topic in years to come.
7.4
Future trends
Of the various bloom control options, co-flocculation of HAB cells with clay is one of the most promising with regards to effectiveness and potential
Mitigation of effects of harmful algal blooms 191 applicability given the Korean and Japanese experience, although many questions remain about environmental (especially the benthic) impacts and financial cost of the treatment. In the US, however, the growing number of studies has defined the type of environmental and bloom conditions where clay treatment may be most effective or beneficial, while benthic studies offer insight into the possible impacts of treatment. This approach has received attention from other places such as Sweden (HagstroÈm and GraneÂli, 2005; Sengco et al., 2005), Hong Kong, and the Philippines. Finding new approaches in chemical control of blooms remains challenging. The need to balance the effectiveness with rapid dilution, specificity, and cost is one of the significant hurdles in this area. Recent tests with barley straw offer some promising options in marine waters. Interest in biological control is growing, especially in Japan and South Korea. Despite numerous reports of possible control agents, however, there has been reluctance to pursue these types of studies owing to the public stigma associated with introducing pathogens, even to control harmful blooms, and the anticipated high cost of producing enough biomass (e.g., grazers). One possibility that has attracted some attention is the concept of enhancing the natural flora of algaecidal bacteria that grow on the surface of macroalgae by culturing the algae along aquaculture sites (pers. comm.., I. Imai, 2006). Another area of general bloom management that is being investigated is bloom prevention. Certainly, this is an ideal method for managing HABs and as the understanding of blooms continue to grow, strategies for preventing them may also become more apparent; however, there are currently no active programs pursuing this strategy.
7.5
Sources of further information and advice
General information about harmful algal blooms · Woods Hole Oceanographic Institution (http://www.whoi.edu/redtide/) · Intergovernmental Oceanographic Commission (http://ioc.unesco.org/ iocweb/index.php)
7.6
References
and CLOERN, J.E. (1992) Trophic interactions and direct physical effects control phytoplankton biomass and production in an estuary. Limnol. Oceanogr., 37 (5): 946±955. ANDERSON, D.M. (1989) Toxic algal blooms and red tides: a global perspective. In: Okaichi, T., Anderson, D.M. and Nemoto, T. (eds), Red Tides: Biology Environmental Science and Toxicology. Elsevier, New York, pp. 11±16. ANDERSON, D.M. (1997) Turning back the harmful red tide. Nature, 388: 513±514. ANDERSON, D.M. (2004) Prevention, control and mitigation of harmful algal blooms: multiple approaches to HAB management. In: Hall, S., Etheridge, S., Anderson, ALPINE, A.E.
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Ecol., 304: 35±49. (1996) Viruses as biological control agents for blooms of marine phytoplankton. Proceedings of the Brown Tide Summit, Ronkonkoma, New York, 20±21 October, pp. 71±76. SUTTLE, C.A., CHAN, A.M. and COTTRELL, M.T. (1991) Use of ultrafiltration to isolate viruses from seawater which are pathogens of marine phytoplankton. Appl. Environ. Microbiol., 57(3): 721±726. SWARTZEN-ALLEN, S.L. and MATIJEVIC, E. (1974) Surface and colloid chemistry of clays. Chem. Rev., 74(3): 385±400. TARUTANI, K., NAGASAKI, K. and YAMAGUCHI, M. (2000) Viral impacts on total abundance and clonal composition of the harmful bloom-forming phytoplankton, Heterosigma akashiwo. Appl. Environ. Microbiol., 66(11): 4916±4920. TARUTANI, K. NAGASAKI, K., ITAKURA, S. and YAMAGUCHI, M. (2001) Isolation of a virus infecting the novel shellfish-killing dinoflagellate, Heterocapsa circularisquama. Aquat. Microb. Ecol., 23(2): 103±111. TAYLOR, F.J.R. (1968) Parasitism of the toxin-producing dinoflagellate Gonyaulax catenella by the endoparasitic dinoflagellate Amoebophrya ceratii. J. Fish. Res. Bd. Canada, 25: 2241±2245. TERLIZZI, D.E., FERRIER, M.D., ARMBRESTER, E.A. and ANLAUF, K.A. (2002) Inhibition of dinoflagellate growth by extracts of barley straw (Hordeum vulgare). J. Appl. Phycol., 14(4): 275±280. TILLMANN, U. (2004) Interactions between planktonic microalgae and protozoan grazers. J. Eukaryot. Microbiol., 51(2): 156±168. UCHIDA, T., YAMAGUCHI, M., MATSUYAMA, Y. and HONJO, T. (1995) The red-tide dinoflagellate Heterocapsa sp. kills Gyrodinium instriatum by cell contact. Mar. Ecol. Prog. Ser., 118: 301±303. WELCH, I.M., BARRETT, P.R.F., GIBSON, M.T. and RIDGE, I. (1990) Barley straw as an inhibitor of algal growth I: Studies in the Chesterfield Canal. J. Appl. Phycol., 2: 231±239. WETSTEYN, L.P.M.J. and PEPERZAK, L. (1991) Field observation in the oosterschelde (The Netherlands) on Coscinodiscus concinnus and Coscinodiscus granii (Bacillariophyceae) infected by the marine fungus Lagenisma coscinodisci (Oomycetes). Hydrobiol. Bull., 25(1): 15±21. WOMMACK, K.E. and COLWELL, R.R. (2000) Virioplankton: viruses in aquatic ecosystems. Microbiol. Mol. Biol. Rev., 1: 69±114. WYNNE, T.T., STUMPF, R.P., TOMLINSON, M.C., RANSIBRAHMANAKUL, V. and VILLAREAL, T.A. (2005) Detecting Karenia brevis blooms and algal resuspension in the western Gulf of Mexico with satellite ocean color imagery. Harmful Algae, 4: 992±1003. YOSHINAGA, I., TAKAHASHI, K. and ISHIDA, Y. (1997) Analysis of algicidal ranges of the bacteria killing the marine dinoflagellate Gymnodinium mikimotoi isolated from Tanabe Bay, Wakayama Pref., Japan. Fish. Sci., 63(1): 94±98. YU, Z., ZOU, J. and MA, X. (1994a) Application of clays to removal of red tide organisms. I. Coagulation of red tide organisms with clays. Chin. J. Oceanol. Limnol., 12(3): 193±200. YU, Z., ZOU, J. and MA, X. (1994b) Application of clays to removal of red tide organisms ± II. Coagulation of different species of red tide organisms with montmorillonite and effect of clay pretreatment. J. Chin. Oceanol. Limnol., 12(4): 316±324. YU, Z., ZOU, J. and MA, X. (1995a) Application of clays to removal of red tide organisms ± III. The coagulation of kaolin on red tide organisms. J. Chin. Oceanol. Limnol., 13(1): 62±70. SUTTLE, C.A.
Mitigation of effects of harmful algal blooms 199 and MA, X. (1995b) Study on the kinetics of clays removing red tide organisms. J. Chin. Oceanol. Limnol., 26(1): 1±6 (Chinese with unpublished English translation). YU, Z., MA, X. and YANG, X. (1995c) Study of main nutrients absorption on clays in seawater. J. Chin. Oceanol. Limnol., 26(2): 208±214 (Chinese with English abstract). YU, Z., SUN, X., SONG, X. and BO, Z. (1999) Clay surface modification and its coagulation of red tide organisms. Chin. Sci. Bull., 44(7): 617±620. YU, Z., ZOU, J.
8 Modelling as a mitigation strategy for harmful algal blooms J. Blanco, Centro de InvestigacioÂns MarinÄas, Spain
Abstract: This chapter reviews the efforts that have been made to develop dynamic models of phycotoxin accumulation in bivalves. The different processes involved in the accumulation are described and their corresponding models are explained. The chapter includes a discussion about the advantages of using models both for research and to increase the safety of marine bivalves, and some information about modelling and modelling tools. Key words: modelling, phycotoxins, paralytic shellfish poisoning, diarrhoetic shellfish poisoning, amnesic shellfish poisoning, ingestion, depuration, detoxification, biotransformation, multicompartmental, kinetics, harmful algal blooms, mitigation.
8.1
Introduction
Toxic phytoplankton outbreaks are one of the most important processes affecting the exploitation of shellfish resources (Shumway, 1989, 1990; Shumway and Cembella, 1993; FernaÂndez et al., 2003). Harvesting bans are frequent in many areas and, in the case of some combinations of shellfish species and toxins, may even last for years. This is the case, for example, of the king scallop Pecten maximus with domoic acid or Spisula solidissima, with paralytic shellfish poisoning (PSP) toxins, which remain toxic for years (Cembella et al., 1993; Blanco et al., 2002), resulting in a severe economic impact on fisheries and preventing this species from being cultured, even when it is, in other aspects, suitable for that activity. Owing to these events, the design and implementation of mitigation strategies is strongly needed. Dealing with shellfish, and especially with bivalves, the strategies should probably be based
Modelling as a mitigation strategy for harmful algal blooms 201 upon two aspects (FernaÂndez et al., 2003): (a) the knowledge of the processes involved in the accumulation of toxicity (which in turn has two aspects: accumulation of toxins and biotransformation), which would allow the design of methods to slow the incorporation of toxins into the bivalves or to accelerate the decrease of toxicity; and (b) the development of predictive capabilities that allow optimisation of the monitoring systems and fit the activities of the producers or fishermen to the predicted evolution of toxicity. As expected, in the early studies of bivalve toxicity and algal toxins, the whole toxin±shellfish system was assumed to be simple: one or few toxins for each type of toxicity behaving in the same way in all bivalve tissues. Simple systems are relatively easy to study because a series of simple hypothesis can be formulated and checked. Notwithstanding, it was soon realised that the interaction between shellfish and algal toxins is complex and species-specific, depending upon both the mollusc and the toxin involved (Shumway, 1990). Real accumulation was much more complex than initially thought, and it has been demonstrated that many toxins and derivatives are frequently involved in any kind of toxicity, that some toxins with different toxic potency could interconvert (Hu et al., 1993; Cembella et al., 1994; MoronÄo et al., 2003), and that the toxins stored in different organs/tissues of the bivalves behave in different ways (Bricelj and Shumway, 1998; Bauder et al., 2001; Blanco et al., 2002). Thus, the system, as it is perceived now, is too complex to be studied adequately by means of the formulation of simple hypotheses and by the observation of simple variables, and consequently other methods are needed ± mathematical modelling is one of those methods.
8.2
Why model the accumulation of toxins in bivalves?
Mathematical models are simplified mathematical representations of systems. Any biological system can be characterised by an input from the environment, a series ± more or less complex ± of interrelated processes and a response that is the result of applying the processes to the input and to the initial state of the system. This kind of modelling, therefore, requires: (a) the conceptual simplification of the system by means of identifying the relevant processes and descriptive variables, and (b) their description by means of sets of equations, in most cases, differential equations. There might be many reasons to model phycotoxin accumulation in bivalves, but probably the most obvious is to predict the time course accumulation and depuration. Dynamic models allow one to make predictions about the change of the state of the system with time, and then allow the estimation, for example, of some relevant features of the accumulation, as the maximum toxicity that the bivalves can reach during an outbreak or the time required for the organisms to depurate the toxins. Nevertheless, other uses can be equally important, as summarising the system behaviour, defining the main processes involved in the system and evaluating their relative importance, quantifying the rates of some
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processes that are difficult to measure by other means, or simply being used as tools to test hypotheses, and to suggest new ones. As already noted, bivalve±algal toxin systems are now known to be very complex. Just as the human brain is not able to process many variables simultaneously, a substantial simplification of the complex reality is always required for clear comprehension. To this end, mathematical models allow the behaviour of the bivalve±toxin system to be summarised by describing the kinetics with a reduced number of variables, parameters and equations. Depuration, for example, can be modelled in many cases with a simple exponential decay equation, and described with that equation plus its two parameters (initial toxicity and depuration rate). This summarised information allows for easy comparisons with other systems/organisms and situations. More complex kinetics, as those sometimes known as biphasic, can also be described with a reduced set of parameters and equations (three equations and five parameters), which, although more difficult to visualise and to compare with other situations than the two-parameter case, is still much easier to deal with than raw data. With systems even more complex than the previous ones, e.g. those including toxins that undergo biotransformation or that are transferred between different body tissues, the use of models is unavoidable, as the direct interpretation of the toxin kinetics is frequently impossible. Models are also useful to study the different processes involved in the accumulation of toxicity mainly in two ways: identification of relevant processes and quantification of rates or variables that are difficult to measure. In the first case, models, being the product of coupling all processes thought to be relevant, are good tools to detect inconsistencies or processes which had not been taken into consideration. These inconsistencies, when a non-integrating approach is used, are very difficult to detect. It is always possible to assume that, when studying part of a system, the unexpected responses that could have been observed are due to relevant variables that have not been taken into account. The use of models minimises these interpretations, as all relevant variables should be taken into account. Following Margalef (1973), models are useful because they fail, and when they fail they point to a relevant aspect of the reality that has not been included, or that has not been correctly formulated, in the model. That was the case, for example of a PSP outbreak in mussels, produced by Gymnodinium catenatum (MoronÄo et al., 1998b) in four Galician RõÂas, in which the failure of the initial modelling efforts to fit the data pointed to the fact that the differences in toxicity of the cells were substantially more important than initially thought. A new model, in which constant phytoplankton toxicity was replaced by a function of the location, fit the observed data much better than the original one. Similarly, when modelling an experiment of accumulation of gonyautoxins (GTXs) (PSP group) in mussels fed Alexandrium minutum (Blanco et al., 2003), biotransformations had to be included in the model in order to fit the observed data, because otherwise the fit was poor for several toxins. Also, the likely existence of two compartments of toxins ± labile and strongly bound ± was suggested by the lack of fit of a one-compartment model (Silvert and Cembella,
Modelling as a mitigation strategy for harmful algal blooms 203
Fig. 8.1
One (a) and two (b) compartment models fit to PSP toxicity depuration from mussels (redrawn in part from Silvert and Cembella, 1995).
1995) (Fig. 8.1). Similarly, the inability to fit their model at the start of an experimental study of domoic acid accumulation led Douglas et al. (1997) to suggest that the toxic phytoplankton involved could have an inhibitory effect on the feeding rate of the scallops studied. Because of their integrated nature, and because they can predict the system output in hypothetical conditions, models are extremely useful in evaluating the relative importance of the involved processes acting simultaneously. By means of changing the values of the parameters of the model, an analysis of sensitivity can be performed and those parameters can be ranked by their repercussion on any aspect of the system response. MoronÄo (2000), for example (Fig. 8.2) found, by this kind of analysis of a two-compartment model of PSP toxicity accumulation in mussels, that the depuration rate of the first compartment is important in regulating the maximum toxicity reached by the mussels in an outbreak, but that it is much less important than the depuration rate of the second compartment in order to regulate the time required to decrease the toxicity below the legal threshold for commercialisation. These kinds of finding are especially useful when planning new research and can contribute substantially to the correct orientation of the research for a specific aim. Models also make it possible to infer the rates of some processes or the values of some variables that cannot be easily studied individually. Toxin biotransformations, for example, are very difficult to assess. Radiolabelled toxins could be used, but obtaining radiolabelled toxins is frequently impossible and, in any case, their use would be expensive and would also require ad hoc equipment and facilities. Using mathematical models, some aspects of the biotransformation kinetics can be studied (Silvert et al., 1998; Blanco et al., 2003), obviously not with the same degree of precision as if radiotracers were used, but in an affordable and relatively easy way, useful for most practical applications. Another important use of models, already cited in the previous section, has to do with our ability to test hypotheses about complex systems and to evaluate our degree of knowledge of the toxicity accumulation system. It is difficult to test hypotheses directly and usually it is necessary to implement models describing
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Fig. 8.2 Some aspects of the sensitivity analysis of a two-compartment model of PSP toxicity accumulation. (a) The effect of the variation of the depuration rate (in day±1) of the first compartment on the accumulated toxicity (y-axis) and on the time required to drop to the maximum allowable toxicity (x-axis). (b) As (a) but with depuration rate of the second compartment.
the hypotheses as tools to determine if the inferences are consistent with existing data. Trying to determine, for example, that biotransformation and not differential depuration of the toxins, is the main process regulating toxicity (Blanco et al., 2003), would probably be an impossible task if models were not used, because of the complexity and dynamism of the processes involved.
8.3 Historical use and development of toxin/toxicity accumulation models Toxins ± and consequently their associated toxicity ± are incorporated into bivalves by means of the ingestion of toxic phytoplankton cells from seawater. After cessation of the ingestion of toxic food, the toxin concentration in the bivalves decreases progressively, falling, after some time, to undetectable levels. This time course of the toxin concentration and toxicity was recognised early in the study of phycotoxins and it was frequently split into two different phases: incorporation ± sometimes called intoxication ± and depuration or detoxification. This division into phases rather than into processes acting simultaneously was probably due, at least in part, to the lack of available tools to study the two processes independently ± toxin acquisition and toxin loss ± that overlapped in time. Nonetheless, the idea of the simultaneity of the two processes was always there, as shown by the attempts to predict the risk of contamination from the concentration of toxic cells in water. Establishing a level of toxic cell concentration with which bivalves are not contaminated implicitly assumes that toxin losses below that level are larger than toxin incorporation. In this context, even when the models have seldom been implemented in this field, implicitly, some simple ones have been used. The phase of incorporation did not garner much attention, but was generally assumed to be linear, with a slope only dependent on the concentration of toxic
Modelling as a mitigation strategy for harmful algal blooms 205 cells. Even if it was not written, the model most frequently used was (e.g. Bricelj et al., 1990): dT=dt Cw Tc F
8:1
where dT=dt velocity of incorporation of toxin; Cw toxic cell concentration in water; Tc toxin or toxicity/cell; F filtration or clearance rate. For depuration, on the other hand, several models were also implicitly used. A linear elimination with constant rate was, for example, assumed when the depuration velocity was expressed as toxicity or toxin concentration lost per day. Formally the assumed model is: dT=dt ÿK
8:2
where T is toxicity, toxin concentration or amount, t is time and K is a depuration constant. This equation describes a zero-order kinetics, in which the velocity of depuration does not depend on the toxin or toxicity accumulated by the bivalve. In some other cases, the depuration was computed using the semidisintegration constant, the slope of the regression of the logarithmically transformed toxicity on time, or it was specifically stated that depuration was directly proportional to the toxicity (Hurst and Gilfillan, 1977; Bricelj and Cembella, 1995). In both cases, a negative exponential relationship between toxicity and time was assumed. They were described by means of the following equation or similar ones: Tt T0 eÿKt
8:3
which corresponds to the differential equation dT=dt ÿKT
8:4
where T is toxicity, toxin concentration or toxin amount, t is time and K is the constant of depuration. Sub-indices, when used, indicate time. The lack of fit of a straight line to the logarithmically transformed data led also to propose biphasic kinetics of depuration, in which two different lines had to be fitted to different portions of the depuration curve (Lassus et al., 1989, 1994; Bricelj et al., 1991). Again the lack of tools to deal with the relative complexity of the kinetics probably forced the adoption of models that were conceptually incorrect, but that allow a reduced number of descriptors (rates) of the time course of the depuration to be determined. In 1992, the first dynamic model of toxin accumulation (domoic acid) was formulated and implemented by Silvert and Subba Rao (1992). This model was soon followed by others of accumulation of other kinds on toxins/toxicities (Blanco et al., 1995, for diarrhoeic shellfish poisoning (DSP) toxins, and Silvert and Cembella, 1995, for PSP toxins), that were initially simple but that increased their complexity with time, as will be shown in the next section.
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8.4 Models of the kinetics of accumulation and transformation of toxins in shellfish 8.4.1 Processes In order to model the accumulation of phycotoxins in bivalves there are several groups of processes that have to be taken into account. The first two, the most obvious, are those that represent the toxin gain and loss by shellfish: toxin ingestion and toxin depuration or elimination. Two others, which are especially important in many cases and that do not involve gain or loss of toxin by the bivalves, are the anatomical redistribution and transformation of toxins. In the discussion that follows, three different groups of models of accumulation will be considered together, for the sake of simplicity. Each group of models deals with a different characteristic of toxic shellfish: toxicity, toxin concentration and toxin content. Obviously the models are not exactly the same, but the early studies focused on toxicity mainly because it was the final objective of prediction and also because there was not sufficient analytical capability to study toxins themselves rather than their effects (many studies have been carried out by means of mouse bioassay). Toxin content is probably the response that should be studied, as it depends only on the amount ingested and on no other variable of the shellfish, while toxin concentration depends not only on the toxin ingested but also on the biomass of the soft tissues of the shellfish, and toxicity depends on the same variables as toxin concentration plus on the proportion and toxic power of each specific toxin. Nevertheless, when biotransformations and weight changes are small enough to be neglected, then the models for the three responses are the same. 8.4.2 Models of toxin incorporation The incorporation of toxins (and toxicity hereafter) into shellfish is usually modelled by assuming both constant rates of toxic cells ingestion and constant cellular toxin contents and usually under the mathematical form, similar to equation [8.1] (Silvert and Subba Rao, 1992; Silvert and Cembella, 1995): dTox=dt feeding rate[cells]water [toxins]cell
8:5
in which Tox toxin amount; feeding rate the rate at which the bivalves withdraw particles from water; [toxins]cell cellular toxin contents of the phytoplankton; and [cells]water concentration of toxic phytoplankton cells in water, which usually varies with time. This model takes into account all ingested toxins wherever they are absorbed through the walls of the digestive systems or where they remain in the gut lumen until they are egested with faeces. When the total amount of toxin that can be accumulated in the gut is considered to be negligible in relation to that accumulated and there is interest, therefore, only on the toxins that pass the gut walls, an additional factor to correct for the non-absorbed toxin can be added: the toxin absorption efficiency (Blanco et al., 1995; MoronÄo et al., 1998b). This factor is equivalent to the absorption efficiency for organic matter that is widely used in physiological studies of bivalves:
Modelling as a mitigation strategy for harmful algal blooms 207 dTox=dt feeding rate [Cells]water [Toxins]cell AE
8:6
where AE is the absorption efficiency. Toxin absorption efficiency is usually assumed to be constant even when, in the case of organic matter, it has been found to be dependent upon the food quality ± which determines the gut passage time (Hawkins et al., 1990) ± and also on an equivalent variable (toxin per unit of volume of seston) in the case of PSP toxins (MoronÄo et al., 2001). This relationship can be included in models by the substitution of the AE constant by a power function of the ratio toxin/seston volume, similar to the one used to predict the absorption efficiency of organic matter (see Hawkins et al., 1990; Navarro et al., 1992; Iglesias et al., 1992). Notwithstanding, owing to the relatively narrow range of variation of AE (usually 40±70% for organic matter, and frequently narrower), its inclusion in the models sometimes does not contribute to improve them substantially, at least from a quantitative point of view. This approach of modelling toxin incorporation gives, in general, good results but has some limitations. The first is a practical one: it is very difficult to correctly sample phytoplankton populations because they are frequently strongly heterogeneous both in space and time. As an example, Silvert and Cembella (1995) found two typical situations while modelling PSP accumulation in mussels, derived from not detecting toxic phytoplankton and from assuming that the populations detected at a precise time are representative of a timespan longer than is the case (Fig. 8.3). MoronÄo (2000) also found, when modelling PSP accumulation in mussels from the RõÂa de Vigo, that the fit of the model obtained in one sampling station was substantially improved when the phytoplankton data corresponding to a nearby (3 km apart) station were used as input for the model, a problem that is very likely linked to the displacement of a front during the tidal cycle. Toxin content/toxicity per phytoplankton cell is also variable in many cases (MoronÄo et al., 1998b, and references therein), and the same is true for filtration/clearance rate, which can be affected by the toxic population because
Fig. 8.3 PSP toxicity in mussel digestive gland (circles) and in suspended matter (mostly phytoplankton), as well as the output of a two-compartment model of toxicity accumulation. Initial anomalous levels of toxicity in mussels (NDP) were probably due to non-detected toxic plankton. A clear overestimation of mussel toxicity by the model (OEP) was probably due to an overestimation of plankton toxicity for that particular period of time (redrawn from Silvert and Cembella, 1995).
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of the specific characteristics of the cells that can enhance or reduce the filtration and clearance rate (reviewed in Shumway and Cucci, 1987; Gainey and Shumway, 1988; Shumway, 1989, 1990; Shumway et al., 1990; Shumway and Cembella, 1993; Blanco-PeÂrez, 2001), because of the amount/type of suspended matter (Bayne et al., 1987, 1993; Iglesias et al., 1992; Navarro and Widdows, 1997), or even because of the environmental conditions associated with the toxic phytoplankton blooms. As an example, Silvert and Subba Rao (1992), suggested that the alteration of the filtration rate might have been the cause of the poor fit of one model to two consecutive Pseudo-nitzschia (domoic acid producer) blooms. Douglas et al. (1997) were also unable to fit a model of accumulation of domoic acid in scallops without assuming that the filtration rate was substantially reduced during the first days of feeding. MoronÄo et al. (2002) could fit a model of PSP toxin accumulation in mussels only after assuming an exceptionally low clearance rate, probably associated with the low salinity water in which the causative species Alexandrium minutum populations developed. Recently, Baron et al. (2006) have replaced the fixed ingestion rate, used in most studies to date, with Qtox, a rate of toxin uptake in an experimental tank, which is dependent (among other variables) on the water flow through the system, on the toxicity and concentration of the toxic phytoplankton in water, and on the decrease of in vivo fluorescence owing to the passage of the water through the system. If we are to obtain good predictions of the attainable toxicity of the episodes, more accurate predictions of the incorporation of toxins are needed. 8.4.3 Models of depuration One-compartment models As already noted, the first developments arose from the need to summarise the depuration kinetics with one or a few parameters, mostly in order to simplify comparisons. Initially, zero-order kinetics, in which the toxin loss did not depend on the toxin amount or concentration in the bivalve, were used. In those cases, the toxin decrease was described by a straight line of the type shown in equations [8.1] and [8.2]. Soon it was realised that those zero-order equations did not correctly describe the depuration kinetics, and a slightly more complex model of depuration was proposed. The new models (Fig. 8.4) (Silvert and Subba Rao, 1992; Blanco et al., 1995; Silvert and Cembella, 1995) assume that toxin levels in shellfish decrease in an amount that is proportional to the accumulated toxin. It can be described by equation [8.4], and its derived kinetics is a first-order one because the change of toxin concentration is dependent upon the toxin level. This equation is the most used, by far, in mathematical models of toxin depuration. Two-compartment models Frequently, the first order model of depuration does not correctly fit the data because an apparently biphasic depuration takes place (Fig. 8.1). In those cases, the initial steps of depuration seem to proceed rapidly and the final ones slowly.
Modelling as a mitigation strategy for harmful algal blooms 209
Fig. 8.4
Conceptual representation of one- and two-compartment models (from Blanco et al., 1997).
This situation has been modelled by means of two-compartment models, in which it is assumed that there are two different pools of toxins, instead of only one, and that each pool is differently bound to bivalve tissues, and consequently it has a different depuration rate (Fig. 8.4). Usually, the first compartment is the one in which toxins are weakly bound to the tissues and is also the compartment that receives the input of toxins from the plankton ingested. The second compartment is the one in which the toxins are strongly bound to the tissues and that receives them from the first compartment. These two-compartment models have usually been described mathematically by a set of two differential equations describing a first-order depuration of the toxins corresponding to that compartment, and also a first-order transfer from the first (fast depurating) to the second (slowly depurating) compartments (Silvert and Cembella, 1995; Blanco et al., 1997, 1999; MoronÄo et al., 1998a,b) (Fig. 8.1): dTox1 =dt ÿTR12 Tox1 ÿ D1 Tox1 dTox2 =dt TR12 Tox1 ÿ D2 Tox2
8:7
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where Toxn is the amount or concentration of toxin (or toxicity), Dn is the depuration rates and TR12 is the transfer rate between compartments 1 and 2. Sub-indexes of Tox and D indicate the compartment. Even when the model can be easily understood, its translation from the mathematical to the organismic level has to be done very carefully. There is no clear and constant correlation between model compartments and organs or tissues in shellfish. The correlation varies with the stage of the toxic outbreak and with the conditions in which it develops. For example, if phytoplankton were highly toxic and the biomass of the accompanying species were low, then a large amount of toxin (or toxicity) will be concentrated in the shellfish gut ± in a large proportion as intact cells ± being therefore, mostly not bound to the tissues. The evacuation of these toxins with faeces would be fast and, in that case, gut lumen would correspond to the first compartment and all other organs and tissues would constitute the second one. In contrast, if phytoplankton had low toxicity and the biomass of accompanying species were large, then, after some time of accumulation, the toxicity in the gut would be quantitatively negligible in relation to that accumulated into shellfish tissues, and consequently it would not be correspond to the first compartment, which would probably correspond reasonably well to the whole digestive gland. Also, when depuration is modelled starting some time after the disappearance of the toxic phytoplankton populations from the water, it is likely that the gut content does not correspond to the first compartment, as after 3 days, in most bivalves, gut evacuation is practically complete or the amount that the gut lumen can store is negligible in relation to the toxin already accumulated in the tissues. In those situations in which the contribution of gut content can be neglected, the correspondences might also be diverse, because it is dependent on species and toxins. In the case of PSP toxins, which, in most species, are distributed among all organs or tissues, digestive gland, that is the fastest depurating organ, may be assumed to be the most important contributor to the first compartment, and all other organs or tissues can be assimilated to the second one. Notwithstanding, in the clam Saxidomus giganteus (Kvitek and Beitler, 1988) or Acantocardia tuberculata (Berenguer et al., 1993), which retain the PSP toxins strongly in the siphon and foot, respectively, these organs would clearly have the heaviest weight in defining the second compartment and therefore the remaining tissues would constitute the first one (even when most of them do not acquire toxins from the environment). In the case of DSP toxins, whose anatomical distribution is usually restricted to the digestive gland, the two compartments would be included in that organ, and consequently they probably represent different cellular types or different areas inside the organ. Assigning a mathematical compartment to an anatomical compartment is therefore a task that should be made with caution, after analysing in detail the behaviour of the particular group of toxins being studied in the species of interest.
Modelling as a mitigation strategy for harmful algal blooms 211 8.4.4 Models with external variables Environmental control or the effects of other external or internal variables have been included in models mainly by replacing some parameters of the above cited models by equations which compute them as a function, usually linear, of the new variables to be taken into account. Silvert and Subba Rao (1992) implemented the effect of temperature on domoic acid depuration in this way, and Blanco et al. (1997, 1999) implemented the effects of temperature, salinity, fluorescence (as an index of phytoplankton abundance), underwater light transmission (as an index of suspended matter) and body weight, also in the same way. More recently, but following the same methodology, Yamamoto et al. (2003) included in their model several variables such as water temperature and shell length of the molluscs. These types of model can be useful in several ways: they can be used to simplify the description of the depuration process or they can be used as a tool to estimate the actual effect of the included variables, by allowing their coefficients to vary while optimising the fit of the model output to the data. In the last case it is especially important to know precisely the kind of model to which the new variables are added, because, as shown by MoronÄo et al. (1998a), an effect can be attributed to the variables only because they are able to partially correct the response of an incorrectly chosen base model (Fig. 8.5). 8.4.5 Multicompartment models Multicompartment models are the logical expansion of two-compartment models in which the correspondence between mathematical and anatomical compartments is forced by asigning mathematical compartments to the most important organs. They are especially useful in the organism with organs that can be easily dissected and analysed as, for example, scallops or large clams. Two-compartment models are usually enough to model depuration with good precision. Notwithstanding, some species have organs or tissues with special characteristics, e.g. their commercial value, as is the case of the adductor muscle and the gonad of scallops, or because they retain strongly some toxins, as the foot and siphon of some clams. To manage the toxic populations of those species, more complex ± multicompartmental ± models are needed, because the depuration kinetics of each organ or main body fraction has to be understood. The implementations have been made in a way similar to that used for twocompartment models, using first-order reactions to describe toxin transport between organs/fractions and depuration of toxins from each of them, both, in the case of domoic acid in scallop Pecten maximus (Blanco et al., 2002), and PSP toxins in the same species (Blanco et al., unpublished results). Using a simplified notation, the equations defining the model have the general form: dTn =dt Acqn ÿ TRnÿm TRmÿn ÿ Depn
8:8
where Tn is the toxin in the compartment n, Acqn is the toxin acquisition from the environment of the compartment n; TRnÿm and TRmÿn are the losses of
Fig. 8.5 Fitting of one- and two-compartment models to PSP and DSP toxin concentration in mussels (whole tissue and digestive gland, respectively) including and not including in the models the effect of the environmental conditions and body weight on the depuration rate. It can be observed that using one-compartment models, the external variables seem to have some effect, but not when two-compartment models were used (from MoronÄo et al., 1998a).
Modelling as a mitigation strategy for harmful algal blooms 213 toxins from compartment n to m, and the toxin gain of compartment n from m; and Depn is the toxin lost by compartment n. These kinds of models have been used to try to describe or predict the depuration and accumulation of toxins by organisms, but also to try to identify the functions of different organs in relation to toxin depuration or the processes involved in the anatomical redistribution of some toxins. Blanco et al. (unpublished results), for example, implemented three different models of PSP accumulation for the king scallop (Fig. 8.6), trying to evaluate the relative importance of the different organs in the kinetics of those toxins. The first model assumed that all organs are able to depurate the toxins, the second assumed that only digestive gland and kidney are able to do that, and finally, a third one that was identical to the second, but that additionally assumed that toxins can be directly lost with the biomass of the gonad during spawning. The sets of equations describing the three models were very similar. Those corresponding to the third one are: dTDG =dt CR TFt AE ÿ DDG TDG ÿ TRDGÿAM TDG ÿ TRDGÿK TDG ÿ TRDGÿF TDG ÿ TRDGÿGM TDG ÿ TRDGÿG TDG 8:9 dTAM =dt TRDGÿAM TDG ÿ TRAMÿK TAM dTK =dt ÿDK TK TRDGÿK TDG TRAMÿK TAM TRFÿK TF TRGMÿK TGM TRGÿK TG dTF =dt TRDGÿF TDG ÿ TRFÿK TF dTGM =dt TRDGÿGM TDG ÿ TRGMÿK TGM dTG =dt TRDGÿG TDG ÿ TRGÿK TG ÿ BLG Ta =Wa where T is the PSP toxin contents, CR the clearance rate of the bivalves, TFt the toxin concentration in phytoplankton, AE the absorption efficiency, D the depuration rate, TR the transference rate, BLG the biomass loss rate at time t, and Ta =Wa is the toxin concentration in the anatomical compartment x. The subindexes indicate the body fraction or the pairs of body fractions to which the toxin contents, or the transfer rates correspond. DG digestive gland, AM adductor muscle, K kidneys, F foot, GM gills and mantle, G gonad. The fit of the first and second models were found to be very similar to each other (Fig. 8.7), but, taking into account that it is difficult to work out an efficient depuration mechanism from gonad, foot or adductor muscle, it seems clear that depuration takes place mainly through digestive gland and kidneys, making clear that the actual cause of the high PSP toxin concentrations found in the scallop kidneys (Lassus et al., 1989) is the input of toxins from other organs, as was hypothesised by Lassus et al. (1992). The third model was similar to the second one with the difference that a toxin loss, equivalent to the loss of gonadal mass, was included. The fit of the model to the data corresponding to the gonad was substantially improved (Fig. 8.7), strongly suggesting that the implemented mechanism is in effect. It is clear that multicompartmental models can be useful tools, even though
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Fig. 8.6 Conceptual multicompartmental model of domoic acid accumulation in scallop. In the model A, all organs can depurate toxins, in model B, only digestive gland and kidney can depurate; and in model C, a loss of toxin from the gonad, that was proportional to its biomass loss, was added to model B.
they cannot solve all situations. Blanco et al. (2002), for example, using a model of anatomical compartmentalisation of domoic acid in the king scallop species, estimate transfer rates between organs of nearly zero, but this finding can be explained by two different mechanisms that the model cannot discriminate: a true absence of transfer between organs or by a very fast transfer of toxin in relation to depuration.
Modelling as a mitigation strategy for harmful algal blooms 215
Fig. 8.7 Fitting of the three models described in Fig. 8.6 to the PSP burden of several body parts of the king scallop. It can be observed that, in most organs, A and B do not differ substantially, that C fits the gonad data much better that the two other models, and that food data cannot be adequatelly modelled.
Another situation in which the estimation of transfer rates by means of multicompartmental models is not reliable happens when the difference of toxin burden between compartments is very large. For example, when transfers of domoic acid to and from the foot of the king scallop ± which contains two or three orders of magnitude less toxin than other organs ± are to be estimated, the actual toxin levels in the foot are smaller than the error with which those in the other organs are estimated, and that makes it impossible to obtain reliable estimations of the involved rates by fitting models to the data. 8.4.6 Models of transformations between toxins It is clear that the most important processes to be considered when modelling toxin accumulation are those that involve toxin gain and/or loss from a particular
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organism or body fraction. Notwithstanding, owing to the metabolic activities of the organisms that accumulate the toxins, to the different physical and chemical environment inside those organisms with relation to the producers, and also to the activity of the phytoplankton enzymes released from the cells during their breakdown into the digestive tract of the shellfish, toxins undergo transformations. Toxin groups usually comprise different compounds that share a common base structure, but differ in the spatial disposition of some radicals or in the kind and number of the radicals. These chemical differences impart different levels of toxicity to the impacted organisms. Transformations, therefore, become especially important for two reasons: they produce changes in toxicity without any change in the toxin content (on a molar basis), and they make it virtually impossible to trace the accumulation (or any process involved in it, such as depuration) of a particular form when it can be interconverted with other forms of the same group of toxins. Several transformations have been modelled to date. In the case of PSP toxins (Fig. 8.8) epimerisation, reductions and decarbamoylation were modelled during the accumulation and depuration by the clam Spisula solidissima (Silvert et al., 1998), and the two first processes in mussels Mytilus galloprovincialis (Blanco et al., 2003), in both species, after the ingestion of Alexandrium cells. Some transformations were considered to be unidirectional (reductions and decarbamoylations) and the other (epimerisation) bidirectional, but all of them were assumed to be first-order reactions. In general, the two models can be described as: dGTX1 =dt ÿK GTX1 E4ÿ1 GTX4 ÿ E1ÿ4 GTX1 ÿ R1ÿ2 GTX1 ÿ DC1 GTX1 dGTX2 =dt ÿK GTX2 E3ÿ2 GTX3 ÿ E2ÿ3 GTX2 ÿ R1ÿ2 GTX1 ÿ DC2 GTX2 dGTX3 =dt ÿK GTX3 E2ÿ3 GTX2 ÿ E3ÿ2 GTX3 ÿ R4ÿ3 GTX2 ÿ DC3 GTX3 dGTX4 =dt ÿK GTX4 E1ÿ4 GTX1 ÿ E4ÿ1 GTX4 ÿ R4ÿ3 GTX2 ÿ DC4 GTX4
8:10
dDcGTX1 =dt ÿK DcGTX1 DC1 GTX1 dDcGTX2 =dt ÿK DcGTX2 DC2 GTX2 dDcGTX3 =dt ÿK DcGTX3 DC3 GTX3 dDcGTX4 =dt ÿK DcGTX4 DC4 GTX4 Here, the equations that defined the basic depuration model were complemented with the different transformations that were assumed to be proportional to the amount of the toxin that undergoes them. Epimerisations were described by Enÿm GTXn , reductions by Rnÿm GTXn , and decarbamoylations by DCn GTXn , where Enÿm are the rates of epimerisation between the toxins indicated in the sub-index, Rnÿm , the reduction rates and DCn the decarbamoylation rates.
Modelling as a mitigation strategy for harmful algal blooms 217
Fig. 8.8 Some transformations of PSP toxins that have been modelled.
In the clam Spisula solidissima, the model had a good fit with the observed data, with the exception of the -epimers (GTX1 and GTX2) (Fig. 8.9), which seem to be overestimated in the central portion of the incorporation. It seems that this strong trend in the case of GTX2 led Silvert et al. (1998) to suggest the possibility that GTX2 were more strongly retained than other toxins of the same group. In the experiment with Mytilus galloprovincialis, Blanco et al. (2003) fit, among others, two models that shared the mathematical formulation of [8.10] but that differ in that in the first model all the transformation rates were zero (no transformations) and in the second one transformations were allowed, but depuration rate was the same for all toxins. The fit of the first model (Fig. 8.10) was not good even when the estimated depuration rates of the four toxins studied were very different from each other. In contrast, the model that included transformations fit the real data well, suggesting that toxin transformation is one of the most important processes in the regulation of the toxin accumulation kinetics, in this case. An additional model allowing transformations but also toxin-dependent depuration rates did not improve the fit and the obtained estimated rates of depuration were very close for the four toxins. Transformations of DSP toxins have also been modelled. The first attempts tried to model the depuration kinetics of okadaic acid and DTX2, including the expected acylation of okadaic acid in bivalves to produce low-polarity derivatives generically known as DTX3 (FernaÂndez et al., 1998). Again first-
Fig. 8.9 Fit of an accumulation model including transformations between PSP toxins in the clam Spisula solidissima, showing the influence of epimerisation (a) and reduction and decarbamoylation (b) (from Silvert et al., 1998).
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Fig. 8.10 Fit of two models of accumulation of PSP toxins in mussels. The model in (a) assumed that toxins depurate at different rates but that there are no transformations. The model in (b) assumed a common depuration rate for all toxins but the existence of epimerizations and reductions (from Blanco et al., 2003).
order reactions were used to describe all processes, but a two-compartment model was needed. dTox1 =dt ÿDT1 Tox1 ÿ T Tox1 ÿ A1 Tox1
8:11
dTox2 =dt ÿDT2 Tox2 ÿ T Tox1 ÿ A2 Tox2 dAcyl1 =dt ÿDTA1 Acyl1 A1 Tox1 ÿ TA Acyl1 dAcyl2 =dt ÿDTA2 Acyl2 A2 Tox2 TA Acyl1 where Tox free toxin concentration; Acyl concentration of 7-acyl derivatives of the free toxins; DT depuration rate of the free toxins; DTA depuration rate of the 7-acyl derivatives; T and TA transference rate of free toxins and 7-acyl derivatives, respectively, from compartment 1 to compartment 2; A acylation rate of the free toxins. Numerical sub-indexes indicate the compartment. The model fit the data in the study well, but some striking features were found (Fig. 8.11). The estimated depuration rate of the low-polarity derivatives
Modelling as a mitigation strategy for harmful algal blooms 219
Fig. 8.11 Conceptual model of DSP toxins accumulation, including ingestion and hydrolysis of low-polarity derivatives (from MoronÄo et al., 2003).
was very high and the estimated acylation rates were very low, thus raising doubts about the real origin of the low-polarity derivatives and, consequently, about the mechanisms involved in the change. It seems likely that the lowpolarity derivatives at the beginning of the experiment were not produced by the mussel but ingested with plankton and quickly hydrolysed to their free acid forms. The high apparent depuration rate derived from the fact that two processes were involved, the toxin elimination from mussels and the hydrolysis to the free acid form. The use of models, and in this case the odd value of some parameters, forced us to consider involved processes that would probably not taken into account otherwise.When a new experiment was modelled (Fig. 8.11) (set of equations 8.12), including the acquisition of low polarity forms from plankton and their corresponding hydrolysis (MoronÄo et al., 2003), the high depuration rate was only partially corrected and some evidences of undetected
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compounds still persisted, showing the ability of models to reveal unknown processes or toxin pools involved in the accumulation or depuration of toxins. dOA1 =dt
N TOA F AE ÿ OA1 D1OA ÿ OA1 A1 ÿ OA1 T12OA CF1 H dCF1 =dt
N TCF F AE ÿ CF1 D1CF H ÿ CF1 T12CF dOAA1 =dt OA1 A ÿ OAA1 D1OAA ÿ OAA1 T12OAA dCF1 dOAA1 dLPCF1 =dt dt dt dOA2 =dt OA1 T12OA ÿ OA2 A2 ÿ OA2 D2OA CF2 H2 dCF2 =dt CF1 T12CF ÿ CF2 D2CF ÿ CF2 H2
8:12
where OA accumulated OA; N Dinophysis acuminata concentration (cells L±1); TOA OA toxin content per D. acuminata cell; F filtration rate (L day±1); AE assimilation efficiency; DOA OA detoxification rate; A OA acylation rate; CF accumulated conjugated forms; H hydrolysis rate; TCF contents of conjugated forms per D. acuminata cell; DCF detoxification rate of conjugated forms; OAA OA acyl-derivatives; DOAA detoxification rate of OA acyl-derivatives; LPCF accumulated low-polarity conjugated forms.
8.5 Applications of modelling for improved shellfish safety and quality One of the most challenging aspects of monitoring systems is to minimise the cost/benefit ratio. One basic rule to maintain low costs is to sample and analyse shellfish only where necessary. Through their ability to predict, models can, and do, contribute substantially to that. For example, in the king scallop, which accumulated 1000 g/g of tissue and assuming a depuration rate of 0.01 day±1 and a one-compartment model, there is no need to sample for at least 326 days, which probably means a substantial reduction of the effort and cost of the monitoring, specially if several locations are affected. If the resources were assigned a priori and the monitoring cost cannot be substantially reduced, then the efforts could be focused on other aspects of the control, such as increasing the spatial resolution of the system or increasing the sampling frequency when the opening level is nearly reached, trying to reduce the impact on harvest and production as much as possible. Yamamoto et al. (2003) evaluated the risk of using toxicity accumulation models to reduce the monitoring effort during the season with low risk of proliferation of Alexandrium populations in Hiroshima Bay, concluding that models can replace direct quantification of toxicity in oysters. Nevertheless, this approach has to be taken with caution in other areas in which the spatial or temporal heterogeneity of the phytoplankton is high, as the measured input to the model might be incorrect.
Modelling as a mitigation strategy for harmful algal blooms 221 Models can also be used as monitoring systems to predict when increases of the risk of intoxication take place due to changes in shellfish or environmental state. It is possible, for example, to predict the change in toxin concentration (and toxicity) of shellfish after a spawning event. If the toxin level was close to, but below, the legal banning threshold, then, if the model predicts an increase in concentration, the toxicity of the shellfish populations should be re-analysed and sampling should be intensified. This situation can probably be observed in domoic acid in the king scallop (Pecten maximus) as its loss with gametes is a small proportion of the toxin that the gonad contains (Blanco et al., unpublished results), or in most species with DSP toxins, because these toxins are nearly always confined to the digestive gland of the bivalves (Stabell et al., 1992; Pillet et al., 1995; Bauder et al., 2001; Hess et al., 2003). If seston concentration in water is low or if it has low quality as food (low digestibility, low percentage of organic matter), shellfish (especially bivalves) lose biomass and consequently the toxin concentration increases (the same amount of toxin in less biomass). This effect can also be predicted if models of growth are implemented coupled with those of toxin accumulation, and consequently, monitoring system can have information available about possible toxin concentration increments due to this cause. Evaluating the consequences of biotransformations is also a contribution of models to the reduction of the risk of intoxication. Both during the process of digestion and after the uptake of the toxins into shellfish tissues, toxins and their derivatives are transformed, producing changes in toxin concentration and toxicity. In some cases, owing to this process, toxicity can increase, especially during the early phases of the depuration, just after the toxic phytoplankton disappeared from the water (Blanco et al., 2005). Models can predict the increase of toxicity and, as in the cases of spawning and biomass loss, they can suggest the needed changes in sampling plans to adequately monitor the shellfish resources.
8.6
Future trends
Models as monitoring tools have great potential; however, to be able to use models more efficiently for prediction, the key issue is not a modelling development but a technological one. More precise estimates of the input to the actual accumulation models are needed. Models need a reliable input, and now, in most cases, the sampling capability is not enough to cope with dynamic phytoplankton populations. New tools to monitor phytoplankton population continuously, or at least frequently, (e.g., fast submersible video-cameras, and/or automated systems of phytoplankton species recognition by image analysis or flow cytometric techniques) have been developed and are starting to be used (Culverhouse et al., 2002, 2003; Sieracki et al., 1998; Lavrentyev et al., 2004; Babin et al., 2005; Anonymous, 2005) and will allow us to model toxin accumulation with much more precision.
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Apart from this technological aspect, there are also other possible means of development. Expanding the actual models by coupling them to other models predicting the toxin concentration, or the toxicity, per cell from the cell number and the environmental conditions would be an important step for the predictive capability of the models. There are only a few models targeting the cellular toxin content (John and Flynn, 2002; Flynn, 2002). Coupling the actual models with those of phytoplankton processes can also be done without previously implementing those of cellular toxin contents, but they probably would not be very useful from a practical standpoint. The development of one or some of the previous models may be of a different nature from those described in this chapter, and especially if the development of dynamic models is very difficult, other kinds of predictive tools, e.g. the combination of several techniques of artificial intelligence, should be investigated.
8.7
Sources of further information and advice
8.7.1 General information There are many good modelling books and I recommend Meerschaert (1999) which gives a good introduction to the subject in a comprehensible manner. More focused on the kind of models required in toxin accumulation than the previous one, Wastney et al. (1999) includes many examples using the free software WinSAAM. Hakanson and Peters (1995) is an interesting book when looking for mixed strategies that allow predictive responses of the target system. Finally, Hilborn and Mangel (1997) give a good introduction to the way in which models can help us to understand natural processes. Much more centred in modelling phycotoxin kinetics, but also including many other modelling activities, the personal web site of William Silvert is a very useful source of consistent information, http://ciencia.silvert.org/models/ and developments by the author's team can also be freely accessed from the web site http://www.cimacoron.org/Cimacoron/EpisodiosToxicos/index.html. 8.7.2 Software There are many options when choosing software tools to model toxin accumulation. Classical programming languages can be used, FORTRAN, C, BASIC, DELPHI, mostly proprietary or open source alternatives such as Python or its science-focused modification SciPy, however, the efficient use of these tools requires a high degree of fluency that is difficult to achieve if programming is not a frequent activity. Other high-power, but less demanding tools, are MATLAB and its relatives SciLab and Octave. MATLAB (Mathworks) is a proprietary high-level programming language complemented with a number of toolboxes designed to facilitate the process of building and fitting models: http://www.mathworks.com,
Modelling as a mitigation strategy for harmful algal blooms 223 h t tp : / / w w w . m a t h w o r k s . c o m /p r o d u c ts / m a t l a b / d e m o s . h t m l , h t t p : / / www.mathworks.com/products/optimization/ http://www.mathworks.com/ products/simulink/demos.html. SciLab (INRIA, ENPC) is a free open source software package that has, additionally to the high-level language, some tools that are similar to MATLAB: http://www.scilab.org/, http://www.scilab.org/com/products/optimization/. Octave is an open source package that clones the basic features of MATLAB but that does not have the functionality that can be obtained by using the toolboxes. Stella http://www.iseesystems.com/, PowerSim, http://www.powersim.com/ products/studio_new.asp, and ModelMaker, http://www.modelkinetix.com/ modelmaker/index.htm are easy-to-use, general-purpose modelling environments. Berkeley mandonna, http://www.berkeleymadonna.com and SAAM II, http://depts.washington.edu/saam2/, can be used also as a general purpose modelling environments but the first has a special interface available that generates kinetic equations and the second has an interface based on compartmental models. All those packages are proprietary. WinSAAM, http:// www.winsaam.com/, is a free software package, specially focused on kinetic simulation.
8.8
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Modelling as a mitigation strategy for harmful algal blooms 225 experts make mistakes? A comparison of human and machine identification of dinoflagellates. Mar. Ecol. Prog. Ser., 247, 17±25. DOUGLAS, D.J., KENCHINGTON, E.R., BIRD, C.J., POCKLINGTON, R., BRADFORD, B. and SILVERT, W. (1997) Accumulation of domoic acid by the sea scallop (Placopecten magellanicus) fed cultured cells of toxic Pseudo-nitzschia multiseries. Can. J. Fish. Aquat. Sci., 54, 907±913. Â NDEZ, M.L., MõÂGUEZ, A., MORON Ä O, A., CACHO, E., MARTõÂNEZ, A. and BLANCO, J. (1998) FERNA Detoxification of low polarity toxins (DTX3) from mussels Mytilus galloprovincialis in Spain. In: Reguera, B., Blanco, J., FernaÂndez, M.L. and Wyatt, T. (eds), Harmful Algae, pp. 449±442. Santiago de Compostela: Xunta de Galicia and Int. Oceanographic Com. of Unesco. Â NDEZ, M.L., SHUMWAY, S.E. and BLANCO, J. (2003) Management of shellfish FERNA resources. In: Hallegraeff, G.M., Anderson, A.D. and Anderson, D.M. (eds), Manual on Harmful Marine Microalgae, pp. 657±692. Paris: Unesco Publishing. FLYNN K.J. (2002) Toxin production in migrating dinoflagellates: a modelling study of PSP producing Alexandrium. Harmful Algae, 1(2), 147±155. GAINEY, L.F.J. and SHUMWAY, S.E. (1988) A compendium of the responses of bivalve molluscs to toxic dinoflagellates. J. Shellfish Res., 7, 623±628. HAKANSON, L. and PETERS, R.H. (1995) Predictive Limnology, Amsterdam, The Netherlands: SPB Academic Publishing. HAWKINS, A.J.S., NAVARRO, E. and IGLESIAS, J.I.P. (1990) Comparative allometries of gutpassage time, gut content and metabolic faecal loss in Mytilus edulis and Cerastoderma edule. Mar. Biol., 105, 197±204. HESS, P., MCMAHON, T., SLATTERY, D., SWORDS, D., DOWLING, G., MCCARRON, M., CLARKE, D.,
and O'CINNEIDE, M. (2003) Use of LC-MS testing to identify lipophilic toxins, to stablish local trends and interspecies differences and to test the comparability of LC-MS testing with mouse bioassay: an example from the Irish Biotoxin Monitoring Programme 2001. In: Villalba, A., Reguera, B., Romalde, J.L. and Beiras, R. (eds), Molluscan Shellfish Safety. Proceedings of the 4th International Conference on Molluscan Shellfish Safety, pp. 57±66. Santiago de Compostela: ConsellerõÂa de Pesca e Asuntos MarõÂtimos, Xunta de Galicia and IOC of Unesco. HILBORN, R. and MANGEL, M. (1997) The Ecological Detective: Confronting Models with Data, Princeton, NJ: Princeton University Press. GIBBONS, W., SILKE, W.
HU, T., DEFREITAS, A.S.W., DOYLE, J., JACKSON, D., MARR, J., NIXON, E., PLEASANCE, S.,
and WRIGHT, J.L.C. (1993) New DSP toxin derivatives isolated from toxic mussels and the dinoflagellates, prorocentrum lima and prorocentrum concavum. In: Smayda, T. and Shimizu, Y. (eds), Toxic Phytoplankton Blooms in the Sea, Developments in Marine Biology, Vol. 3, Amsterdam: Elsevier, pp. 507±512. HURST, J.W. and GILFILLAN, E.S. (1977) Paralytic shellfish poisoning in maine. Tenth National Shellfish Sanitation Workshop. Hunt Valley, Maryland, 1977. IGLESIAS, J.I.P., NAVARRO, E., ALVAREZ JORNA, P. and ARMENTIA, I. (1992) Feeding, particle selection and absorption in cockles Cerastoderma edule (L.) exposed to variable conditions of food concentration and quality. J. Exp. Mar. Biol. Ecol., 162, 177± 198. JOHN, E.H. and FLYNN, K.J. (2002) Modelling changes in paralytic shellfish toxin content of dinoflagellates in response to nitrogen and phosphorus supply. Mar. Ecol. Prog. Ser., 225, 147±160. QUILLIAM, M.A., WALTER, J.A.
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and BEITLER, M.K. (1988) A case for sequestering of paralytic shellfish toxins as a chemical defense. J. Shellfish Res., 7, 629±636. LASSUS, P., FREMY, J.M., LEDOUX, M., BARDOUIL, M. and BOHEC, M. (1989) Patterns of experimental contamination by Protogonyaulax tamarensis in some French commercial shellfish. Toxicon, 27, 1313±1321. LASSUS, P., BARDOUIL, M., LEDOUX, M., MURAIL, I., BOHEC, M., TRUQUET, P., FREMY, J.M. and ROHMER, V. (1992) Paralytic phycotoxin uptake by scallops (Pecten maximus). Aquat. Living Resour., 5, 319±324. LASSUS, P., LEDOUX, M., BARDOUIL, M., BOHEC, M. and ERARD, E. (1994) Kinetics of Alexandrium minutum Halim toxin accumulation in mussels and clams. Natural Toxins, 2, 329±333. LAVRENTYEV, P.J., MCCARTHY, M.J., KLARER, D.M., JOCHEM, F. and GARDNER, W.S. (2004) Estuarine microbial food web patterns in a Lake Erie coastal wetland. Microbial Ecology, 48(4), 567±577. MARGALEF, R. (1973) Some critical remarks on the usual approaches to ecological modelling. Inv. Pesq., 37, 621±640. MEERSCHAERT, M.M. (1999) Mathematical Modelling, 2nd edn. San Diego, CA: Academic Press. Ä O, A. (2000) Ph.D. Thesis. CineÂtica de intoxicacio MORON Ân-desintoxicacioÂn de toxinas PSP en Mytilus galloprovincialis, Santiago de Compostela: Universidad de Santiago de Compostela. KVITEK, R.G.
 NDEZ, M.L., FRANCO, J.M., MARTIÂNEZ, A., REYERO, I., MIÂGUEZ, A., CACHO, E. Ä O, A., FERNA MORON
and BLANCO, J. (1998a) PSP and DSP detoxification kinetics in mussel, Mytilus galloprovincialis: effect of environmental parameters and body weight. In: Reguera, B., Blanco, J., FernaÂndez, M.L. and Wyatt, T. (eds), Harmful Algae, pp. 445±448. Santiago de Compostela: Xunta de Galicia and IOC of Unesco. Ä O, A., MANEIRO, J., PAZOS, Y. and BLANCO, J. (1998b) Modelling the accumulation of MORON PSP toxins in Galician mussels: results and perspectives. In: Reguera, B., Blanco, J., FernaÂndez, M.L. and Wyatt, T. (eds), Harmful Algae, pp. 441±444. Santiago de Compostela: Xunta de Galicia and IOC of Unesco. Ä O, J., FRANCO, J., MIRANDA, M., REYERO, M.I. and BLANCO, J. (2001) The effect of MORON mussel size, temperature, seston volume, food quality and volume-specific toxin concentration on the uptake rate of PSP toxins my mussels (Mytilus galloprovincialis Lmk). J. Exp. Mar. Biol. Ecol., 257, 117±132. Ä O, A., AREÂVALO, F., LION, M., MANEIRO, J., PAZOS, Y., SALGADO, C. and BLANCO, J. MORON (2002) Bloom of Alexandrium minutum Halim in the Galician (NW Spain) RõÂasBaixas and its repercussion on mussels Mytilus galloprovincialis. In: Villalba, A. (ed.), Fourth International Conference on Molluscan Shellfish Safety. Book of Abstracts, Santiago de Compostela: ConsellerõÂa de Pesca. Xunta de Galicia. Â NDEZ, M.L., MANEIRO, J., PAZOS, Y., SALGADO, C. and BLANCO, Ä O, A., AREÂVALO, F., FERNA MORON J. (2003) Accumulation and transformation of DSP toxins in mussels Mytilus galloprovincialis LMK during a toxic episode caused by Dinophysis acuminata. Aquat. Toxicol., 62, 269±280. NAVARRO, E., IGLESIAS, J.I.P. and ORTEGA, M.M. (1992) Natural sediment as a food source for the cockle Cerastoderma edule (L.): effect of variable particle concentration on feeding, digestion and the scope for growth. J. Exp. Mar. Biol. Ecol., 156, 69±87. NAVARRO, J.M. and WIDDOWS, J. (1997) Feeding physiology of Cerastoderma edule in response to a wide range of seston concentrations. Mar. Ecol. Prog. Ser., 152, 175± 186.
Modelling as a mitigation strategy for harmful algal blooms 227 and HOUVENAGHEL, G. (1995) Patterns in long term accumulation of okadaic acid and DTX-1 in blue mussels, Mytilus edulis, experimentally fed with the DSP-containing alga Prorocentrum lima. In: Lassus, P., Arzul, G., Erard, E., Gentien, P. and Marcaillou, C. (eds), Harmful Marine Algal Blooms, pp. 487±492. Paris: Lavoisier, Intercept Ltd. SHUMWAY, S.E. (1989) Toxic algae: a serious threat to shellfish aquaculture. World Aquaculture, 20, 65±74. SHUMWAY, S.E. (1990) A review of the effects of algal blooms on shellfish and aquaculture. J. World Aquacult. Soc., 21, 65±104. SHUMWAY, S.E. and CEMBELLA, A.D. (1993) The impact of toxic algae on scallop culture and fisheries. Rev. Fish. Sci., 1, 121±150. SHUMWAY, S.E. and CUCCI, T.L. (1987) The effects of the toxic dinoflagellate Protogonyaulax tamarensis on the feeding and behaviour of bivalve molluscs. Aquat. Toxicol., 10, 9±27. SHUMWAY, S.E., BARTER, J. and SHERMAN-CASWELL, S. (1990) Auditing the impact of toxic algal blooms on oysters. Environmental Auditor, 2, 41±56. SIERACKI, C.K., SIERACKI, M.E. and YENTSCH, C.S. (1998) An imaging-in-flow system for automated analysis of marine microplankton. Mar. Ecol. Prog. Ser., 168, 285±296. SILVERT, W.L. and CEMBELLA, A.D. (1995) Dynamic modelling of phycotoxin kinetics in the blue mussel, Mytilus edulis, with implications for other marine invertebrates. Can. J. Fish. Aquat. Sci., 52, 521±531. SILVERT, W. and SUBBA RAO, D.V. (1992) Dynamic model of the flux of domoic acid, a neurotoxin, through a Mytilus edulis population. Can. J. Fish. Aquat. Sci., 49, 400± 405. SILVERT, W., BRICELJ, M. and CEMBELLA, A. (1998) Dynamic modelling of PSP toxicity in the surfclam (Spisula solidisssima): multicompartmental kinetics and biotransformation. In: Reguera, B., Blanco, J., FernaÂndez, M.L. and Wyatt, T. (eds), Harmful Algae, pp. 437±440. Santiago de Compostela: Xunta de Galicia and IOC of Unesco. STABELL, O.B., STEFFENAK, I. and AUNE, T. (1992) An evaluation of the mouse bioassay applied to extracts of `diarrhetic' shellfish toxins. Food Chem. Toxicol., 30(2), 139±144. WASTNEY, M.E., PATTERSON, B.H., LINARES, O.A., GREIF, P.C. and BOSTON, R.C. (1999) Investigating Biological Systems Using Modeling: Strategies and Software, edn. San Diego, CA: Academic Press. YAMAMOTO, T., FLYNN, K.J. and TAKAYAMA, H. (2003) Application of a two-compartment one-toxin model to predict the toxin accumulation in Pacific oysters in Hiroshima Bay, Japan. Fish. Sci., 69(5), 944±950. PILLET, S., PEREIRA, A., BRAEKMAN, J.-C.
9 Metals and organic contaminants in bivalve molluscs W.-X. Wang, HKUST, Hong Kong
Abstract: This chapter discusses the accumulation of chemical contaminants in bivalve shellfish and the implications of this accumulation on the safety assessment of shellfish. It focuses primarily on metals because of the extensive available studies on the bioaccumulation of metals in bivalve molluscs, although some representative organic contaminants are also briefly mentioned. Among the metals, a few representative non-essential and toxic metals such as Cd are discussed, as recently there has been an increasing concern about Cd concentrations in bivalves due to new safety standards in shellfish. Key words: metals, organic contaminants, shellfish, bioaccumulation, safety standards.
9.1
Introduction
Over the past several decades there has been considerable concern regarding the contamination of shellfish with metals and organics for two main reasons. The first is the concern about safe consumption of shellfish by humans because of the pollution of coastal waters where shellfish grow and are harvested. This has been especially true with the increasing industrial activity in many countries. Secondly, many shellfish (e.g., mussels and oysters) are employed as biological monitors of coastal contamination in such programs as the early Mussel Watch Program in the 1970s and later the National Status and Trend Program in the US as well as the International Mussel Watch Program (Goldberg et al., 1978; Farrington et al., 1983; O'Connor, 1992). Several sources of metals can contribute to metal accumulation in marine shellfish, including metal in the water
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(or the dissolved phase), metal from prey (such as phytoplankton, small protozoans, bacteria), and metal in sediment. Marine bivalves are known to pump significant amounts of seawater (e.g., tens to hundreds of liters each day by individual filter-feeding bivalves, such as mussels and oysters) and thereby to process particles because of their extremely high filtration activity. Thus, there is a very close interaction between bivalves and the chemicals in the water, many of which are considered to be particle reactive and to remain in the particles. These contaminants are accumulated through both dissolved phase uptake and particulate ingestions by shellfish. Exposure to toxic substances may result from direct consumption by humans. There are numerous chemical contaminants that can cause potential toxicity, including inorganic chemicals (metals or metalloids such as Cd, As, Hg, Pb) and organic compounds (polychlorinated biphenyls ± PCB, dioxins, chlorinated hydrocarbons ± PAH). In addition, new chemicals are also now widespread in many coastal and estuarine waters (such as estrogens and hormones ± many of which are identified as endocrine disruptors, and antibiotics). These chemicals are either produced naturally or anthropogenically and are eventually accumulated in marine food chains as a result of bioaccumulation or biomagnification (e.g., trophic transfer). Various factors, such as the geographic location (the source of input), the type of contaminant, and biological factors such as feeding pattern determine the bioaccumulation and concentrations of these chemicals in shellfish. This review discusses the accumulation of chemical contaminants in bivalve shellfish and the implications of this accumulation on the safety assessment of shellfish, and focuses primarily on metals because of the extensive available studies on the bioaccumulation of metals in bivalve molluscs, although some representative organic contaminants are also briefly mentioned. Among the metals, a few representative non-essential and toxic metals such as Cd are discussed as recently there has been an increasing concern about Cd concentrations in bivalves due to new safety standards in shellfish.
9.2
Metal concentrations in bivalve molluscs
There are remarkable differences in the concentrations of metals found in different species of bivalve molluscs. The best-known examples are the differences in metal concentrations in mussels and oysters collected simultaneously at the same locations by the National Status and Trends Programs (O'Connor, 1992). The concentrations of Ag, Cd, Cu, Ni, Zn, which tend to bind with the Scontaining ligands (e.g., proteins), are typically much higher in American oysters (Crassostrea virginica) than in mussels (Mytilus sp.), whereas the concentrations of Se, Pb, Hg, Cr, and As in mussels are higher than those found in oysters collected from the same sites. The concentrations of organic compounds such as total PCB and PAH are also higher in mussels than they are in oysters. There have been numerous reports on the concentrations of metals and organics in different
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species of marine bivalves. Wang and Wong (2006) recently quantified the seasonal variations of concentrations of five trace metals/metalloids (Ag, Cd, Cu, Se, and Zn) for 1 year in the black mussels Septifer virgatus and rocky oysters (Saccostrea cucullata) collected from a rocky shore in Clear Water Bay in the eastern part of Hong Kong, which is subjected to significant influence from ocean currents. The bay is considered relatively pristine without significant impacts from anthropogenic activity. There were striking differences in the metal concentrations in the black mussels and the oysters. The concentrations of Class B type or borderline metals (Ag, Cu, Cd, Zn) in the oysters were higher than those in the mussels, whereas the concentrations of Se were comparable in both species. Even among the taxonomically close species, metal concentrations can also vary considerably. For example, the Ag concentrations in the black mussels were generally very low (0.01±0.06 g g±1), in strong contrast to the Ag concentrations measured in the common mussel Mytilus edulis collected from different regions of US coastal waters (geometric mean of 0.17 g g±1, O'Connor, 1992). Ng and Wang (2005a) measured the Ag concentrations in the green mussel Perna viridis collected from coastal (eastern) and estuarine (western) sites in Hong Kong. The Ag concentrations in the estuarine population were 0.04± 0.17 g g±1, and much higher concentrations were found in the coastal population (0.21±0.28 g g±1). The lower Ag concentrations in the estuarine population were attributed to higher Ag efflux from the mussels collected from the estuarine site (see below). Based on such differences among closely related bivalve species, future studies should examine the mechanisms underlying the accumulation and natural levels of Ag in bivalves. Cadmium (Cd) concentrations also vary considerably between bivalve species. Pigeot et al. (2006) surveyed Cd concentrations in different species of bivalves, crustaceans, and gastropods from a bay along the Atlantic coast of France. Among the many different species of bivalves, the Cd concentrations differed by a factor of 10. Scallops (Chlamys varia) contained the highest Cd concentrations (11.8 g g±1) whereas the clams Macoma balthica contained the lowest Cd concentrations. Even among the different clam species, Cd concentrations ranged from 0.3 g g±1 in Tapes philippinarum to 8.3 g g±1 in Modiolus barbatus. In a recent study, Belcheva et al. (2006) surveyed the Cd concentrations in the digestive glands of the Japanese scallop Patinopecten yessoensis from the Sea of Japan. They found that Cd concentrations in the digestive gland increased significantly with age and as high as 150 g g±1 dry weight of Cd was found in the digestive glands of scallops collected from unpolluted areas. Clearly, much research is needed to understand the mechanisms underlying these high Cd concentrations and their ecotoxicological significance. It is also very important to recognize that the high concentrations of metals in seafood are not necessary related to pollution ± a fact that has often been overlooked by shellfish buyers, sellers and regulators. For instance, it has been known for decades that oysters are able to accumulate extremely high levels of Zn and Cu in their tissues (Rainbow, 1990; Rainbow and Philips, 1993). Thus, even in pristine areas, the accumulation of metals by molluscs may be
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problematic and is likely affected by environmental, seasonal, and/or chemical factors. Recently, increasing numbers of studies have attempted to understand the body concentrations of metals in bivalves from contrasting marine environments or populations. Such studies have relied mainly on the biokinetic model and the measurements of kinetic parameters described in the kinetic model, in which the concentration of a metal in the bivalves is controlled by the balance between uptake, elimination and growth, as described by the following equation (Thomann, 1981; Wang et al., 1996; Luoma and Rainbow, 2005): dCt ku Cw AE IR Cf ÿ
ke g Ct 9:1 dt where Ct is the metal concentration in an animal at time t, ku is the uptake rate constant from the dissolved phase (L g±1 day±1), Cw is the metal concentration in the dissolved phase (g L±1), AE is the metal assimilation efficiency from the dietary phase, IR is the weight-specific ingestion rate of the animal (g g±1 day±1), Cf is the metal concentration in the dietary phase (g g±1), ke is the efflux rate constant (day±1), and g is the growth rate constant of the animal (day±1). The kinetic model can be used to describe the time-course of the bioaccumulation dynamics: Ct Css 1 ÿ eÿ
ke gt
9:2
where Css is the metal concentration in bivalves achieved under steady-state conditions: ku Cw AE IR Cf 9:3 Css ke g A typical application is to compare field-measured metal concentrations in bivalves and model-predicted values after kinetic quantifications in the laboratory coupled with geochemical measurements in the field. Ke and Wang (2001) compared the bioaccumulation of Cd and Zn in an estuarine oyster (Crassostrea rivularis) and a coastal oyster (Saccostrea glomerata). Parameters quantified in this study included the dietary assimilation efficiency, dissolved uptake rate, and efflux rate. The uptake rate constants quantified for Cd and Zn were the highest among different bivalve species, likely because of their very high filtration rates (Wang, 2001). A major difference was observed in the efflux rate constant, which was higher in the estuarine oysters than in the coastal oysters. The bioenergetic-based kinetic model demonstrated that estuarine and coastal oysters differ in their strategies in accumulating high metal concentrations in their tissues. The estuarine oyster C. rivularis achieves a high concentration of Cd and Zn presumably by the high metal influx rate from both the aqueous and particulate phases. In contrast, the extremely low metal efflux rate may have caused the accumulation of metals in the coastal oysters to a high level, despite their influx rates from both the aqueous and particulate sources being lower than those in estuarine oysters. This study also predicted a likely Cd and Zn concentration in S. glomerata of 3.3±10.0 g g±1 and 1130±7500 g g±1, which is
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similar to the actual measured metal concentrations (2±17 g g±1 for Cd, and 1600±7700 g g±1 for Zn). Ng and Wang (2005b) took a similar modeling approach but quantified the differences in Cd biokinetics from two populations of green mussels, Perna viridis, from two sites (eastern and western) in Hong Kong with contrasting hydrological conditions. Mussels collected from the western site had three to six times higher Cd tissue concentration than did the eastern population collected during two seasons (wet summer and dry winter), but the salinity was lower in the western site only during the summer. The Cd uptake rate constant from the dissolved phase was higher in the western population during the summer owing to a much lower salinity, but it was comparable during the winter. Dietary uptake of Cd was similar in both populations, while assimilation of Cd was lower from the ingested radiolabeled seston than from diatoms. The efflux of Cd remained comparable between the two populations during the two seasons (0.02±0.03 day±1). The model implies that the faster influx of Cd from the aqueous phase caused the higher body Cd concentrations in the western population. The kinetic model was able to predict four to five times difference in Cd concentrations in the two populations of mussels. This study highlighted the differences in Cd accumulation kinetics in different populations of mussels, likely caused by the different physical environments. Thus, both Cd geochemistry and Cd uptake kinetics play a role in the bioaccumulation differences between the populations. In contrast to Cd, the Ag concentration in the green mussels collected from the eastern population of Hong Kong were two to six times higher than that from the western population collected in different seasons (Ng and Wang, 2005a). Such differences in the body burdens between the two populations may similarly be explained by both the Ag geochemistry and metal physiology of Ag. Although the Ag uptake from the aqueous phase was faster in the western population during all seasons, no significant relationship between Ag uptake and salinity was found. Dietary uptake of Ag was similar in both populations. In contrast, the efflux of Ag was twice as fast in the western population during the wet and dry seasons, which may partially explain the lower Ag tissue burden in the mussels. Consequently, it is clear that different biokinetics (dissolved uptake, efflux, assimilation) contribute to different metal concentrations in different species of bivalves or within a single species but from different populations. Kinetic modeling approach can identify the key processes leading to metal accumulation in marine shellfish. Linkage of such modeling to the management of shellfish safety has not yet been explored, but this certainly remains a possibility. For example, in order to reduce the levels of toxic metals in shellfish (a management issue), some of the key processes can be scrutinized after the modeling study. Producers of shellfish may then be able to modify these key processes to ensure that the metal levels are below the safety limits. Clearly, modeling will remain a promising approach for the shellfish industries and regulators.
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9.3
233
Internal speciation of metals in bivalve molluscs
It is important to recognize that the total metal concentrations in bivalve molluscs may not necessarily provide useful information for risk assessments for human consumption. The internal distribution of metals in the tissues (e.g., subcellular pools) is equally important in risk assessments, but this distinction has just started to receive attention. The subcellular distribution of metals in bivalves has been examined using biochemical fractionation techniques (Wallace and Lopez, 1996; Wang et al., 1999; Ettajani et al., 2001; Blackmore and Wang, 2002; Wallace et al., 2003). These studies generally separated the metals either into five subcellular pools (cellular debris, metallothionein-like proteins (MTLP), heat sensitive proteins (HSP), metal-rich granules (MRG), and organelle fractions; Wallace and Lopez, 1996). Other studies also separated metals into three subcellular pools (insoluble, soluble, and weakly bound metals; Ettajani et al., 2001; Miao and Wang, 2006). Several factions can be combined (e.g., the metal-sensitive fraction from cellular debris, organelles and HSP, the biologically detoxified metal fraction from MT and MRG, and the trophically available metal fraction from organelles, HSP and MT) (Wallace et al., 2003). The subcellular distribution of metals also critically affects their trophic transfer to predators such as marine gastropods and fish (Cheung and Wang, 2005; Zhang and Wang, 2006). For example, the dietary AEs were higher in animals fed the heat-stable protein fraction or the heat-sensitive protein fraction than in those fed the insoluble fractions. These recent studies demonstrated that metals associated with the insoluble fraction or MRG had a lower bioavailability than metals associated with the soluble fraction. Different metals generally have contrasting associations with different subcellular pools, depending on the bivalve species, exposure history, and other conditions. It appears that one major fate of metals in bivalves is to move into the insoluble fraction (or MRG, cellular debris), although some metals are able to induce MT as their storage sites. Cd is mainly associated with the insoluble fraction in marine mussels, but MT is also an important ligand binding to Cd. In the green mussel P. viridis collected from Hong Kong coastal waters, for example, 60±80% of Cd was distributed in the insoluble fraction, and 5±25% was associated with the MTLP fraction, but the percentages varied with the collection site and season (Ng and Wang, 2005b). In the clam Ruditapes philippinarum, about 50±60% of Cd was found in the insoluble fraction, as compared to 20±30% in the MTLP fraction (Ng and Wang, 2004). In the clam Mactra veneriformis collected from a contaminated bay in Northern China, the distribution varied with the degree of contamination, with the more contaminated clams accumulating significantly higher levels of Cd (21%) in MTLP fraction than did the individuals from the less-contaminated site (Shi and Wang, 2004a). Cadmium distribution in oysters (Saccostrea cucullata) is also comparable to that observed in mussels and clams, with most Cd in the insoluble fraction (Cheung and Wang, 2005). Similar to Cd, the majority of Zn is found in MRG or the insoluble fraction in mussels, clams, oysters (Blackmore and Wang, 2002; Baudrimont et al., 2003; Shi and Wang, 2004a; Cheung and Wang, 2005).
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In green mussels, much less Zn is found in the soluble fraction, and 28 ëC
Maximum hours from harvest to temperature control 36 14 12 10
hours hours hours hours
therefore might be diminished by summer closure of harvest waters as has been proposed for the US Gulf states. As an alternative to outright summer closure and its ancillary economic hardship, a time±temperature matrix has been developed. Time±temperature matrices are management devices that specify the time required from harvest to temperature control (e.g., ice, refrigeration) at various water temperatures. A time±temperature matrix approach is in effect for the management of V. vulnificus (Table 10.2) for the US Gulf Coast (NSSP, 2003). During the summer, for example, oysters must be cooled within 10 h of initial harvest. The application of post-harvest processing may help diminish risk of infection by pathogens. Post-harvest processing includes various techniques such as relaying, depuration, irradiation, pasteurization, high hydrostatic pressure (HHP), and freezing. Relayed shellfish are initially harvested in a contaminated area and transferred to an unpolluted environment where they potentially eliminate their pathogens. Relaying typically occurs as a transfer of shellfish from a lower-salinity Conditionally Approved area to a higher-salinity Approved area. Improved water quality or higher salinity can be used to improve shellfish quality. Microbial indicators and pollution-associated bacteria are obvious candidates for diminution by relaying; however, the higher-salinities of Approved waters can diminish the numbers of some naturally occurring pathogens. For example, Cook and Ellender (1986) exposed eastern oysters to bacteria and viruses and relayed them to relatively unpolluted waters. Fecal coliforms were generally reduced >99.9% within 5 days, and S. typhimurium was eliminated in 25ù) of seaside bays of Maryland and Virginia (Andrews et al., 1962; Haskin et al., 1966; Sindermann, 1990) and in Long Island Sound, New York (Sunila et al., 2002). Recently, plasmodia, but not spores, of Haplosporidium costale were detected in low prevalence and intensity of infection in Crassostrea virginica from several locations in the Southern Gulf of St. Lawrence, Atlantic coast of Nova Scotia and Bras d'Or Lakes in Cape Breton, Nova Scotia, Canada (Stephenson et al., 2003). Impact on the host Because of the difficulty in distinguishing SSO from MSX, epidemiological data may be subjected to certain inaccuracy. Mortalities typically occur 10 months after exposure to the pathogen ± a striking characteristic of the disease is the sudden onset of mortalities in May and the abrupt termination of mortalities about mid-June (Andrews and Castagna, 1978; Andrews, 1984). Mortalities of native stocks in the restricted enzootic area average 20±50%. The route of infection and life cycle remain unknown (Sindermann, 1990).
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11.2.4 Paramyxeans Marteilia refringens (Aber disease of the European flat oyster) and Marteilia sydneyi (QX disease of the Sydney rock oyster) These paramyxean parasites have caused recurring seasonal (summer) mass mortalities of the edible flat oyster, Ostrea edulis, in Europe (Grizel et al., 1974; Marteil, 1976; Alderman, 1979) and of the Sydney rock oyster, Saccostrea glomerata, in Australia (Perkins and Wolf, 1976). Infection with Marteilia refringens and M. sydneyi are regarded as major concerns for molluscan aquaculture and are listed as notifiable to the OIE (OIE, 2006). Marteilia refringens occurs in the Southern Atlantic Europe from Brittany to Portugal, and in the Mediterranean. Marteilia sydneyi is found in coastal estuaries of southern Queensland and northern New South Wales, Australia (Adlard and Ernst, 1995). Several species, Ostrea edulis, Mytilus edulis, M. galloprovincialis are susceptible to infection with Marteilia refringens (Berthe et al., 2000) although mussels appear more predominantly infected by M. maurini, a sister species of M. refringens (Le Roux et al., 2001). In addition, a copepod species, Paracartia grani, has been identified with potential role in the life cycle of Marteilia refringens (Berthe et al., 1998; Audemard et al., 2001, 2002). Besides Saccostrea glomerata, M. sydneyi possibly also infects Striostrea mytiloides and Saccostrea forskali (Berthe et al., 2004a). Impact on the host In oysters, infection with Marteilia refringens results in a poor condition index with glycogen loss (emaciation), discoloration of the digestive gland, cessation of growth, tissue necrosis, and increased mortalities (Grizel et al., 1974). In mussels, Marteilia most often occurs without causing any clinical disease. Factors triggering onset of disease are not clearly established, but may be related to environmental stress such as temperature and/or stock differences in disease resistance (Berthe et al., 2004a). Digestive tropism is also a feature of infection with Marteilia sydneyi, with the host oyster showing signs that include poor condition factor and shrunken body. Death is attributed to direct blockage of the digestive gland by the parasite and consequent host starvation (Wolf, 1979). While M. refringens is completing its development in the digestive tract with luminal tropism, M. sydneyi development through to sporogenesis involves connective tissue location (Kleeman et al., 2002a). In enzootic areas, control of marteiliosis is attempted by curtailing the planting of oysters during the period of transmission and by growing oysters in high-salinity waters (35±37 ppt) to limit the development of Marteilia refringens (Berthe et al., 2004a). 11.2.5 Xenohaliotis californiensis (withering syndrome of abalone) Withering syndrome is a chronic wasting disease of abalone (Haliotis spp.) which was first observed in the Channel Islands in the mid-1980s (Haaker et al., 1992). This fatal disease played a significant role in the demise of southern Californian black abalone and may also be contributing to the lack of recovery
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of other species following their severe depletion by over-fishing. The causative agent of withering syndrome is an intracellular Rickettsiales-like prokaryote that infects the gastrointestinal epithelium of abalone (Friedman et al., 2000; Moore et al., 2002). Although final taxonomic position has not been established for this new genus and new species, denomination includes `candidatus'; full approval of Xenohaliotis californiensis is however expected. The disease is reported from California, USA, Baja California, Mexico, and Chile (Altstatt et al., 1996; Finley and Friedman, 2000). The pathogen has also been recorded more recently from Iceland, Ireland and Spain with no reports of significant mortality rates. It affects several species of abalone, including Haliotis cracherodii, H. rufescens, H. corrugata, H. fulgens, H. sorenseni and possibly H. discus hannai which are susceptible to infection with X. californiensis. Impact on the host Xenohaliotis californiensis affects all sizes of abalone and causes lethargy, retracted visceral tissues, atrophy of the foot muscle and eventual death. The disease has progressively spread throughout the Californian Channel Islands, causing severe reduction of abalone populations on six of the eight Channel Islands by 1992 and closure of the Californian black abalone fisheries in 1993. Survivors of disease outbreaks appear to gain some resistance to the disease (Friedman et al., 2003b). In affected culture facilities, the severity of the disease may be curtailed if water temperature is reduced to about 15 ëC or less (Moore et al., 1999). In addition, it has been shown that intramuscular injection and oral administration of oxytetracycline were effective in reducing the losses of infected abalone (Friedman et al., 2003a). 11.2.6 Perkinsus marinus and P. olseni The genus Perkinsus includes parasites that infect molluscs worldwide, some of which are associated with serious diseases and mass mortalities (Villalba et al., 2004). The first Perkinsus species was described in Crassostrea virginica (Mackin et al., 1950; Sparks, 2005). Since then, several new species have been added to this genus (Berthe et al., 2004b; Villalba et al., 2004). Perkinsus marinus is found on the East coast of the USA (Andrews, 1996) from Maine to Florida, and along the Gulf of Mexico coast to the Yucatan Peninsula (Burreson et al., 1994). The disease was also introduced to Hawaii (Kern et al., 1973). Perkinsus olseni occurs on the coasts of South Australia and New South Wales in Australia (Lester and Davis, 1981), and several Perkinsus spp. were found in various molluscan species from the Great Barrier Reef and Western Australia (Goggin and Lester, 1995; Hine and Thorne, 2000). Over 67 molluscan species are known worldwide as species susceptible to infection with parasites of the genus Perkinsus. The species described in Europe and Asia, Perkinsus atlanticus (Azevedo, 1989), has close taxonomic relationships with P. olseni and synonymy was proposed, with the name P. olseni having priority (Goggin, 1994; Murrell et al., 2002).
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Perkinsus qugwadi is found in the scallop Patinopecten yessoensis on the west coast of Canada (Blackbourn et al., 1998) while P. chesapeaki occurs in the soft-shell and arctic clams, Mya arenaria and Macoma balthica, on the east coast of the USA (McLaughlin et al., 2000). Perkinsus mediterraneus was recently described in Ostrea edulis in Spain (Casas et al., 2004). A more complete list of host species and valid Perkinsus species can be found in Villalba et al. (2004). Two species among the seven which have been described, Perkinsus marinus and P. olseni, are given particular attention because of their impact on susceptible species. Impact on the host Progression of Perkinsus spp. infections through the host tissues causes lesions that may lead to host death. Sublethal effects may cause interference with host energy fluxes, which may, in turn, lead to slower growth rate and poor condition of market size individuals. Infection causes host weakening so that it becomes increasingly difficult for affected individuals to overcome any other adverse conditions (Villalba et al., 2004). Pacific oyster, Crassostrea gigas, and triploid Suminoe oysters, C. ariakensis, were found to be less susceptible (although not completely resistant) to infection with Perkinsus marinus (Meyers et al., 1991; Calvo et al., 2000). Greater resistance to infection and disease in C. gigas may be attributed to elevated cellular and humoral activities that may degrade the parasite more effectively, and/or lower plasma protein levels that may limit parasite growth (La Peyre et al., 1995). Perkinsus olseni proliferates in tissues and may produce pustules in the foot and mantle of Haliotis rubra and H. laevigata (Goggin and Lester, 1995). Transmission of the parasite occurs directly between individual molluscs. Infection by Perkinsus olseni in wild Haliotis rubra is positively correlated with both water temperature and size of abalone (Lester et al., 2001). With respect to P. marinus, management strategy consist in reducing the density of oysters, and harvesting or transferring oysters to low salinity areas (lower than 9 ppt) before water temperatures increase to 15±20 ëC; however, infections may persist for years in low salinity areas (Burreson and Ragone Calvo, 1996). Fallowing of infected oyster beds has not proven effective but early harvest is considered to avoid mass mortalities caused by P. marinus (Butsic et al., 2000). Cold winter temperatures may limit the natural spread of this pathogen to northern areas. For Haliotis rubra and H. laevigata, stress such as high temperature (e.g. 20 ëC) may predispose abalone to disease. Infected Haliotis rubra but not H. laevigata appear to contain, and possibly eliminate, the infection at 15 ëC and during the winter (Lester and Davis, 1981; Goggin and Lester, 1995). Disease outbreaks in abalone culture facilities were controlled by isolating infected tanks, removing off the infected abalone, followed by washing equipment in fresh water (Goggin and Lester, 1995). Cyclohexamide was found to be effective in reducing Perkinsus marinus infections in live oysters without killing the host, although infections were not completely eliminated (Calvo and Burreson, 1994; Villalba et al., 2004). There is no known therapeutics to control Perkinsus olseni
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infection in abalone. Efforts to develop C. virginica stocks resistant to Perkinsus marinus are still in progress (Stickler et al. 2001).
11.3
Diagnostic methods
Until the mid-1980s, the standard diagnostic methods used for the detection of molluscan pathogens were cytology, histology and electron microscopic techniques (Miahle et al., 1995; OIE, 2005b). Polyclonal or monoclonal antibodies raised against specific pathogens have had a limited contribution in the biotechnological advancement of diagnostics for molluscan diseases despite several attempts in opening that avenue. This may have several causes among which we would here emphasise two. Most of the pathogens of significance, with exception of Perkinsus species, are non-culturable organisms. For a long time, inclusion of positive controls in diagnostic assays has been extremely difficult, and impossible in commercially available kits such as enzyme-linked immunosorbent assay (ELISA), for example. Another impediment to popular use of antibodies as diagnostic reagents may certainly lie in the fact that taxonomy of the significant pathogens has long been uncertain, hence rendering specificity of antibodies difficult to assess. However, with the increasing need to improve levels of sensitivity and specificity, as well as to gain more powerful tools for investigating the biology of these pathogens, DNA-based molecular assays were developed such as polymerase chain reaction (PCR), in situ hybridization, and real-time PCR for several prominent disease agents (Berthe et al., 1999). Probes were published for pathogens belonging to Mikrocytos (Adlard and Lester, 1995; Carnegie et al., 2003a), Bonamia (Carnegie et al., 2000, 2003b; Cochennec et al., 2000; Corbeil et al., 2006b; Marty et al., 2006), Perkinsus (Robledo et al., 2000; Yarnall et al., 2000; Audemard et al., 2004;), and Marteilia (Le Roux et al., 1999, 2001; Kleeman and Adlard, 2000). Although some comparative studies have been conducted (Pernas et al., 2001, Kleeman et al., 2002b; Diggles et al., 2003), true validation of these reagents and procedures is still limited (Berthe, 2000; TheÂbault et al., 2005). Positive features of these molecular assays include low cost, speed, sensitivity, reliability and specificity. Although these assays have limitations of their own, they have shown to be invaluable as complementary tools for the diagnosis and study of molluscan pathogens at the molecular level.
11.4 Effects of shellfish disease on the international shellfish industry 11.4.1 Economic losses due to molluscan diseases Impacts of diseases on the shellfish industry include direct loss because of high mortality rates; reduced yield per individual or stock, early harvest before animals have reached market sizes, culling of abnormal animals; reduced
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demand for the product because of lower quality and aesthetic values, and possible regulatory response of closing production areas, or withholding products from markets owing to unknown effects of a disease on humans (Sindermann, 1990). The industry may also suffer from indirect effects such as loss of market position for the product, negative effects on consumers due to association with `disease' image, especially when used in news media; lost or reduced crops affecting the local economy; cost of research and management activities aimed at mitigating the effects of disease; and reduced attractiveness of the affected shellfish species to recreational fishermen, with consequent losses to coastal service industries (Sindermann, 1990). When facing health issues with actual or potential economic losses, producers must implement new strategies for the industry to remain viable and sustainable. Some of these strategies include control of movement of animals on to the farms or farming sites; destruction of clinically sick animals; the reduction in animal size for marketing, disinfection of rearing facilities; planting in areas less subject to disease; planting of seed stock at a lower density to reduce likelihood of disease transmission; fallowing prior to re-stocking; removing animals from a diseased area for several years, thereby eliminating the pathogen population; replacing susceptible stocks or species with more resistant animals (Sindermann, 1990; Subasinghe et al., 2001). Regarding the consumption side of the market, diseases in shellfish have significant negative impacts on marketability and consumer confidence. For example, a disease outbreak can result in a reduction in demand for products that look, taste or smell abnormal ± with subsequent disinterest in related products. The reduced availability of particular products often results in higher prices, which often motivate the search for less expensive substitute foods (Sindermann, 1990). 11.4.2 International trade and health issues Over the past several decades aquaculture has expanded, intensified and diversified, based heavily on movements of live aquatic animals and animal products (broodstock and seed). This growth has been prompted, in particular, by depletion of fisheries resources, development of niche markets, and certainly world trade liberalisation. This trade, facilitated by improved transportation efficiency, is now recognised as having played a pivotal role in the introduction and spread of pathogens and diseases into new populations and geographic regions (Subasinghe et al., 2001). There is now convincing evidence of the serious socioeconomic, environmental and international trade consequences arising from global trans-boundary movements of aquatic animal pathogens. Per annum figures of the financial consequences of disease losses in aquaculture species reach, for some countries, hundreds of millions of US dollars (Chinabut, 1994; Alday de Graindorge and Griffith, 2001; Wei, 2000). A global estimate by the World Bank in 1997 of financial losses to aquaculture, due to disease, was in the range of US$3 billion
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per annum (Subasinghe et al., 2001). In addition to the obvious effects of largescale aquaculture losses on rural communities, diseases also cause considerable financial impact on investor confidence (Subasinghe et al., 2001). Interventions to prevent or mitigate impacts of disease include early diagnosis, treatment, control or eradication, ecosystem management, advanced breeding programs, restocking programmes and quarantine measures such as health certification and movement restrictions. Other possible actions include risk assessments, regional agreements, collaboration and complementary legislation, closed cycle production, specific pathogen-free broodstock and hatcheries, specialized diagnostic capabilities, research, development and training in improved diagnostic capabilities, development of disease information systems, a `zoogeographic' approach to quarantine and health certification and development of resources and infrastructure (Humphrey, 2001).
11.5
Reducing disease in molluscan aquaculture
11.5.1 Breeding programmes Because of their higher market value, structure of the private sector and farming operations, genetic advances in aquaculture species are further advanced for finfish than for bivalve molluscs (Gosling, 2003). However, bivalve populations exhibit an enormous degree of genetic variation making them good candidates for enhancement of desirable traits, such as disease resistance, through a selective breeding programme. This has even been seen by some as the way forward in reducing the impact of diseases. In order to obtain individuals possessing suitable disease-resistance traits that can be maintained through generations, breeding schemes such as individual selection or family selection are established (Gosling, 2003). These schemes aim at selecting desired traits while minimising the loss of other beneficial traits such as growth and high survival rates. Some recent examples of successful disease resistance programme in oysters include the following projects. Decline in American oyster, Crassostrea virginica, populations caused by the parasite Haplosporidium nelsoni (MSX) (95% mortality of native oysters in Delaware Bay between 1957 and 1959) prompted the US government to develop a MSX-resistant oyster breeding programme. Since 1960 strains of oysters have been produced that are resistant to this disease. This has been achieved by exposing lines (F1 generation), developed from survivors of the 1957±59 epizootics, to MSX-infected waters for a period of 33 months. Survivors of the F1 line were then used to produce the F2 line and these were again challenged as before. This procedure was followed for several generations (Ford and Haskin, 1987). IFREMER, the French Research Institute for Exploitation of the Sea, has also undertaken a breeding programme that aims at selecting a flat oyster, Ostrea edulis, resistant to the protozoan parasite Bonamia ostreae which caused a serious decline in the flat oyster industry in the early 1980s and onward. The
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selected strains have shown better survival in Bonamia-affected areas compared to both natural spat and hatchery-reared wild spat. In addition, some cross-bred lines have shown superior growth rates (Baud et al., 1997; Naciri-Graven et al., 1998; 2000). A selection for bonamiasis-resistant flat oyster has also been successful in Ireland (Hugh-Jones, 1999; Culloty et al., 2001). This study, however, did not find any correlation between resistance and growth rate. In Australia, the QX disease in Sydney rock oysters (Saccostrea glomerata), caused by the paramyxean protozoan Marteilia sydneyi, can cause up to 95% mortality in the summer months. Since 1996 the New South Wales State Government has been selectively breeding Saccostrea glomerata for resistance to QX disease. The breeding programme has had a specific affect on the defensive phenoloxidase enzyme system of oysters. The study has shown that the breeding for resistance has resulted in negative selection against one form of phenoloxidase (Bezemer et al, 2006). 11.5.2 Transgenic animals In recent years, more research has been directed towards the introduction of foreign DNA into the germ line of aquatic animals with the aim of improving traits favourable to commercialisation. Significant progress has been made with transgenic fish species which showed improved growth performance by overexpressing growth hormones (Devlin et al., 1995; Dunham et al., 1992; Chen et al., 1993; Rahman et al., 2001). In addition, transgenic zebra fish have been developed as models for the study of vertebrate development, diseases and genetics (Ju et al., 1999; Kobolak and Muller, 2003) as well as for environmental studies (Mattingly et al., 2001). More recently, transgenic fishes expressing antimicrobial and immuno-modulatory molecules have shown increased resistance to some pathogens (Dunham et al., 2002; Weifeng et al., 2004). Currently, transgenic work with molluscs is not as advanced as in fish, partly due to the non-existence of continuous molluscan cell lines. However, some progress has been achieved by identifying homologous and heterologous promoters that transiently express reporter molecules in primary cultures of molluscan cells (Cadoret et al., 2000; Boulo et al., 2000). Technologies using gene transfer to embryos and eggs have also shown promise (Buchanan et al., 2001; Cadoret, 1992; Cadoret et al., 1997a,b; Mialhe et al., 1997; Powers et al., 1995) but need to be improved to become cost effective and produce higher survival rates in recipient animals before commercialisation can be considered feasible. Disease-resistant molluscs obtained through transgenesis of genes encoding non-specific immune molecules such cytotoxic proteins and antimicrobial peptides (Cadoret et al., 2000; Roch, 1999; Robledo and Vasta, 2003) or pathogen-specific molecules such as RNA interference, would be of tremendous value for the aquaculture industry as losses due to diseases would be greatly reduced.
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11.5.3 Management strategies An effective health management programme must cover all aspects of aquaculture activity, from the production stock at the farm level to issues at the international level and must be based on open communication and multidirectional information exchange and feedback. At the local level, farmers must have access to fisheries officers trained in health management and field-level disease surveillance and reporting. Farms or growing areas must have suitable environmental conditions (high water quality, manageable predation/fouling, etc.) and there must be consistent and accurate record keeping (source of stock, feed, growth records, weather conditions, etc.). Such records are invaluable for determining the source or nature of a disease outbreak, assisting in accurate and rapid diagnosis, decision making and providing appropriate intervention and control measures (Subasinghe et al., 2001). National aquatic animal health programmes, biosecurity and effective implementation of codes of practice must also be established to ensure nationwide health guidelines and policies. National strategies on aquatic animal health are, thus, becoming essential programmes which contain important elements such as policy, legislation and enforcement, risk analysis, pathogen lists, information systems, health certification, quarantine, disease surveillance, monitoring and reporting, zoning, emergency preparedness, research, institutional structure, human resources development, and regional and international cooperation. These components are selected by countries on the basis of the objectives of their national plan. Clearly, dealing with day-to-day situations in farms, pond/farm health management is of prime importance in preventing, controlling and even eradicating serious diseases. Therefore, activities such as health monitoring, diagnosis, quarantine, reporting, communication, emergency planning and response must be undertaken not only at the country/state level but also, and before anything, at the locality or farm level. While ongoing efforts at certification of service providers (both human and facilities) and strengthening cooperation between fisheries and veterinary authorities are expected to further reduce the impact of aquatic animal diseases, empowering farmers to manage disease and other risks will be the key to success under conditions of everdecreasing public funds.
11.6
Future trends
In many regions of the world, increasing demand for aquaculture products in conjunction with diminishing seafood supplies from capture fisheries will continue to stimulate the further development and growth of the aquaculture sector. During the past few decades, aquaculture has not only expanded, diversified, intensified but also made significant technological advances. The potential of this development to enhance local food security, alleviate poverty and improve rural livelihoods has been well recognised (FAO, 2002).
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There is a strong need for further education to increase farmers' awareness of disease and its prevention. Routine surveillance for current and emerging diseases needs to be maintained or implemented. Regional disease surveillance programmes, cooperation between industries and the regulatory authorities and contingency planning should be established, incorporating a network of reference diagnostic laboratories, which in turn need to be capable of supporting health certification efforts, using standardised protocols that are adopted throughout the region. Finally, within farming communities, communication needs to be more effective, so that vital management information can be rapidly disseminated. Again, this needs to be addressed at both the national and regional levels (Alday de Graindorge and Griffith, 2001).
11.7
Sources of further information and advice
11.7.1 OIE reference laboratories · Virginia Institute of Marine Science (Dr Eugene Burreson), Gloucester Point, Virginia, 23062, USA: http://www.vims.edu/abc/ · Fisheries and Oceans Canada (Dr Susan Bower), Pacific Biological Station: http://www.pac.dfo-mpo.gc.ca/sci/aqua/default_e.htm · IFREMER, Laboratoire de Genetique et Pathologie (Dr Isabelle Arzul): http://www.ifremer.fr/anglais/program/proge.htm 11.7.2 Disease information · Bonamia ostreae: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ bonostoy_e.htm · Bonamia exitiosa: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ bonamoy_e.htm · Bonamia roughleyi: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ mikrouoy_e.htm · Marteilia refringens: http://www.ifremer.fr/crlmollusc/page_labo/leaflets/ marteilia_refringens1.htm · Marteilia sydneyi: http://www.dpi.nsw.gov.au/research/areas/productionresearch/aquaculture/outputs/2004/output-225 · Mikrocytos mackini: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ mikmacoy_e.htm · Perkinsus marinus: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ pmdoy_e.htm · Perkinsus olseni: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/pages/ perkolab_e.htm · Withering syndrome of abalone: http://www.pac.dfo-mpo.gc.ca/sci/shelldis/ pages/fwsab_e.htm
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11.7.3 Other web-based resources · AquaNet: https://www.mun.ca/virt/aquanet/English/index.php · Australian Government Department of Agriculture, Fisheries and Forestry: http://www.affa.gov.au/ · CEFAS: OIE Collaborating Centre for Information on Aquatic Animal Diseases. http://www.cefas.co.uk/oie.html · CSIRO Livestock Industries, Australian Animal Health Laboratory: http:// www.csiro.au · Network of Aquatic Centers in Asia-Pacific (NACA): http://www.enaca.org/ · Office International des Epizooties Aquatic Animal Health Code ± 2005: http://www.oie.int/eng/normes/fcode/en_sommaire.htm · Office International des Epizooties Manual of Diagnostic Tests for Aquatic Animals ±2006: http://www.oie.int/eng/normes/fmanual/A_summry.htm
11.8
References
and ERNST I (1995), `Extended range of the oyster pathogen Marteilia sydneyi', Bull Eur Ass Fish Path, 15, 119±121. ADLARD R D and LESTER R J G (1995), `Development of a diagnostic test for Mikrocytos roughleyi, the aetiological agent of Australian winter mortality of the commercial rock oyster, Saccostrea commercialis (Iredale & Roughley)', J Fish Dis, 18, 609± 614. ALDAY DE GRAINDORGE V and GRIFFITH D (2001), `Ecuador', in: Shrimp Farming and the Environment. Thematic Review on Management Strategies for Major Diseases in Shrimp Aquaculture. A Component of the WB/NACA/WWF/FAO Programme on Shrimp Farming and the Environment. Report of the Workshop held in Cebu, Philippines, 28±30 November 1999 (http://www.enaca.org/shrimp). ALDERMAN D J (1979), `Epizootiology of Marteilia refringens in Europe', Mar Fish Rev, 41, 67±69. ALTSTATT J M, AMBROSE R F, ENGLE J M, HAAKER P L, LAFFERTY K D and RAIMONDI P T (1996), `Recent declines of black abalone Haliotis cracherodii on the mainland coast of central California', Mar Ecol Prog Ser, 142, 185±192. ANDREWS J D (1984), `Epizootiology of haplosporidian diseases affecting oysters', Comp Pathobiol, 7, 243±269. ANDREWS J D (1996), `History of Perkinsus marinus, a pathogen of oysters in Chesapeake Bay 1950±1984', J Shellfish Res, 15, 13±16. ANDREWS J D and CASTAGNA M (1978), `Epizootiology of Minchinia costalis in susceptible oysters in seaside bays of Virginia's eastern shore 1959±1976', J Invertebr Pathol, 32, 124±138. ANDREWS J D and WOOD J L (1967), `Oyster mortality studies in Virginia. VI. History and distribution of Minchinia nelsoni, a pathogen of oysters in Virginia Chesapeake', Science 8(1), 1±13. ANDREWS J D, WOOD J L and HOESE H D (1962), `Oyster mortality studies in Virginia. III. Epizootiology of a disease caused by Haplosporidium costale Wood and Andrews', J Insect Pathol, 4, 327±343. ADLARD R D
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and WOLF P H (1976), `Fine structure of Marteilia sydneyi sp. n. Haplosporidian pathogen of Australian oysters', J Parasitol, 62, 528±538. PERNAS M, NOVOA B, BERTHE F, TAFALLA C and FIGUERAS A (2001), `Molecular methods for the diagnosis of Marteilia refringens', Bull Eur Ass Pathol, 21(5), 200±208. PICHOT Y, COMPS M, TIGEÂ G, GRIZEL H and RABOUIN M A (1980), `Recherche sur Bonamia ostreae gen. n., sp. n, parasite nouveau de l'huõÃtre plate Ostrea edulis L.', Rev. Trav. Inst. PeÃche Marit, 43(1), 131±140. POWERS D A, KIRBY V L, COLE T and HEREFORD L (1995), `Electroporation as an effective means of introducing DNA into abalone (Haliotis rufescens) embryos', Mol Mar Biol Biotechnol, 4, 369±375. PERKINS F O
RAHMAN M A, RONYAI A, ENGIDAW B Z, JAUNCEY K, HWANG G-L, SMITH A, RODERICK E, PENMAN D, VARADI L and MACLEAN N (2001), `Growth and nutritional trials on transgenic Nile tilapia containing an exogenous fish growth hormone gene', J Fish Biol, 59, 62±78. RENAULT T and NOVOA B (2004), `Viruses infecting bivalve mollusks', Aquat Living Resour, 17(4), 397±409. RENAULT T, STOKES N A, CHOLLET B, COCHENNEC N, BERTHE F and BURRESON E M (2000), `Haplosporidiosis in the Pacific oyster, Crassostrea gigas, from the French Atlantic coast', Dis Aquat Org, 42, 207±214. ROBERT R, BOREL M, PICHOT Y and TRUT G (1991), `Growth and mortality of the European oyster Ostrea edulis in the Bay of Arcachon (France)', Aquat Living Resour, 4, 265±274. ROBLEDO J A F and VASTA G R (2003), `Characterization of the Crassostrea virginica slc11a gene', J Shellfish Res, 22(1), 353. ROBLEDO J A F, COSS C A and VASTA G R (2000), `Characterization of the ribosomal RNA locus of Perkinsus atlanticus and development of a polymerase chain reactionbased diagnostic assay', J Parasitol, 86(5), 972±978. ROCH P (1999), `Defense mechanisms and disease prevention in farmed marine invertebrates', Aquaculture, 172, 125±145. SINDERMANN C J (1990), Principal Diseases of Marine Fish and Shellfish, 2nd Edition, Volume 2 Diseases of Marine Shellfish. San Diego, CA, pp. 103±105 and 443±445, Copyright 1990, with permission from Elsevier. SPARKS A K (2005), `Observations on the history of non-insect invertebrate pathology from the perspective of a participant', J Invertebr Pathol, 89, 67±77. STEPHENSON M F, MCGLADDERY S E, MAILLET M, VENIOT A and MEYER G (2003), `First reported occurrence of MSX in Canada', J Shellfish Res, 22, 355. STICKLER S M, WAGNER E, ENCOMIO V G and LAPEYRE J F (2001), `Natural dermo resistance and its role in the development of hatcheries for the Gulf of Mexico', J Shellfish Res, 20, 557 (Abstract). SUBASINGHE R P, BONDAD-REANTOSO M G and MCGLADDERY S E (2001), `Aquaculture development, health and wealth', in: R P Subasinghe, P Bueno, M J Phillips, C Hough, S E McGladdery and J R Arthur, eds. Aquaculture in the Third Millenium. Technical Proceedings of the Conference on Aquaculture in the Third Millenium, Bangkok, Thailand, 20±25 February 2000. NACA, Bangkok and FAO, Rome, pp. 167±191. SUNILA I, STOKES N A, SMOLOWITZ R, KARNEY R C and BURRESON E M (2002), `Haplosporidium costale (seaside organism), a parasite of the eastern oyster, is present in Long Island Sound', J Shellfish Res, 21, 113±118. THEÂBAULT A, BERGMAN S, POUILLOT R, LE ROUX F and BERTHE F C J, (2005), `Validation of in
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situ hybridisation and histology assays for the detection of the oyster parasite Marteilia refringens', Dis Aquat Org, 65(1), 9±16. VILLALBA A, REECE K S, ORDAS M C, CASAS S M and FIGUERAS A (2004), `Perkinsosis in molluscs: A review', Aquat Living Resour, 17, 411±432. WEI Q (2000), `Country Report. People's Republic of China', in: Asia-Pacific Regional Aquatic Animal Health Management Programme. Final Workshop under FAO/ NACA TCP/RAS/9065 (A) ± `Assistance for Responsible Movement of Live Aquatic Animals' in Cooperation with OIE and the Ministry of Agriculture, China P.R., Bejin, China PR, 27±30 June 2000. China PR National Strategy. WP No. 04. WEIFENG M, YAPING W, WENBO W, BO W, JIANXIN F and ZUOYAN Z (2004), `Enhanced resistance to Aeromonas hydrophila infection and enhanced phagocytic activities in human lactoferrin-transgenic grass carp (Ctenopharyngodon idellus)', Aquaculture, 242, 93±103. WOLF P H (1979), `Life cycle and ecology of Martelia sydneyi in the Australian oyster, Crassostrea commercialis', Mar Fish Rev, 41, 70±72. YARNALL H A, REECE K S, STOKES N A and BURRESON E M (2000), `A quantitative competitive polymerase chain reaction assay for the oyster pathogen Perkinsus marinus', J Parasitol, 86(4), 827±837.
12 Hazard analysis and critical control point programs for raw oyster processing and handling V. Garrido and S. Otwell, University of Florida, USA
Abstract: Bivalve molluscs such as oysters destined for raw consumption share the common problem of potential microbial and chemical contamination of the edible portion of the product due to the natural filtration of surrounding waters that provide essential oxygen and food. This chapter reviews SSOP and HACCP programs for the safe handling of raw oysters Key words: raw oysters, HACCP, SSOP.
12.1
Introduction
The production of molluscan shellfish in the US is divided into four regions: Gulf of Mexico, South Atlantic, North Atlantic and Pacific regions. Each region is very distinct regarding species produced, climate conditions, production methods and growing seasons. In 2005, the US recorded total landings of 29 133 tonnes of shellfish composed primarily of oysters, clams and mussels.1 The Gulf of Mexico region, from the west coast of Florida through Texas, produces approximately 31% of all the wild harvested shellfish as well as a good portion of the aquaculture clams in the United States market. Primarily, the Gulf region specializes in eastern oyster (Crassostera virginica) production and even though the production is year round, the bulk of the harvest is conducted during the colder months of the year (October±April). The second largest producing region is the Pacific region, from California through Washington State, where a variety of shellfish species including pacific oysters (Crassotrea gigas), clams and mussels are produced via wild harvest as well as aquaculture or farming with the
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harvest season going year round. The North Atlantic region, a distant third, has a very limited production owing to a combination of factors such as a very narrow harvest season and a shrinking access to open harvest areas. This region produces primarily mussels, clams and few eastern oysters (approximately 2%). Regardless of the production region, all bivalve molluscs destined for raw consumption share the common problem of potential microbial and chemical contamination of the edible portion of the product due to the natural filtration of surrounding waters that provide essential oxygen and food. In some cases the amount and type of microorganism or chemicals that can accumulate and reside in the mollusc during the filtration can present potential food safety issues for certain consumers that prefer to eat the molluscan products raw. The majority of oyster illnesses involving potential microbial pathogens could be prevented by proper cooking. In most cases, the food safety problem is the recipe rather than the product. Epidemiologically, molluscan shellfish have been linked to illnesses such as Vibrio spp. infections, hepatitis A and Norwalk or Norwalk-like viral infections or shellfish toxins such as amnesic, diarrheic, neurotoxic and paralytic shellfish poisoning.2,3 These food safety problems, except for the Vibrio spp. infections, are related to contaminated harvest areas. Vibrios vulnificus and Vibrio parahaemoliticus are indigenous, naturally occurring organisms that thrive in the marine environment where the bivalves are grown. These bacteria are found to be in higher numbers during the warmer months of the year when there are a higher number of reported molluscan-borne illnesses.4,5 Vibrio bacteria reside in the shellfish and in the environment and can multiply during harvest, transportation and processing if the cold chain is not maintained. Most of the food safety problems associated with raw oysters are easily controlled by harvesting from `approved' waters followed by proper refrigeration until consumed. Based on the history of persistent illnesses due to consumer preference for raw oysters, oyster production/processing is one of the most stringently regulated segments of seafood commerce in the world. In addition to the basic regulatory requirements as for all food processing (current good manufacturing practice (cGMP), sanitation standard operating procedures (SSOP), etc.), oysters are subject to harvest area classifications, routine water quality monitoring and a series of other requirements dictated by the National Shellfish Sanitation Program (NSSP) as part of the Interstate Shellfish Sanitation Conference (www.issc.org) in the US. The NSSP sets the minimum standards for regulatory compliance within interstate commerce as well as guidance on how to accomplish them. These standards range from how to classify the harvest areas depending on rainfall and contaminant impact, time and temperatures from harvest to refrigeration at the primary processor, how to certify and what certification should be given to each processor depending on its activities, and how to process, label and distribute the oysters. The most recent mandate for hazard analysis and critical control point (HACCP) was an additional layer of regulatory controls.
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HACCP principles were developed during the 1960s as part of an optimization of the food safety program for NASA's Space Program.6 Some large food corporations in the US opted to use it as a model to maximize safety and minimize production loss. HACCP was not viewed as a zero-risk system, but it provided additional safety assurance without unreasonable analytical costs. HACCP identified and prevented the occurrence of problems rather than finding them after the fact through traditional quality control evaluations and analysis of the finished products. The intent was to prevent problems rather than rely on detection alone. Although HACCP was initially designed to focus on food safety, some operations and inspection programs decided to use it with a blend of safety and quality attributes. Regardless of the focus or blend, all HACCP programs were based on seven steps or principles as adopted by CODEX7 and the National Advisory Committee on Microbiological Criteria for Foods (NACMCF).8 For the program to work effectively, it is imperative that all seven principles are defined appropriately and that all are based on realistic and science based information. These principles, as defined by the NACMCF,8 are as follows: 1. Conduct hazard analysis ± the process of collecting and evaluating information on hazards associated with the food under consideration to decide which are significant and must be addressed in the HACCP plan. 2. Determine the critical control points (CCP) ± critical control points are steps or areas of the process at which control can be applied and is essential to prevent or eliminate a food safety hazard or to reduce it to an acceptable level. 3. Establish critical limits (CL) ± maximum and/or minimum value to which a biological, chemical or physical parameter must be controlled at a CCP to prevent, eliminate or reduce to an acceptable level the occurrence of a food safety hazard. 4. Monitor each CCP ± to conduct a planned sequence of observations or measurements to assess whether a CCP is under control and to produce an accurate record. 5. Establish corrective actions (CA) ± procedures to be followed when a deviation occurs. 6. Establish verification procedures ± activities, other than monitoring, that determine the validity of the HACCP plan and that verify the system is correctly implemented. 7. Establish record-keeping and documentation procedures ± to maintain written documentation that the monitoring activities are being conducted and that the CCP is under control or a CA was applied. In 1994, the US Food and Drug Administration (FDA) published a proposed rule which mandated that all seafood processors in the US as well as all importers had to develop and implement HACCP base programs to provide more self-control for the safety of their processes and ultimately their products.9 After a series of workshops and comment periods, the FDA published the final
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HACCP regulation in December of 1995 (21 CFR parts 123 and 1240) known as the `Procedures for the Safe and Sanitary Processing and Importing of Fish and Fisheries Products' (http://www.cfsan.fda.gov/~comm/haccpsea.html). They provided a 2-year grace period to allow industry to complete appropriate training, develop their HACCP programs and finish the implementation of the programs. The European Union published and implemented their HACCP regulations in 1991 prior to the US10 and eventually similar HACCP mandates were adopted about most seafood-producing nations.
12.2
HACCP for oyster production and safety
All processors of raw molluscan shellfish are required by local, state and federal regulations to follow some basic foundation or prerequisite programs (Table 12.1). Even though some of these programs do not require a written format, all require monitoring and some have record-keeping mandates. For example, all processors must comply with the Good Manufacturing Practices (GMP), which in the US are published by the FDA under the Code of Federal Regulations Title 21 Part 110 (http://www.cfsan.fda.gov/~acrobat/cfr110.pdf), but no documentation has to be generated. On the other hand, in accordance with the 2005 NSSP Model Ordinance,11 the shellfish industry is required to provide a written recall program to include all activities and steps necessaries to successfully trace and recall potentially hazardous products. Assuming compliance with all of the prerequisite programs, the HACCP program can be developed and implemented focusing only in the food safety aspects of the process and product. Initially, the HACCP mandate assures the program includes complementary SSOP which requires that all the oyster processors maintain SSOP monitoring and record-keeping across eight key sanitation areas and all necessary corrections. The HACCP mandate recommends that all seafood processors should develop and implement a written SSOP but requires written monitoring records of the sanitation activities and all necessary corrections related to the eight key sanitation areas as defined by the same rule (21 CFR part 123.11 sec. 1±8). These key areas are as follows: 1. Safety of the water that comes into contact with food or food-contact surfaces or is used in the manufacturing of ice. 2. Condition and cleanliness of food contact surfaces, including utensils, gloves and outer garments. 3. Prevention of cross-contamination from insanitary objects to food, food packaging material and other food contact surfaces, including utensils, gloves, and outer garments, and from raw products to cooked products. 4. Maintenance of hand washing, hand sanitizing and toilet facilities. 5. Protection of food, food packaging material, and food contact surfaces from adulteration with lubricants, fuel, pesticides, cleaning compounds, sanitizing agents, condensate, and other chemical, physical and biological contaminants.
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6. Proper labeling, storage and use of toxic compounds. 7. Control of employee health conditions that could result in the microbiological contamination of food, food packaging materials and food contact surfaces. 8. Exclusion of pests from the food plant. After successfully developing and implementing all base programs in the processing plant, the processor can develop the HACCP plan. The plan begins with a detailed description of the product or products produced as presented to the costumers, the way the product is distributed and how the product is prepared before consumption. Following the product description, a complete process flow diagram (Fig. 12.1) should be written to include all relevant steps Table 12.1 Prerequisite programs required to provide controls and oversight of the processing plant environment and help build the foundation for the HACCP program Program name
Program description
Mandated or recommended
Recall/traceability program
Description of how to identify, trace back, segregate and remove from the market any product that has been deemed as adulterated or unfit for human consumption
Mandated as written for oyster processors by the NSSP Model Ordinance
Sanitation standard operating procedures (SSOP)
Description of cleaning and sanitation procedures, monitoring and corrections used in the processing plant
Recommended written SSOP program Mandated monitoring and record-keeping of sanitation activities covering the eight key sanitation areas by the FDA's HACCP regulations
Good manufacturing practices
Description of plant environment and personnel in regards to plant infrastructure, equipment and utensil design and construction as well as personnel behavior
Mandatory compliance with cGMP (21 CFR part 110) but no written program is necessary
Master sanitation program
Monitoring and scheduled maintenance of equipment and areas not addressed in the SSOP such as compressors, evaporator pans in cooling units, etc.
Recommended
Training program
Schedule and description of materials to be used in personnel training on a yearly basis
Recommended training schedule for processing and sanitation personnel
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Fig. 12.1
Process flow diagram.
in the processing from procurement of the raw materials through processing, handling and storage to shipping. Although a diagram is not mandated, it is strongly recommended because it can help the processor visualize and confirm the actual processing and handling steps occurring at the plant. In addition, this diagram will be a useful training tool for new employees.
12.3
HACCP plan for processing of frozen raw oysters
Freezing has been introduced as a PHP (post-harvest processing) method to reduce or eliminate potential Vibrio pathogens in oysters destined for raw consumption. Requirements for PHP processing were introduced in the 2001 Interstate Shellfish Sanitation Conference (ISSC) meeting as part of the Vibrio vulnificus (Vv) management plan. This plan has the goal of reducing the rate of Vv illnesses by 60% for the years 2007 and 2008 (average) in comparison with a baseline obtained from the average illness rate obtained from the years 1995±1999 (0.306/
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million).12 The ISSC offers guidance on how to properly evaluate and validate any post-harvest process that targets the reduction of Vibrio in shellfish. According to the 2005 NSSP Guide for Control of Molluscan Shellfish, the process needs to consistently reduce the Vibrio load by 3.52 logs before it could be claimed as a PHP.13 The freezing method selected for this example is intended to yield frozen oysters as whole or half-shell products destined for raw or cooked consumption. The product description is eastern oysters (Crassotrea virginica) harvested from the northern Gulf of Mexico and post-harvest processed to reduce the Vibrio vulnificus loads to non-detectable levels (less than 30 most probable number (MPN)/g) as defined by the FDA and ISSC.13 The methods of storage and distribution would be as frozen product for general public consumption.
12.4
Hazard analysis
Hazard analysis is the initial step of developing an effective HACCP program. The intent is to prioritize the hazards, biological, chemical or physical, that must be monitored and controlled during the handling and production of the oyster product. The hazard analysis should follow all steps described in the process flow-diagram and evaluate each one for all potential food-safety hazards that may be introduced or augmented at each step. A properly conducted hazard analysis should consider all food safety hazards but should only have control measurements for those hazards that are `reasonably likely' to occur. In other words, those that have a good chance of occurring during normal industry handling practices and could cause an illness when the oyster is consumed ± significant safety hazard. As shown in Table 12.2, the hazard analysis should consider all factors related to the product from harvest to consumption but will concentrate on the steps where control could be applied, from receiving the raw ingredients until shipping the finished product. In the example, a series of hazards are identified based on experience, published illness data, regulatory requirements as well as scientific publications, but only few are considered as a significant safety hazard. Control measures are given for those significant safety hazards and a CCP will be defined for its control.
12.5
Identify the critical control points (CCP)
For every significant safety hazard identified during the hazard analysis, there must be one or more CCPs to control the hazard. In the PHP oyster example, three CCPs were identified to control a series of hazards: (1) receiving step controls the potential chemical and microbial contamination from the harvest areas while receiving; (2) cold storage to help maintain the levels of pathogenic bacteria to an acceptable level through temperature controls; and (3) PHP step (finished product frozen storage) controls the targeted pathogen reduction to improve the safety of the finished product.
Table 12.2
Example: Hazard Analysis Worksheet for Shucker Packer Frozen oysters Post-harvest processing (PHP)*
Firm Name: Firm Address: Signature: Date:
Product description:
Nitrogen frozen shellstock ± whole and half-shell
Method of storage and distribution: Intended use & consumer:
Frozen (0 ëF/ÿ18 ëC) Raw/cooked
(1) Ingredient/ Processing step
(2) Identify potential hazards introduced, controlled or enhanced at this step
(3) Are any potential food-safety hazards significant? (Yes/No)
(4) Justify your decision for column 3
(5) What preventative measure(s) can be applied to prevent the significant hazards?
(6) Is this step a critical control point? (Yes/No)
Receiving
Biological: Pathogenic bacteria
Yes
Pathogenic bacteria can be concentrated by shellstock in their natural environment. Shellstock are consumed raw.
Only receive shellstock harvested from Yes approved waters and properly tagged by a licensed harvester. Only receive shellstock product from certified dealers listed in current Interstate Certified Shellfish Shippers List (ICSSL).
Virus
Yes
Shellstock are known to concentrate environmental contaminants and have been related to viral outbreaks.
Only receive shellstock harvested from Yes approved waters and properly tagged by a licensed harvester. Only receive shellstock product from
Pathogen growth
Yes
Shellstock are assumed to be eaten raw. Pathogens could grow if time/temperature abuse occurred during harvest and transportation.
certified dealers listed in current ICSSL. Yes Only receive shellstock product properly refrigerated at 45 ëF or below if from a certified dealer or within the time-temperature matrix if from a harvester.
Chemical: Natural toxins
Yes
Shellstock are known to concentrate environmental contaminants.
Yes
Industrial and boat chemicals could contaminate the shellstock during harvest and transport.
Only receive shellstock harvested from Yes approved waters and properly tagged by a licensed harvester. Only receive shellstock product from certified dealers listed in current ICSSL. Only receive shellstock harvested from Yes approved waters and properly tagged by a licensed harvester. Only receive shellstock product from certified dealers listed in current ICSSL.
Yes
Thermal abuse could increase pathogenic bacteria.
Chemical contamination
Physical: None Product cold storage
Biological: Pathogen growth Physical: None Chemical: None
Maintain storage cooler temperature below 45 ëF.
* This hazard analysis and HACCP plan are for illustrative purposes only and may not be applicable to all PHP operations.
Yes
Table 12.2 Ingredient/ Processing step
Continued Identify potential hazards introduced, controlled or enhanced at this step
Processing ± Biological: washing/grading/ Pathogen growth culling shellstock Chemical: None
Are any potential food-safety hazards significant? (Yes/No)
Justify your decision for column 3
No
Time in this operation is too short to promote significant pathogen growth.
No
Time±temperature is not likely to occur due to rapid freezing step. Will be controlled by effective sanitation and good manufacturing practices.
Physical: None Freezing ± post- Biological: harvest processing Pathogen growth (PHP) Pathogen contamination Chemical: None Physical: None
No
What preventative measure(s) can be applied to prevent the significant hazards?
Is this step a critical control point? (Yes/No)
Culling, shucking, Biological: glazing, and Pathogen growth packing
No
Not likely to occur due to low temperatures.
No
Will be controlled by effective sanitation and good manufacturing practices.
Biological: Pathogen growth
No
Not likely to occur due to low frozen temperatures.
Vibrio reduction
Yes
If product is not stored for the validated length of time, vibrios will not be reduced to nondetectable levels.
Pathogen contamination Chemical: None Physical: None Finished product frozen storage
Chemical: None Physical: None
Store frozen oysters for time determined by validation study.
Yes
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12.6
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Definition of the critical limits (CL)
During this step, critical limits are assigned for each CCP to serve as guidance to assure control of the monitored parameters. The identified CL can be one parameter, such as maximum refrigeration temperatures, that would minimize the growth of the targeted bacteria, or multiple parameters such as harvest area, time of harvest and product temperature to assure an acceptable bacteria level in the oysters. In the oyster PHP HACCP example (Table 12.3), several CLs are defined for each CCP; at the receiving step, for example, only product harvested from an open area or processed by a certified dealer is acceptable to be received. Also we identify the temperature of the product and/or conveyance as 7 ëC (45 ëF) as appropriate if the product is received from a certified dealer or the proper time interval from harvest to delivery if the product is received directly from the harvester. At cold storage, a maximum temperature of 7 ëC (45 ëF) is required in the cooler (2005 NSSP Model Ordinance). During the identified PHP freezing step a minimum time in the freezer is necessary to reduce the targeted bacterial pathogen to acceptable levels.
12.7
Designate monitoring procedures
The monitoring procedures must specify four things: (1) what parameter will be monitored, (2) how will it be done, (3) when or how frequently it will be conducted, and (4) by whom? What will be monitored is defined by the control measure that was elected during the hazard analysis. It is the parameter to be monitored, usually being a measurement or observation that will render a real-time result. Usually the monitoring activities rely on physical or chemical measurements, or visual observations that will give the processor an immediate or rapid result. Microbiological analysis is not usually used during monitoring for the high cost and time dependent reactions, taking up to 6 days to yield a confirmatory result. Microbiological analyses are normally reserved for validation and verification activities. In the PHP oyster example, the What is the oyster tag and ambient temperature of the truck at receiving, ambient temperature in the cold storage room, and duration of frozen storage as defined by a previous process validation. How the monitoring will be conducted will also be defined by the control measurements adopted during the hazard analysis. If temperature is the parameter to monitor, then a calibrated thermometer needs to be used. In the example, the company uses calibrated thermometers and visual observations to accomplish their monitoring requirements. When defines the frequency of monitoring. The ideal monitoring frequency is continuous to better determine trend and to prevent deviations before the CL is reached. Continuous monitoring is preferred but not always possible owing to the nature of the activity or operational cost, so the frequency needs to be determined based on the variability of the parameter being monitored and how close to the critical limits is the operating level. In
Table 12.3
Example: Model HACCP Plan for Frozen Oysters ± PHP Product description: Frozen shellstock ± Whole and half-shell
Firm Name: Firm Address:
Method of storage and distribution: Frozen at 0 ëF (ÿ18 ëC) Consumer and intended use: By the general public ± To be eaten raw or fully cooked
Signature: Date: Critical control point (CCP)
Significant hazard(s)
Critical limits for each Preventive Measure
Receiving
· Pathogen growth and/or contamination
· Open water harvest
· Virus · Natural toxins · Chemical contamination
Monitoring What
· Truck temperature no higher than 45 ëF
· Obtain from state shellfish agency by · Harvester tag w/ phone/fax/ complete information internet and (including harvest document harvest area closures time) and re-openings · Harvester license for applicable or aquaculture harvest areas certificate · Visual check · Proper dealer tag and dealer listing in current ICSSL
(Receipt from certified dealers)
· Truck temperature at receipt
· Harvest tag or dealer tag properly completed (harvest times within the time matrix) · Must be licensed harvester or certified dealer
· Open water harvest
How
· Harvest date
Frequency
Who
· Each closure and re-opening
Area supervisor
· 5 containers per lot container per delivery
Corrective action
Verification
Records
Reject oysters that do not meet critical limits or redirect oysters not in compliance with the time-torefrigeration matrix to `for shucking only' application
· Quarterly calibration of temperature monitoring device
· Harvest area status record/ closure record
· Records review within one week of activity
· Receiving temperature record
· Receiving record
· Thermometer calibration/ accuracy verification record
Table 12.3
Continued Monitoring
Critical control point (CCP)
Significant hazard(s)
Critical limits for each Preventive Measure
Shellstock storage cooler
Pathogen growth
Cooler temperature shall be no higher than 45 ëF
Finished product frozen storage
Vibrio vulnificus For Vibrio vulnificus Product freezing reduction (Vv) levels of less date and inventory than 30 MPN/g, hold rotation product in frozen storage for a minimum of 28 days For Vv levels of less than 3 MPN/g, hold product in frozen storage for a minimum of 56 days
Corrective action
Verification
Records
What
How
Frequency
Who
Cooler temperature
Visual check of cooler thermometer
Twice daily during working hours & once daily during non-operating days when product is in cooler
Area supervisor
· Follow corrective actions listed belowy
· Quarterly calibration of cooler temperature monitoring device, or verification with calibrated thermometer · Records review within one week of activity
· Cooler temperature record · Thermometer calibration record · Corrective action record
Area supervisor
· If critical limit is not met, extend the frozen storage time until it meets the critical limit
· Weekly records review · Monthly end product testing for nondetectable Vvz · Quarterly calibration of temperature monitoring device
· Frozen inventory record · Corrective action record · Verification record · Thermometer calibration record · PHP validation report
Inventory control Every lot (date-in and dateout of storage)
* This hazard analysis and HACCP plan are for illustrative purposes only and may not be applicable to all PHP operations y Corrective action shellstock: If cooler temperature is >45 ëF, check meat temperature. If shellstock meat temperature is between 45 and 50 ëF, the product will be placed in a working cooler at 45 ëF or less. If cooler temperature is >45 ëF and shellstock meat temperature is >50 ëF the product will be will be placed in a working cooler at 45 ëF or less, isolated and evaluated to insure the product is safe. z Three samples collected after validation storage time in frozen storage and tested for Vibrio vulnificus levels (ISSC defines non-detectable as less than 30 MPN/g Vv while California defines non-detectable as less than 3 MPN/g).
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the PHP example, the frequency varies from each occurrence to twice daily depending on the CCP. Who is defined as the best trained individual to conduct the activity. The person charged with this task needs to have some understanding of how the parameter will be monitored and how to react in case a deviation is encountered at the time of the activity. An individual name is not necessary because a specific person may not be available at all times. Typically, area supervisor, quality control personnel or dock master are the designations you will find in a HACCP plan. In the example, the area supervisor is the position in charge of conducting the monitoring.
12.8
Corrective action (CA)
Every time a deviation from the CL occurs, a CA needs to be taken and a record must be generated. The corrective actions can be from as simple as evaluate and release, all the way to destroying the affected products depending on the magnitude and nature of the deviation. Here are some recommended steps to be followed: · · · · · ·
Identify and isolate the affected product. Evaluate the affected product. Take necessary measurements on the affected products. Destroy the affected products that cannot be handled by any other way. Identifying the cause of the problem. Apply a temporary and permanent correction of the process.
12.9
Specify verification (and validation) procedures
Verification procedures include a series of procedures that are separated by time of occurrence. The activities that occur `before' the HACCP plan is implemented are called validation, the activities that occur `during' the implementation are verification. All these activities are executed to determine if the plan has been developed and implemented correctly. These activities include: · · · · · ·
Evaluation of the HACCP elements. Equipment and process validation. Monitoring equipment calibrations. Record review. Targeted in-process and finished product sampling and analysis. Process and program audits.
The US FDA regulation defines specific frequencies for the validation and verification activities such as yearly HACCP plan evaluation and at least weekly
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record reviews. Monitoring equipment calibration frequency is left to the manufacturer's recommendation. In the PHP example, the verification activities recommended are weekly record reviews, monthly microbial analysis of finished product and quarterly monitoring for equipment calibration. Process and equipment validations were conducted prior to implementation of the HACCP plan and are kept on file for future reference or evidence to justify selection of a particular process or practice, e.g. the validation of nitrogen freezing and frozen storage as an effective process to reduce Vibrio vulnificus in oysters.
12.10
Specified records
Records are the heart of the HACCP program. They give the processor a reference to better optimize operations by tracking performance and changes, plus providing evidence for the execution of the HACP plan. HACCP records can include the HACCP plan and base documents (SSOP, GMP, recall, etc.), monitoring records, CA records and verification records. The Appendix includes examples of HACCP and sanitation records.
12.11 1.
References
(2007). Annual Commercial Landing Statistics ± http://www.st.nmfs.gov/st1/commercial/landings/annual_landings.html 2. FOOD AND DRUG ADMINISTRATION (2001). Fish and Fisheries Products Hazards and Controls Guidance: Third Edition. http://www.cfsan.fda.gov/~comm/haccp4.html 3. NATIONAL SEAFOOD HACCP ALLIANCE FOR TRAINING AND EDUCATION (2006). Compendium of Fish and Fisheries Product Processes, Hazards and Controls http://seafood.ucdavis.edu/haccp/compendium/compend.htm 4. WORLD HEALTH ORGANIZATION (WHO) AND FOOD AND AGRICULTURE ORGANIZATION (FAO) (2005). Risk assessment of Vibrio vulnificus in raw oysters: interpretative summary and technical report. http://www.who.int/foodsafety/publications/micro/ mra8.pdf 5. FOOD AND DRUG ADMINISTRATION (2005). Quantitative Risk Assessment on the Public Health Impact of Pathogenic Vibrio parahaemolyticus in Raw Oysters ± http://www.cfsan.fda.gov/~acrobat/vpra.pdf 6. FAO (1998). Food Quality and Safety Systems ± A Training Manual on Food Hygiene and the Hazard Analysis and Critical Control Point (HACCP) System ± Section 3 7. FAO/WHO (2001). Codex Alimentarius ± Food Hygiene ± Basic Texts, Second Edition 8. NATIONAL ADVISORY COMMITTEE ON MICROBIOLOGICAL CRITERIA FOR FOODS (NACMCF) (1998). Hazard analysis and critical control point principles and application guidelines. Journal of Food Protection, Vol. 61, No. 9, pp. 1246±1259 9. FOOD AND DRUG ADMINISTRATION (1994). Proposal To Establish Procedures for the Safe Processing and Importing of Fish and Fishery Products; Proposed Rule. Federal Register: January 28. http://www.cfsan.fda.gov/~lrd/fr940128.html NATIONAL MARINE FISHERIES SERVICES
HACCP programs for raw oyster processing and handling 10.
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EUROPEAN UNION COUNCIL DIRECTIVE,
22 July 1991. Laying down the Health Conditions for the Production and the Placing on the Market of Fishery Products 91/ 493/EEC Directive. http://www.cfsan.fda.gov/~acrobat/hp91493.pdf 11. FOOD AND DRUG ADMINISTRATION (2005). National Shellfish Sanitation Program (NSSP) Guide for the Control of Molluscan Shellfish. http://www.cfsan.fda.gov/ ~ear/nss3-toc.html 12. FOOD AND DRUG ADMINISTRATION (2005). NSSP Guide for the control of Molluscan Shellfish. IV. Guidance Documents, chapter IV. Naturally Occurring Pathogens. http://www.cfsan.fda.gov/~ear/nss3-44.html#p04 13. FOOD AND DRUG ADMINISTRATION (2005). NSSP Guide for the control of Molluscan Shellfish. II. Model Ordinance XVI. Post Harvest Processing. http:// www.cfsan.fda.gov/~ear/nss3or16.html
Appendix: Examples of HACCP and sanitation records The following pages show examples of HACCP and sanitation documentation.
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13 Biofouling and the shellfish industry D. I. Watson, University College Cork, Ireland and S. E. Shumway and R. B. Whitlatch, University of Connecticut, USA
Abstract: The impacts of biofouling on shellfish and aquaculture can be extreme and sometimes devastating. Biofouling is a major problem within the shellfish aquaculture industry, particularly with regard to the resultant increased labor costs and reduced value of product. The chapter discusses the impacts of biofouling, economic costs involved during culture, processing, the value of the end product, and the techniques employed to mitigate and remove biofouling organisms. While biofouling can have drastic impacts on the quality of product, most issues are associated with marketing and reduced aesthetic value. Biofouling does not impact the safety of the product for human consumption. Key words: biofouling, shellfish, seafood quality, shellfish safety.
13.1
Introduction
Impacts of biofouling on molluscs (= shellfish) and aquaculture can be extreme and sometimes devastating. Biofouling affects all structures, both natural and artificial, immersed in the marine environment, although often in different ways (Glasby, 1999; Glasby and Connell, 1999; Railkin, 2004). Living organisms, such as gastropod and bivalve molluscs, can provide a massive amount of hard substratum that can be settled upon by larvae of marine organisms. In fact, in many benthic and intertidal areas, the surface area provided by shellfish may be equal to, or greater than, the inert substratum available (Railkin, 2004). Settling sedentary organisms generally recruit to both living and inert substrata (D'Antonio, 1985). Such biofouling of living substrata by both epibionts and endobionts is a major problem within the shellfish aquaculture sector, particularly with regard to the resultant increased labor costs and reduced value
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of product (Adams et al., in preparation). While bivalves and other marine species have various natural defensive techniques to reduce recruitment of epiand endobionts (see Wahl et al., 1998), shells invariably end up fouled by marine organisms ± this fouling may in turn facilitate the settlement of certain epibionts that capitalise on the protective abilities of the bivalves (Forester, 1979, Pitcher and Butler, 1987). Within the shellfish aquaculture sector epi/endobionts affect cultured organisms in different ways (see below), and can seriously affect the costs involved during culture or processing; biofouling may also significantly affect the value of the end product (Adams et al., in preparation). In this chapter we will discuss the impacts of biofouling on molluscan shellfish handling and safety, and the techniques employed to mitigate and remove biofouling organisms
13.2
Biofouling and shellfish
13.2.1 Affected bivalves Impacts of biofouling on shellfish and shellfish culture vary with geographic location, shellfish species, habitat, and method of culture. Impacts of fouling differ between the intertidal (e.g., mussels, oysters), subtidal (e.g., scallops, oysters, mussels, abalone), and soft-sediments (e.g., scallops) (see Plate I). Permanently immersed animals and culture gear tend to be more heavily fouled than others. Bivalves which are infaunal (e.g., clams) have few or no epibionts in comparison with exposed shellfish which can provide a substratum for the larvae of fouling species. Intensity of biofouling can vary between the two valves of an individual bivalve, and overall intensity of valve fouling can differ between wild and cultured bivalves. Often, cultured organisms must be removed from their natural conditions so that they can be cultured in a more economically feasible fashion. This has important implications for development of fouling communities on the shells. For instance, the scallop Pecten maximus is a benthic species that lives on soft-sediment with its lower valve imbedded within the sediment and with its upper valve covered by a thin sediment layer. The valves (particularly the lower) are, therefore, essentially unavailable as a settlement platform for biofouling larvae (Lodeiros and Himmelman, 2000) with wild stocks only being fouled on the upper valve (Ivin et al., 2006; Schejter and Bremec, 2007). The lower valve, however, is not entirely unavailable to certain species, such as some endobionts (Evans, 1969). When scallops are cultured off the seafloor in trays or nets, both valves become available to biofouling organisms. Biofouling in culture situations can lead to increased stress or feeding competition for the scallops (VeÂlez et al., 1995) which may result in reduced growth (Lodeiros and Himmelman, 2000). Such valve specificity by foulers is also seen in some species of oysters where the sponges Cliona viridis and C. celata infest only the lower valve of the flat oyster, Ostrea edulis (Rosell et al., 1999). Such potential for differences in
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the impacts of fouling among individuals of species growing either naturally in beds or within aquaculture structures have financial implications with regard to harvest, maintenance and marketability. 13.2.2 Type and extent of fouling As noted by Wahl et al. (1998), few species are solely involved in the fouling of other living organisms and all groups of fouling species, be they shelled (e.g., barnacles, serpulid worms, and other bivalves), soft-bodied foulers (e.g., ascidians), or borers (e.g., boring sponges) can impact wild and cultured bivalves and their fisheries. Fouling organisms can be species-specific with regard to host selection (Rosell et al., 1999). Sponge endobionts were rare on mussel shells (Mytilus edulis), but prevalent on oyster shells (O. edulis) from similar water depths; thus such species-specificity is not necessarily controlled by the environment in which the bivalves are found. The extent of fouling often varies with the age of the bivalve (Scardino et al., 2003) and the ability of a given bivalve species to defend itself against foulers. Fouling species can be attracted to, or inhibited by, certain conditions on the substrata including light and shade conditions, color (James and Underwood, 1994), and surface texture. Fisher (pers. comm.) noted that on the exterior surface of the oyster shells, worm penetration sites were routinely associated with crevices and growth ring (radial ridges) areas in the shell profile, and suggested that these changes in the shell profile may facilitate recruitment of Polydora planktonic larvae during the settling out stage of worm life history) and chemistry (Lapointe and Bourget, 1999; Thomason et al., 1994). These factors can be particularly important during the early developmental stages of a fouling community (Bourget et al., 1994). Shell surface texture and chemistry appear to play an important role in defining the settlement levels of fouling organisms; however, this is very dependent upon the scale of surface roughness encountered (see Scardino et al., 2003 and references therein). No single surface texture appears to prevent recruitment of a wide range of fouling organisms, and certain organisms even prefer to recruit to smooth surfaces (Scardino and de Nys, 2004). Many bivalves possess physical or behavioral defences against fouling. The periostracum of bivalve shells is composed of a proteinaceous matrix, which Wahl et al. (1998) proposed as a potential defensive mechanism against fouling organisms and may also assist in the preventative formation of biofilms (Scardino and de Nys, 2004). The periostracum can be textured (e.g., Mytilus galloprovincialis) or smooth (e.g., Pinctada imbricate) (Scardino et al., 2003). Although no single periostracum texture will inhibit all foulers, certain speciesspecific interactions have been identified for some bivalves (Bers and Wahl, 2004; Bers et al., 2005). The periostracum also acts as a physical barrier against some boring organisms (Harper and Skelton, 1993) and it is this property which is being emulated in the development of new synthetic coatings designed to protect pearl oysters from biofouling (de Nys and Ison, 2004). It has also been
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noted that the presence and effectiveness of this system in protecting bivalves against fouling organisms diminishes with age in some species (Guenther et al., 2006; Scardino et al., 2003; Wahl et al., 1998) through the deterioration of this barrier by environmental factors. Other systems employed by bivalves include the cleaning of their shells and the facilitation of certain foulers that deter the recruitment of other fouling organisms. Foot-sweeping among bivalves, such as mussels, assists in the physical removal of new settlers (Thiesen, 1972), and reduces the build-up of a fouling community on the shell. This behavior is most effective in small mussels (100 000
50 75 110 130 140 140 145 145 150
35 45 50 55 55 55 60 60 60
20 23 25 26 27 27 27 27 30
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prevalence (usually 2, 5, or 10%). Using a population of 1000 candidate SPF shrimp stock in primary quarantine as an example, the table indicates that 130 and 55 specimens will be required for testing for prevalence rates of 2% and 5%, respectively. Because of their high sensitivity and specificity, samples taken for molecular (PCR/RT-PCR) or antibody-based tests may be combined as pooled samples of no more than five individuals for testing. Thus, the 130 and 55 samples from the above illustration will constitute 26 and 11, respectively, pooled samples for actual testing. With expensive assays such as PCR/RT-PCR this constitutes a very significant savings in diagnostic testing. The program, FreeCalc Version2 will calculate the sample size required for different levels of statistical confidence at assumed prevalence rates in populations. FreeCalc can also adjust sample size when the sensitivity and specificity of the screening test to be used is known to be less than 100%. FreeCalc Version 2 and the supporting documentation is available from the Australian Centre for International Agricultural Research (ACIAR) at http://www.aciar.gov.au/web.nsf/doc/JFRN5J46ZY. In the context of SPF stock development following the ICES guidelines (Sindermann, 1988, 1990) (Table 16.4 and Fig. 16.7), samples are taken at specific times during primary and secondary quarantine and tested for the specific pathogens listed in Tables 16.1 and 16.2. This activity is within the definition of targeted surveillance. Unusual pathologies or unexpected mortalities, however, are also sampled and investigated in SPF stock development, which is an activity that falls within the definition of general (passive) surveillance. After the criteria set forth in the ICES guidelines have been met and a particular stock is declared SPF of specified diseases/pathogens, maintenance of SPF status requires that the domesticated SPF stocks be the subject of a routine surveillance program (Lightner, 2003b, 2005; OIE, 2006a). To be functional, a SPF breeding program must have a surveillance program with both targeted and general (passive) surveillance components.
16.6
Biosecurity and the culture of wild seed/broodstock
Although some applications of biosecurity principles are possible in an industry that uses wild stocks for seed production, consistency in preventing the introduction of diseases and pathogens is problematic because of a variety of problems inherent in having laboratory testing performed. Such problems may include limitations to the accuracy and sensitivity of the test(s) used, representative sampling and sample sizes needed for statistical confidence, and problems with getting the required samples to diagnostic laboratories, tested, and reported within what is often a relatively short period of time between the time the wild seed stock is collected or spawned and the time by which transport and/or stocking must occur. Furthermore, the prevalence and severity of infection of significant pathogens in wild populations may be quite low, making their detection a difficult task. These factors lead frequently to false
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negative results when wild stocks (nauplii, larvae, postlarvae, or broodstock) are sampled and screened using even the most sensitive molecular methods available. Hence, while more sensitive and accurate diagnostic tests are becoming available each year, no test is likely ever to be 100% accurate (OIE, 2006b). The best way to be sure of the pathogen status of any given shrimp stock is have control of the stock and to monitor it for specific pathogens over time, thus building a documented history of that particular stock as being free of specific pathogens. This is the concept of programs that develop domesticated lines of SPF shrimp, and this is the principal reason why domesticated lines of the Pacific white shrimp, Litopenaeus vannamei, so rapidly became the dominant shrimp species farmed in the world within 5 years after their introduction to East and SE Asia (FAO, 2006).
16.7 Biosecurity through environmental control and best management practices A variety of environmental and best management practice strategies have been adopted for the control of viral and other significant excludable diseases in penaeid shrimp aquaculture (Lee and O'Bryen, 2003; Lightner, 2005; Scarfe et al., 2006). These strategies range from the use of improved culture practices (i.e., where sources of virus contamination are reduced or eliminated, source water is treated, filtered, and aged to remove potential vectors, culture ponds are cleaned, plowed, and fallowed and treated between crops, routine sanitation practices are improved, stocking densities are reduced, etc.) to stocking domesticated SPF or SPR shrimp stocks. Some opportunistic disease agents (e.g., certain Vibrio spp.) are part of the shrimp's normal microflora, but can become deadly pathogens in `stressed' shrimp. `Stress' in shrimp is a poorly defined condition that is difficult to measure, and it has more causes that are not well understood. Its causes can range from the shrimp being subjected to environmental extremes to over- or under-feeding. Most penaeid shrimp have the best culture performance (i.e., growth and food conversion efficiency) at water temperatures near their upper tolerance limit for a particular life stage of the species (Lightner, 2003b, 2005). Farms and management practices must recognize this factor, be tailored to benefit from the effect, and mitigate it when `stress' and disease could result from water temperatures becoming too high for too long. Hence, the farm siting, culture system design, the quality of feed used, stocking density, the farm's routine management practices, and other factors can have a profound effect on the amount of `stress' to which farmed shrimp stocks are subjected. Therefore, diseases due to `abiotic' agents (i.e., `stress', toxicants, environmental extremes, nutritional imbalances, etc.), or those due to opportunistic `biotic' agents that are either commonly present in the culture environment or part of the shrimp's normal microflora, are not excludable and should not be among the listed disease agents to be excluded in a biosecurity plan for a facility, compartment, zone, etc., or in an SPF stock domestication and
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development program. The management of such diseases, however, through farm design, the use of appropriate feeds and feed application, and the quality of overall management are nonetheless essential components of successful shrimp farming. Because these topics are beyond the scope of the present review, the authors refer the reader to reviews published elsewhere in which this topic has been thoroughly reviewed (Browdy and Jory, 2001; Lee and O'Bryen, 2003; Scarfe et al., 2006).
16.8
Conclusions
In the wake of the epizootics, due principally to the shrimp viruses TSV and WSSV that swept through the main penaeid shrimp growing regions of both Asia and the Americas, there has been a paradigm change in what the industry farms and how it is done. SPF, SPR, and biosecurity were terms seldom heard in shrimp farming establishments a decade ago, but today these terms, and the concepts and practices they represent, are increasingly being applied by the global shrimp farming industry. The application of SPF, SPR, and biosecurity concepts to many of the existing types of shrimp farming, as they have been applied to poultry for example, is not something that can be accomplished easily or in the short term. The industry has thousands of hectares of farms and hundreds of hatcheries (Rosenberry, 2001) which were not designed to afford managers with much of an opportunity to totally prevent particular pathogens from being introduced and becoming established, or to exclude them during normal farming activities even if SPF shrimp are stocked. Nonetheless, with the use of applicable elements of the concepts of SPF, SPR, and biosecurity much can be done to reduce losses due to particular pathogens by utilizing `seed stocks' that are free of the major pathogens of concern and by modifying existing farms and their management routines to apply biosecure practices. The progress made by the shrimp farming industry by changing from farming mostly shrimp stocks obtained directly or indirectly from wild stocks to culturing domesticated stocks of L. vannamei which are SPF for the major shrimp diseases and which have been improved through selective breeding for desirable performance characteristics including disease resistance, has contributed significantly to making the industry more sustainable and environmentally responsible. These remarkable changes and the trend away from farming wild shrimp to culturing only domesticated SPF shrimp are likely to continue well into future.
16.9
Acknowledgments
Grant support for the author of this review was provided by the United States Marine Shrimp Farming Consortium under Grant No. 2004-38808-02142 and from Hatch Project ARZT-136860-H-02-135 (both through the Cooperative State Research, Education and Extension Service (CSREES), US Department of
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Agriculture, and special grants from Darden Restaurants (Orlando, FL) and Morrison Enterprises (Hastings, NE).
16.10
References and further reading
(1983). Construction of broodstock, maturation, spawning, and hatching systems for penaeid shrimps in the Centre Oceanologique du Pacifique. In: J.P. McVey (ed.) CRC Handbook of Mariculture. Volume 1 Crustacean Aquaculture. CRC Press, Boca Raton, FL, pp. 105±121. ARGUE, B. and A. ALCIVAR-WARREN (1999). Genetics and breeding applied to the penaeid shrimp farming industry. In R.A. Bullis and G.D. Pruder (eds) Controlled and Biosecure Production Systems. Evolution and Integration of Shrimp and Chicken Models. Proceedings of a Special Session, Sydney, Australia, 27±30 April 1999. World Aquaculture Society, Baton Rouge, LA, pp. 29±53. BARTLEY, D.M., R. SUBASINGHE, and D. COATES (1996). Draft Framework for the responsible use of introduced species. European Inland Fisheries Advisory Commission. EIFAC/XIX/96/Inf.8. BELL, T.A. and D.V. LIGHTNER (1984). IHHN virus: infectivity and pathogenicity studies in Penaeus stylirostris and Penaeus vannamei. Aquaculture 38: 185±194. BELL, T.A. and D.V. LIGHTNER (1988). A Handbook of Normal Shrimp Histology. Special Publication No. 1, World Aquaculture Society, Baton Rouge, LA. BONDAD-REANTASO, M.G., S.E. MCGLADDERY, I. EAST, and R.P. SUBASINGHE (2001). Asia Diagnostic Guide to Aquatic Animal Diseases. FAO Fisheries Technical Paper 402/2, Food and Agriculture Organization of the United Nations, Rome, Italy. BRADLEY-DUNLOP, D., C.R. PANTOJA, and D.V. LIGHTNER (2004). Development of monoclonal antibodies for detecting necrotizing hepatopancreatitis in penaeid shrimp. Dis. Aquatic Organisms 60: 233±240. BRAY, W.A., A.L. LAWRENCE, and J.R. LEUNG-TRUJILLO (1994). The effect of salinity on growth and survival of Penaeus vannamei, with observations on the interaction of IHHN virus and salinity. Aquaculture 122: 133±146. BROCK, J.A. (1991). An overview of diseases of cultured crustaceans in the Asia Pacific region. In Fish Health Management in Asia-Pacific. Report on a Regional Study and Workshop on Fish Disease and Fish Health Management. ADB Agriculture Department Report Series No. 1. Network of Aquaculture Centres in Asia-Pacific. Bangkok, Thailand, pp 347±395. BROCK, J.A. (1992). Current diagnostic methods for agents and diseases of farmed marine shrimp. In W. Fulks and K. Main (eds) Proceedings of the Asian Interchange Program Workshop on the Diseases of Cultured Penaeid Shrimp in Asia and the United States, Honolulu, HI, 27±30 April 1992. The Oceanic Institute, Honolulu, HI, pp. 209±231. BROCK, J.A. (1997). Special topic review: Taura syndrome, a disease important to shrimp farms in the Americas. World J. Microbiol. Biotechnol. 13: 415±418. BROCK, J.A. and B. LEAMASTER (1992). A look at the principal bacterial, fungal and parasitic diseases of farmed shrimp. In J. Wyban (ed.) Proceedings of the Special Session on Shrimp Farming, Orlando, Florida, 22±25 May 1992, World Aquaculture Society, Baton Rouge, LA, pp. 212±226. BROCK, J.A., and D.V. LIGHTNER (1990). Diseases of crustacea. Diseases caused by AQUACOP
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microorganisms. In O. Kinne (ed.) Diseases of Marine Animals, Vol. III, Biologische Anstalt Helgoland, Hamburg, Germany, pp. 245±349. BROCK, J.A. and K. MAIN (1994). A Guide to the Common Problems and Diseases of Cultured Penaeus vannamei. Oceanic Institute, Honolulu, HI. BROCK, J.A., D.V. LIGHTNER, and T.A. BELL (1983). A review of four virus (BP, MBV, BMN, and IHHNV) diseases of penaeid shrimp with particular reference to clinical significance, diagnosis and control in shrimp aquaculture. Proceedings of the 71st International Council for the Exploration of the Sea, C.M. 1983/Gen:10/1±18. BROWDY, C.L. and D.E. JORY (EDS) (2001). The New Wave, Proceedings of the Special Session on Sustainable Shrimp Culture 2001. The World Aquaculture Society, Baton Rouge, LA. BROWDY, C.L., J.D. HOLLOWAY, C.O. KING, A.D. STOKES, J.S. HOPKINS, and P.A. SANDIFER (1993). IHHN virus and intensive culture of Penaeus vannamei: effects of stocking density and water exchange rates. J. Crustacean Biol. 13: 87±94. BULLIS, R.A. and G.D. PRUDER (EDS) (1999). Shrimp Biosecurity: Pathogens and Pathogen Exclusion. Controlled and Biosecure Production Systems. Evolution and Integration of Shrimp and Chicken Models. Proceeding of a Special Session, World Aquaculture Society, Sydney, Australia, Oceanic Institute, Honolulu, HI. CARR, W.H., J.M. SWEENEY, and J.S. SWINGLE (1994). The Oceanic Institute's specific pathogen free (SPF) shrimp breeding program: preparation and infrastructure. In U.S. Marine Shrimp Farming Program 10th Anniversary Review, Gulf Coast Research Laboratory Special Publication. Ocean Springs, MS. Gulf Research Reports 1, pp. 47±54. CARR, W.H., J.N. SWEENEY, L.M. NUNAN, D.V. LIGHTNER, H. H. HIRSCH, and J.J. REDDINGTON (1996). (The) use of an infectious hypodermal and hematopoietic necrosis virus gene probe serodiagnostic field kit for the screening of candidate specific pathogen-free Penaeus vannamei broodstock. Aquaculture 147: 1±8. CHANRATCHAKOOL, P., J.F. TURNBULL, S.J. FUNGE-SMITH, I.H. MACRAE, and C. LIMSUWAN (1998). Health Management in Shrimp Ponds. Aquatic Animal Health Research Institute, Dept. of Fisheries, Kasetsart University, Bangkok, Thailand. CRANE, M. ST. J. and J.A.H. BENZIE (EDS) (1999). Proceedings of the Aquaculture CRC International Workshop on Invertebrate Cell Culture, 2±4 November 1997, University of Technology, Sydney, Australia. Methods Cell Science 21(4): 171±272. ERICKSON, H.S., M. ZARIN-HERZBERG, and D.V. LIGHTNER (2002). Detection of Taura syndrome virus (TSV) strain differences using selected diagnostic methods: diagnostic implications in penaeid shrimp. Dis. Aquatic Organisms 52: 1±10. FAO (2006). State of World Aquaculture. FAO Fisheries Technical Paper 500, Food and Agriculture Organization of the United Nations, Rome. FAO/NACA/UNEP/WB/WWF (2006). International Principles for Responsible Shrimp Farming. Network of Aquaculture Centres in Asia-Pacific (NACA). Bangkok, Thailand. FAUQUET C.M., M.A. MAYO, J. MANILOFF, U. DESSELBERGER, and L.A. BALL (2005). Virus Taxonomy. Classification and Nomenclature of Viruses. Eighth Report of the International Committee on Taxonomy of Viruses. Elsevier Academic Press, Amsterdam. FEGAN, D.F. and H.C. CLIFFORD III (2001). Health management for viral diseases in shrimp farms. In C.L. Browdy and D.E. Jory (eds) The New Wave, Proceedings of the Special Session on Sustainable Shrimp Culture. Aquaculture 2001. The World Aquaculture Society, Baton Rouge, LA, pp. 168±198.
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(1997). Major viral diseases of the black tiger prawn (Penaeus monodon) in Thailand. World J. Microbiol. Biotechnol. 13: 433±442. FLEGEL, T.W. (2006). Detection of major penaeid shrimp viruses in Asia, a historical perspective with emphasis on Thailand. Aquaculture 258: 1±33. FLEGEL, T.W. and V. ALDAY-SANZ (1998). The crisis in Asian shrimp aquaculture: current status and future needs. J. Appl. Ichthyol 14: 269±273. FLEGEL, T.W.
FLEGEL, T.W., D.F. FEGAN, S. KONGSOM, S. VUTHIKOMUDOMKIT, S. SRIURAIRTANA, S.
BOONYARATPALIN, C. CHANTANACHOOKIN, J.E. VICKERS, and O.D. MACDONALD (1992). Occurrence, diagnosis and treatment of shrimp diseases in Thailand. In W. Fulks and K.L. Main (eds) Diseases of Cultured Penaeid Shrimp in Asia and the United States. The Oceanic Institute, Honolulu, HI, pp. 57±112. FORESTER, J.R.M. and T.W. BEARD (1974). Experiments to assess the suitability on nine species of prawns for intensive cultivation. Aquaculture 3: 355±368. FULKS, W. and K.L. MAIN (EDS) (1992). Diseases of Cultured Penaeid Shrimp in Asia and the United States. The Oceanic Institute, Honolulu, HI. GAA (GLOBAL AQUACULTURE ALLIANCE) (1999a). Shrimp white spot virus confirmed in Central America. GAA Newsletter, 2, issue 2. GAA (GLOBAL AQUACULTURE ALLIANCE) (1999b). Shrimp white spot disease in Latin America an update. GAA Newsletter, 2, issue 3. HASSON K.W., D.V. LIGHTNER, L.L. MOHNEY, R.M. REDMAN, B.T. POULOS, J. MARI, and J.R. BONAMI (1999). The geographic distribution of taura syndrome virus (TSV) in the Americas: determination by histology and in situ hybridization using TSV-specific cDNA probes. Aquaculture 171: 13±26. HENNIG, O.L., C.R. PANTOJA, S.M. ARCE, B. WHITE-NOBLE, S.M. MOSS, and D.V. LIGHTNER (2005). Development of a specific pathogen free population of the Chinese fleshy prawn Fenneropenaeus chinensis. Part 2: Secondary quarantine. Aquaculture 250: 579±585. HUTCHINS, D.L., G.W. CHAMBERLAIN, and J.C. PARKER (1979). Length±weight relations for several species of penaeid shrimp cultured in ponds near Corpus Christi, Texas. Proc. World Maricult. Soc. 10: 565±570. ICES (1995). ICES Code of Practice on the Introductions and Transfers of Marine Organisms 1994/Code de Conduite du CIEM pour les Introductions et Transferts d'Organismes Marins 1994. ICES Cooperative Research Report No. 204. Copenhagen, Denmark. ICES (2004). ICES Code of Practice on the Introduction and Transfers of Marine Organisms 2004, at www.ices.dk/reports/general/2004/ICESCOP2004.pdf. JIMENEZ, R. (1992). Sindrome de Taura (Resumen). In Acuacultura del Ecuador. Camara Nacional de Acuacultura, Guayaquil, Ecuador, pp. 1±16. JOHNSON, S.K. (1990). Handbook of Shrimp Diseases. Texas A&M University, College Station, TX. JOHNSON, S.K. (1995). Handbook of Shrimp Diseases. Texas A&M University, College Station, TX. KALAGAYAN G., D. GODIN, R. KANNA, G. HAGINO, J. SWEENEY, J. WYBAN, and J. BROCK (1991). IHHN virus as an etiological factor in runt-deformity syndrome of juvenile Penaeus vannamei cultured in Hawaii. J. World Aquaculture Soc. 22: 235±243. LEE, C.S. and P.J. O'BRYEN (EDS) (2003). Biosecurity in Aquaculture Production Systems: Exclusion of Pathogens and Other Undesirables. The World Aquaculture Society, Baton Rouge, LA.
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(1983). Diseases of cultured penaeid shrimp. In J.P. McVey (ed.) CRC Handbook of Mariculture. Vol. 1. Crustacean Aquaculture. CRC Press, Boca Raton, FL, pp. 289±320. LIGHTNER D.V. (1988). Diseases of cultured penaeid shrimp and prawns. In C.J. Sindermann and D.V. Lightner (eds) Disease Diagnosis and Control in North American Marine Aquaculture. Elsevier, Amsterdam, pp. 8±127. LIGHTNER, D.V. (1992). Shrimp viruses: diagnosis, distribution and management. In J. Wyban (ed.) Proceedings of the Special Session on Shrimp Farming, Orlando, Florida, 22±25 May 1992. World Aquaculture Society, Baton Rouge, LA, pp. 238±253. LIGHTNER, D.V. (1993a). Diseases of penaeid shrimp. In J.P. McVey (ed.) CRC Handbook of Mariculture: Crustacean Aquaculture, 2nd edn. CRC Press, Boca Raton, FL, pp. 393±486. LIGHTNER, D.V. (1993b). Non-infectious diseases of Crustacea with an emphasis on cultured penaeid shrimp. In J.A. Couch and J.W. Fournie (eds), Advances in Fisheries Science: Pathobiology of Marine and Estuarine Organisms. CRC Press, Boca Raton, FL, pp. 343±358. LIGHTNER, D.V. (ED.) (1996a). A Handbook of Shrimp Pathology and Diagnostic Procedures for Diseases of Cultured Penaeid Shrimp. World Aquaculture Society, Baton Rouge, LA. LIGHTNER, D.V. (1996b). Epizootiology, distribution and the impact on international trade of two penaeid shrimp viruses in the Americas. Rev. Sci. Tech. Office Int. Epizooties 15: 579±601. LIGHTNER, D.V. (1999). The penaeid shrimp viruses TSV, IHHNV, WSSV, and YHV: current status in the Americas, available diagnostic methods and management strategies. J. Appl. Aquacult. 9: 27±52. LIGHTNER, D.V. (2002). Advances in diagnosis and management of shrimp virus diseases in the Americas. In Y. Inui and E.R. Cruz-Lacierda (eds) Disease Control in Fish and Shrimp Aquaculture in Southeast Asia ± Diagnosis and Husbandry Techniques. SEAFDEC, Aquaculture Department, Iloilo, Philippines, pp. 7±33. LIGHTNER, D.V. (2003a). The penaeid shrimp viral pandemics due to IHHNV, WSSV, TSV and YHV: history in the Americas and current status. In Y. Sakai, J.P. McVey, D. Jang, E. McVey and M. Caesar (eds) Proceedings of the Thirty-second US Japan Symposium on Aquaculture. US±Japan Cooperative Program in Natural Resources (UJNR). US Department of Commerce, NOAA, Silver Spring, MD, and at http:// www.lib.noaa.gov/japan/aquaculture/aquaculture_panel.htm LIGHTNER, D.V. (2003b). Exclusion of specific pathogens for disease prevention in a penaeid shrimp biosecurity program. In C.-S. Lee and P.J. O'Bryen (eds) Biosecurity in Aquaculture Production Systems: Exclusion of Pathogens and Other Undesirables. The World Aquaculture Society, Baton Rouge, LA, pp. 81±116. LIGHTNER, D.V. (2005). Biosecurity in shrimp farming: pathogen exclusion through use of SPF stock and routine surveillance. J. World Aquaculture Soc. 36: 229±248. LIGHTNER D.V. and C.R. PANTOJA (2001). Biosecurity in shrimp farming. In M.C. Haws and C.E. Boyd (eds) Methods for Improving Shrimp Farming in Central America. USDA, Hurricane Mitch Reconstruction Project, UCA Press, Managua, pp. 123± 165 (in English and Spanish). LIGHTNER, D.V. and R.M. REDMAN (1985). A parvo-like virus disease of penaeid shrimp. J. Invertebr. Pathol. 45: 47±53. LIGHTNER, D.V. and R.M. REDMAN (1991). Hosts, geographic range and diagnostic procedures for the penaeid virus diseases of concern to shrimp culturists in the LIGHTNER, D.V.
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17 Selective breeding of penaeid shrimp S. M. Moss and D. R. Moss, Oceanic Institute, USA
Abstract: An estimated 2.67 million metric tonnes of farmed shrimp were produced in 2005 with an estimated value greater than US$10.6 billion. Despite the economic importance of farmed shrimp, the global shrimp farming industry has been slow to adopt approaches to genetic improvement which are prevalent in more mature meat-producing industries. In the early 1990s, a number of research and commercial shrimp breeding programs emerged, and these programs generated basic information about the quantitative genetics of shrimp. Researchers quantified heritability estimates for commercially important traits, and generated information about phenotypic and genetic variation, phenotypic and genetic correlations, and genotype environment interactions. Importantly, these programs provided evidence that selective breeding of shrimp can be effective in improving traits of commercial importance, such as growth and disease resistance. Currently, there are shrimp breeding programs in the Americas, Asia, Australia, New Caledonia, Madagascar, and the Middle East. Although selective breeding offers tremendous opportunity for increased production and profitability to the shrimp farmer by improving commercially important traits, there are significant obstacles to the large-scale adoption of genetic improvement strategies. The value of selectively bred shrimp cannot be fully realized if shrimp are grown in environments where virulent pathogens exist. It is important for farmers using genetically improved shrimp to adopt cost-effective, biosecure strategies to mitigate the risk of pathogen introduction into their grow-out ponds so that the genetic potential of their crop can be fully realized. In addition, care must be taken to ensure that founder stocks for shrimp breeding programs come from genetically diverse populations. If not, problems associated with inbreeding depression may arise, resulting in reduced production and profitability. The sustainability of the global shrimp farming industry will be predicated on the use of selectively bred, genetically diverse populations of specific pathogenfree shrimp stocked in biosecure environments. Key words: penaeid shrimp, selective breeding, genetic improvement, growth, disease resistance, specific pathogen-free.
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Introduction
Shrimp belonging to the family Penaeidae include commercially important species inhabiting tropical and sub-tropical waters around the world (BaileyBrock and Moss, 1992). These shrimp are cultured worldwide and generate significant foreign exchange for the major shrimp-farming countries in Asia and the Americas. According to the Food and Agriculture Organization of the United Nations (FAO), an estimated 2.67 million metric tonnes (MT) of farmed penaeid shrimp were produced in 2005 with an estimated value greater than US$10.6 billion (Fig. 17.1; FAO, 2007). Despite the economic importance of farmed shrimp, the global shrimp farming industry has been slow to adopt genetic improvement strategies that are prevalent in more mature meat-producing industries, such as the poultry and swine industries. This has resulted in production inefficiencies and reduced profits for shrimp farmers. Historically, farmers have relied on the capture of wild shrimp to stock their ponds (see Chapter 16). Wild shrimp are caught as postlarvae or broodstock. Postlarvae are stocked directly into ponds, whereas broodstock are spawned in captivity to produce postlarvae. Wild-caught shrimp pose a serious risk to the industry because they may be carriers of virulent pathogens which can be spread, both horizontally and vertically, throughout a shrimp culture facility or a shrimp farming region. Pandemics caused by White spot syndrome virus (WSSV) and other viruses have resulted in significant economic losses in Asia and the Americas, and these losses can be attributed, in part, to the use of infected, wild-caught shrimp (Chapter 16; Lightner, 2003). Another significant disadvantage of culturing wild-caught shrimp is the inability of the farmer to benefit from selective breeding or other genetic improvement strategies. Selective breeding of terrestrial animals has resulted in
Fig. 17.1 Global production (tonnes) of farmed shrimp from 1950 to 2005 (FAO, 2007).
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dramatic improvements in growth, feed conversion efficiency, and reproductive performance over successive generations. For example, the chickens that we eat today grow twice as fast on half the amount of feed as the chickens of 50 years ago (Boyle, 2001), and this improvement is due, in large part, to the selective breeding practices of poultry breeders. However, benefits accrued from the genetic improvement of shrimp lag far behind those realized in more mature meat-producing industries, despite the fact that many penaeid shrimp possess characteristics amenable to selective breeding, including high fecundity, a relatively short generation time (9±12 months), and ease of captive reproduction. Selective breeding offers tremendous opportunity for increased production and profitability to shrimp farmers by improving commercially important traits. As the economic benefits of selective breeding become more compelling, it is likely that the global shrimp farming industry will invest significant resources in selective breeding programs to produce genetically superior stocks. This chapter introduces some basic concepts of selective breeding and reviews results from research and commercial shrimp breeding programs. The review focuses on commercially important traits such as growth and disease resistance. The chapter concludes with a commentary on future trends and identifies additional sources of information on the genetic improvement of aquatic animals, including penaeid shrimp.
17.2
Selective breeding
17.2.1 Basic concepts Most traits of interest to shrimp farmers are quantitative traits which have phenotypes that exhibit a continuous distribution within a population, such as weight and length. Quantitative traits typically are polygenic and may be controlled by hundreds or thousands of genes (Tave, 1993), although a few genes may have disproportionately large effects on an animal's phenotype with many genes having smaller effects (Orr, 1999). Recently, researchers reported that growth heterosis in oyster larvae was controlled by about 350 genes or 1.5% of the oyster's genome (Hedgecock et al., 2007), and this represents one of the first estimates of the number of genes that determine growth in any animal. Phenotypes of quantitative traits also are influenced by environmental factors. Environmental factors relevant to shrimp breeders (and breeders of other aquaculture species) include stocking density, water temperature, water quality, and feed quality. To improve an animal's performance by selection, it is helpful to understand the phenotypic variance of a population in the context of its component parts. Phenotypic variance (VP ) can be observed and measured and is the sum of three components: genetic variance (VG ), environmental variance (VE ), and the interaction between the genetic and environmental variance (VGE ). VP VG VE VGE
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VG is the variance component of greatest interest to breeders and is the sum of three components: additive genetic variance (VA ), dominance genetic variance (VD ), and epistatic genetic variance (VI ). V G VA V D V I VA is the genetic variance component that is due to the additive effect of all the genes in an organism. Importantly, additive genetic effects are transferred from parents to offspring and it is this variance component that breeders try to exploit through selective breeding. VD is the genetic variance component that is due to the interaction of genes at each locus (i.e. dominance) and is not inherited, although breeders can exploit this variance component by crossbreeding. VI is the genetic variance component that is due to the interaction of genes across loci (i.e. epistasis) and is generally not considered in a genetic improvement program (see Falconer and Mackay, 1996, for a detailed discussion of the different components of phenotypic variance). Because VA is transferred from parents to offspring, this variance component is used to estimate the heritability of quantitative traits. Heritability (h2) describes the percentage of VP that is inherited in a predictable manner (Falconer and Mackay, 1996), and estimates of h2 are critically important when designing a breeding program, predicting a response to selection, or calculating an individual's breeding value (Gjedrem and Olesen, 2005). Heritability in the narrow sense is defined as the ratio between VA and VP : h2 VA =VP Broad-sense heritability is defined as the ratio between VG and VP and includes dominance and epistatic variance components which, as indicated previously, are not passed from parent to offspring. Heritability estimates range from 0 to 1. As a general rule, traits with h2 below 0.2 are considered lowly heritable and may not be amenable to selection (Bourdon, 2000). Traits with h2 between 0.2 and 0.4 are considered moderately heritable, and traits with h2 greater than 0.4 are highly heritable and are very amenable to selection. It is important to note that h2 above 0.7 are rare (Bourdon, 2000). It is also important to note that h2 estimates for a given trait are not immutable. Heritability can vary among populations owing to differences in gene frequencies and environmental conditions, and can change within populations owing to changes in gene frequencies, environmental changes, and age of the animal (Falconer and Mackay, 1996; Gjedrem and Olesen, 2005). Heritability estimates are used by breeders to predict how a population will respond to selection for a given trait, and this information is valuable in assessing the efficacy of different breeding strategies. In a simple, single-trait breeding program, parents with a desirable phenotype are selected from a base population and mated to produce a population of offspring. A breeder can predict the extent to which the trait will improve in the offspring by calculating a response to selection (R). It is important to note that deviations from the
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predicted response result from VD and VI effects, environmental changes (VE or VGE ), or sampling error due to a small sample size (Tave, 1993). The response to selection is the product of h2 for a given trait and the selection differential (S): R Sh2 The selection differential is the difference in the mean phenotypic value (for a given trait) between selected parents and the base population, and essentially describes the phenotypic superiority (or inferiority) of selected individuals relative to the population from which they were selected (Tave, 1993). The selection differential is the product of the selection intensity (i) and the phenotypic standard deviation of the trait of interest (p ): S ip Selection intensity refers to the difference in standard deviation units between selected parents and the base population and is related to the proportion of animals selected as parents. High selection intensities can be achieved in fecund animals such as the penaeid shrimp which can produce more than 150 000 eggs per spawn for the Pacific white shrimp, Penaeus (Litopenaeus) vannamei, and over a million eggs per spawn for the giant tiger prawn, Penaeus monodon (Treece, 2000). The selection differential is also influenced by the standard deviation (or variance) of the trait of interest. Highly variable traits provide a greater opportunity for genetic improvement, whereas less variable traits are more difficult to improve because there is little variation from which to select (i.e. phenotypes are similar). In summary, the goal of any genetic improvement program is to improve the quality or performance of the target species with an expectation that this will improve profitability for the farmer. For quantitative traits, a breeder must understand the different components of phenotypic variance associated with the trait of interest to determine which genetic improvement strategy is most appropriate. If the breeder wants to exploit VA , then selective breeding becomes the strategy of choice. 17.2.2 Selective breeding programs for shrimp Over the past several years, there has been an increasing trend among shrimp farmers to stock their ponds with postlarvae produced from captive broodstock in an effort to minimize the negative impacts of disease (Crocos and Moss, 2006). The disease status of captive broodstock can be controlled, to a significant extent, using specific pathogen-free (SPF) shrimp in conjunction with a comprehensive biosecurity strategy designed to minimize the introduction and spread of pathogens in a maturation/hatchery facility (Chapter 16; Lotz, 1997; Lightner, 2003). SPF shrimp are free of specified pathogens and SPF status is contingent on the level of biosecurity where the shrimp are maintained (Moss et al., 2003a). Currently, only SPF populations of P. vannamei are commercially available on a large scale, and this factor has contributed to P. vannamei
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usurping P. monodon as the most commonly cultured shrimp species in the world. In 2000, an estimated 630 984 MT of farmed P. monodon were produced globally, whereas only 145 387 MT of farmed P. vannamei were produced during the same year (FAO, 2007). However, in 2005, farmed P. vannamei production increased to 1 599 423 MT and this represents a 1000% increase over 5 years. During the same period, farmed production of P. monodon increased to 723 172 MT, representing only a 14.6% increase. Historically, shrimp farmers in the Americas have cultured P. vannamei, so this dramatic species shift has occurred primarily in Asia where more P. vannamei are now produced than in the Western Hemisphere (Moss, 2004). Following the initial development of SPF populations of P. vannamei in the early 1990s (Wyban et al., 1993), a number of research and commercial shrimp breeding programs were established, mostly in the Western Hemisphere. These programs generated basic information about the quantitative genetics of penaeid shrimp, including h2 estimates, estimates of phenotypic and genetic variance, phenotypic and genetic correlations, and genotype environment interactions. Importantly, these programs provided evidence that selective breeding of shrimp can be effective in improving commercially important traits. Currently, there are shrimp breeding programs in the Americas, Asia, Australia, New Caledonia, Madagascar, and the Middle East (Rosenberry, 2006; Clifford and Preston, 2006). The primary traits of interest for shrimp breeders are growth and resistance to viral pathogens (Clifford and Preston, 2006). Shrimp breeders can use several different selection strategies to improve these traits, including individual (mass) and family selection. Individual selection is based on an individual's own phenotype or performance, and individuals are either culled or selected based on their phenotype relative to the population mean. Family selection is based on mean family performance and can be divided into between-family selection and within-family selection. Between-family selection relies on a comparison of family means where entire families are either culled or selected based on mean family performance. Within-family selection is based on the ranking of an individual's performance within each family, and individuals are either culled or selected based on their relationship to their own family mean (see Tave, 1993, and Gjedrem and Thodesen, 2005, for a detailed discussion on different selection strategies). The most appropriate selection strategy depends on a number of factors, including the h2 of the trait under selection and the degree of environmental variance. Individual selection typically is used to improve traits with high h2 , whereas family selection is used to improve traits with low h2 or when there are uncontrollable sources of environmental variance (VE or VGE ). Sib selection, a form of between-family selection, is used when shrimp have to be sacrificed in order to measure the trait of interest (e.g., tail to body weight ratio). This form of selection is particularly useful in SPF breeding programs designed to improve disease resistance (see Section 17.2.4). With this approach, the decision to cull or select a particular family is based on the phenotypic performance of shrimp
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that are exposed to the pathogen of interest in a disease-challenge test. Unexposed, SPF siblings from the best-performing families are then used to propagate the next generation. Individual selection can result in rapid genetic gains in a short period of time for those traits with high h2 . In addition, the cost required to implement a breeding program based on individual selection is relatively low compared with family-based selection programs. However, a significant disadvantage in using individual selection is the potential for inbreeding because the genetic relationships among the different families of a breeding population typically are unknown. Inbreeding results from the mating of individuals who are related by ancestry and can cause a reduction in heterozygosity within a population (Falconer and Mackay, 1996, see Section 17.2.5). As the level of inbreeding accumulates in a population, inbreeding depression can occur resulting in a reduction in mean phenotypic performance of certain traits (typically fitnessrelated traits such as survival and fecundity). Rapid accumulation of inbreeding can be prevented in individual selection programs by keeping a large number of broodstock in the breeding population (~50 pairs of broodstock per generation; Bentsen and Olesen, 2002). Inbreeding can be controlled using family selection because a breeder can avoid mating closely related individuals (e.g. brothers and sisters or first cousins). This can be accomplished by keeping families physically separated or by using markers to differentiate among families, and by maintaining accurate pedigree records. In many shrimp breeding programs, visible implant elastomer (VIE) tags are used as markers to identify families (Godin et al., 1996). These tags are made of a non-toxic, colored elastomer which is injected into the shrimp's connective tissue just under the endocuticle and can be read through the exoskeleton. Molecular markers, such as microsatellites, can also be used to determine parentage (Jerry et al., 2004). Microsatellites are short, non-coding DNA sequences which are repeated many times throughout an organism's genome, and these markers are particularly effective for parentage testing because they are codominant, highly polymorphic, and inherited in a Mendelian fashion (Argue and Alcivar-Warren, 2001). Because the use of molecular markers precludes the need for separate rearing areas or the use of physical tags, high selection intensities can be attained, resulting in more rapid genetic gains. However, the use of molecular markers requires the application of technologies typically unavailable to most shrimp farmers (e.g., polymerase chain reaction), and may be cost prohibitive. Currently, there are a number of shrimp breeding programs that use family selection and physical tags to improve growth and disease resistance. More sophisticated breeding programs are beginning to emerge which rely on the computation of estimated breeding values (EBVs) using computer software for the integration and analysis of pedigree and performance data (e.g., best linear unbiased prediction (BLUP) analysis, Castillo-JuaÂrez et al., 2006, and see Gjerde, 2005, for more information about BLUP analysis and other methods used to estimate breeding values). As indicated previously, only SPF populations of P. vannamei are commercially available on a large scale, so it has been
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the species of choice for most commercial breeding programs, particularly in the Western Hemisphere. There are ongoing breeding programs for other penaeid species, including Penaeus (Fenneropenaeus) chinensis (Wang et al., 2006), Penaeus (Litopenaeus) stylirostris (Goyard et al., 1999), Penaeus (Marsupenaeus) japonicus (Preston et al., 2004), and P. monodon (Coman and Preston, 2008), although most of these programs are research-oriented or are at early stages of commercial development. 17.2.3 Selection for growth A major goal of many selective breeding programs is to improve growth of the target species. In the context of shrimp farming, faster growth will either increase the number of crops per year, thereby increasing annual yield (kg/ha/ year), or increase the weight of shrimp at harvest, resulting in higher prices per kg for the farmer. Selecting for faster growth also may improve other commercially important traits by indirect selection, such as feed conversion efficiency (Goyard et al., 2001) and pond survival (Gitterle et al., 2005a). Pond survival may improve because shrimp are more susceptible to certain pathogens at a smaller size (Brock et al., 1997, but see Lotz, 1997). Growth can be expressed in terms of absolute, relative, or specific growth rate (Hopkins, 1992). In the context of shrimp farming, absolute growth rate typically is used and is expressed as weight gain over a defined time period (e.g., g/day, g/ week). Shrimp breeders often express growth as `weight at age' or harvest weight, and this measure is meaningful when shrimp are stocked into a test environment at the same time and at a similar age or size. Not surprisingly, there is a high phenotypic and genotypic correlation (>0.85) between harvest weight and growth rate (Dr John Rocha, Aquatec, pers. comm.), and either trait can be used as a selection criterion in a breeding program. In addition to measuring shrimp weight, breeders can use morphometric correlates of weight in a selection program if there is a high positive correlation between the traits of interest. For example, the correlation coefficient between shrimp tail weight and depth of the sixth abdominal segment in P. stylirostris and P. vannamei was 0.95 and 0.85, respectively (Lester, 1983), and it was suggested that this morphometric measurement could serve as a suitable selection criterion to improve tail weight in a breeding program. Potential advantages in using morphometric correlates of weight include ease and accuracy of data acquisition (Lutz, 2001). Because shrimp do not exhibit linear growth throughout their life, the age or size at which growth is measured becomes an important consideration for a shrimp breeder. If the goal of a breeding program is to increase shrimp weight after a certain number of days of culture, shrimp should be selected at that time unless there is a high positive correlation between shrimp weight at harvest and at other time intervals. For example, there was a high positive genotypic correlation in weight of P. monodon after 30 and 40 weeks of culture (rG 0:97; Kenway et al., 2006). However, the strength of this correlation decreased as the time interval increased from 30 weeks to 54 weeks of culture (rG 0:73). In a
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study with P. vannamei, the phenotypic correlation between market weight (~20 g) and broodstock weight (>35 g) was moderate (rP 0:42; Argue et al., 2000). If the goal of a breeding program is to select the fastest-growing shrimp to market weight, shrimp should be selected at market weight rather than as broodstock. Many penaeid species exhibit sexual growth dimorphism where females grow faster than males (Diaz et al., 2001; Hennig et al., 2003; Argue et al., 2002; Otoshi et al., 2003; Kenway et al., 2006). Sexual dimorphism for growth typically occurs when shrimp are 13±18 g for P. vannamei (Chow and Sandifer, 1991; Moss et al., 2002; PeÂrez-Rostro and Ibarra, 2003a) and 13±28 g for P. monodon (Cheng and Chen, 1990; Hansford, 1991; Kenway et al., 2006). Although the cause of sexual growth dimorphism in penaeid shrimp is unclear, it may involve physiological differences between the sexes rather than behavioral ones (Hansford and Hewitt, 1994; Moss and Moss, 2006). Shrimp breeders can account for sexual growth dimorphism in data analyses by treating sex as a fixed effect to correct for mean and variance differences in growth- and size-related traits (PeÂrez-Rostro and Ibarra, 2003a; Kenway et al., 2006). Because female shrimp grow faster than males at a certain size/age, the use of an all-female population could increase farm production by 5±15% per crop (Argue and Alcivar-Warren, 2001), and use of a mono-sex population could provide seedstock and broodstock suppliers with a mechanism to protect their valuable germplasm. Despite these benefits, little research has been done on sex reversal in shrimp and the genetic mechanism controlling sex in penaeid shrimp is not well understood. In an effort to explore the possibility of skewing the sex ratio of P. vannamei through selective breeding, Argue et al. (2002) reported that the heritability estimate for sex ratio was not significantly different from zero. These results suggest that it is not possible to produce a higher percentage of females through selective breeding. However, sex ratio may be altered by manipulating the androgenic gland or exposing shrimp to exogenous hormones (Nagamine et al., 1980; Sagi and Cohen, 1990; Malecha et al., 1992; but see Moss et al., 2003b). Heritability estimates for growth- and size-related traits (i.e. weight, length, growth rate, etc.) have been reported in a number of penaeid species (Table 17.1) including P. vannamei (Carr et al., 1997; Argue et al., 2002; PeÂrez-Rostro and Ibarra, 2003a,b, Gitterle et al., 2005a), P. stylirostris (Goyard et al., 2002), P. monodon (Benzie et al., 1997; Jarayabhand et al., 1998; Kenway et al., 2006), and P. japonicus (Hetzel et al., 2000). In general, h2 estimates for growth are considered moderate to high (h2 0:2) for penaeid shrimp and they typically are associated with large standard errors. Traits with moderate to high h2 estimates should respond well to selection because a significant portion of the phenotypic variance (VP ) is inherited (see Section 17.2.1). There are a number of published studies that indicate significant improvement in shrimp growth can be made through selective breeding. In a study with P. vannamei, Argue et al. (2002) reported that selected shrimp were 21% and 23% larger at harvest than unselected control shrimp after one generation of selection when reared in a
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Table 17.1 Heritability estimates (h2 SE) for growth- and size-related traits in commercially important penaeid shrimp Species
Trait
h2 SE
Reference
P. vannamei P. vannamei
Weight at ~11 g Weight at ~23 g
Carr et al. (1997) Argue et al. (2002)
P. vannamei
Weight at 29 wk
0.42 0.84 1.19 0.34
P. vannamei
Total length at 29 wk Weight at ~20 g
P. vannamei P. stylirostris P. monodon
Growth rate from ~5 to 17 g Weight at 57 mg Weight at 449 mg
P. monodon
P. monodon P. japonicus
Total length at 25 days old Total length at 65 days old Weight at 30, 40, and 54 wk Weight at ~6 months
0.15 0.43 (raceway) 0.59 (pond) 0.18
0.28 0.18 0.24 0.05 (line 1) 0.17 0.04 (line 2) 0.11 0.12 0.56 0.10 0.39 0.15
0.02 0.03 0.00 0.00 0.06
0.07 0.04
(sire), (dam) (sire), (dam)
PeÂrez-Rostro and Ibarra (2003a) PeÂrez-Rostro and Ibarra (2003a) Gitterle et al. (2005a) Goyard et al. (2002) Benzie et al. (1997)
Jarayabhand et al. (1998)
0.55 0.07, 0.45 0.11, Kenway et al. (2006) 0.53 0.14 0.23 Hetzel et al. (2000)
raceway and pond, respectively. Kenway et al. (2006) concluded that selecting the largest 40% of P. monodon at 30 weeks should increase 30-wk weight by 10%. Hetzel et al. (2000) reported a mean selection response for growth of 10.7% after one generation of selection in P. japonicus, despite a low selection intensity of 29%. Goyard et al. (2002) indicated that, after five generations of mass selection, P. stylirostris exhibited a 21% increase in growth rate compared to unselected control shrimp. Clearly, significant improvement in growth can be achieved if growth is the only trait being selected. However, if shrimp breeders want to improve two or more traits simultaneously, information about correlated responses becomes important. If there is a significant negative correlation between growth and other trait(s) under selection, the rate and magnitude of genetic improvement for growth will be compromised. Information about phenotypic and genotypic correlations for growth and other commercially important traits in penaeid shrimp are limited (Table 17.2). Argue et al. (2002) found a negative genotypic correlation ( SE) between mean family harvest weight and mean family survival to Taura syndrome virus (TSV) exposure in P. vannamei (rG ÿ0:46 0:18). Similarly, Moss et al. (2005) reported a significant negative phenotypic correlation between mean family weight and mean family survival to TSV (rP ÿ0:15), based on performance data from 587 full-sib
Table 17.2
Phenotypic (rP ) and genotypic (rG ) correlations ( SE) between commercially important traits in penaeid shrimp
Species
Traits
P. P. P. P.
Weight Weight Weight Weight
vannamei vannamei vannamei vannamei
and and and and
grow-out survival survival to TSV exposure survival to TSV exposure growout survival
P. vannamei
Weight and survival to WSSV exposure
P. monodon P. monodon
Weight at 16 wk and survival at 16±35 wk Weight at 30 wk and survival at 16±35 wk
rP SE
rG SE
ÿ0.18 ÿ0.08 0.07 ÿ0.15
ÿ0.46 0.18 0.42 0.40 ÿ0.55 ÿ0.64 ÿ0.05 0.12
(line 1) (line 2) 0.18 (line 1) 0.19 (line 2) 0.18 0.22
Reference Carr et al. (1997) Argue et al. (2002) Moss et al. (2005) Gitterle et al. (2005a) Gitterle et al. (2005b) Kenway et al. (2006) Kenway et al. (2006)
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families of P. vannamei. Gitterle et al. (2005b) reported a negative genotypic correlation between harvest weight and survival to WSSV exposure in two different lines of P. vannamei (rG ÿ0:55 0:18 and ÿ0:64 0:19). These results suggest that there may be a trade off between growth and disease resistance in shrimp (see further discussion in Section 17.2.4). Conversely, Gitterle et al. (2005a) found a positive correlation between the estimated mean full-sib family breeding value for harvest weight and pond/tank survival in two different lines of P. vannamei (rG 0:42 and 0.40). These data suggest that selecting for growth will result in improvements in growout survival, provided viral pathogens such as TSV or WSSV are not present. For P. monodon, Kenway et al. (2006) reported that genetic correlations between family survival and weight at age were generally low with large standard errors. Growth (and other commercially important traits) may be affected by the interaction between an organism's genotype and its environment (G E interaction). If these interactions are significant, breeders may need to develop different lines of shrimp for each unique rearing environment. Significant G E interactions tend to occur when differences between or among genotypes are large, or when there are large differences between or among test environments (Falconer and Mackay, 1996). Little published data are available on G E interactions for growth in shrimp, although Gitterle et al. (2005a) reported a low genotype test environment interaction for harvest weight in P. vannamei reared in ponds and tanks. Similarly, PeÂrez-Rostro and Ibarra (2003b) reported an insignificant G E interaction for harvest size in P. vannamei, although the lack of significance may have resulted from little genotypic differences among the families under evaluation or a limited amount of time that the shrimp spent in the different test environments. In contrast, Coman et al. (2004) reported a small but significant genotype density interaction for growth in P. japonicus, although only six families were evaluated. These authors suggested that selection of genotypes for rapid growth at low density may not produce genotypes with rapid growth at high density. In summary, significant improvement in shrimp growth can be made through selection. Harvest weight is a good candidate trait to select for because it is easy to measure, is highly correlated with growth rate, and is economically important. In addition, harvest weight may be positively correlated with other commercially important traits such as grow-out survival. Although published data on G E interactions for growth-related traits are limited, existing data suggest that shrimp which grow well in one environment will also grow well in other environments. The lack of a significant G E interaction precludes the need to develop multiple fast-growing shrimp lines for different rearing conditions. However, there may be a significant G E interaction for grow-out survival (Dr John Rocha, Aquatec, pers. comm.) and additional research is needed to explore this relationship. Further research also is needed to explore the relationship between growth and disease resistance, although there are breeding strategies available to improve both traits simultaneously even if they are negatively correlated.
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17.2.4 Selection for disease resistance Shrimp farmers have suffered significant economic losses over the past 15 years due to various diseases (Lightner, 2003; Flegel, 2006) and these losses have catalyzed fundamental changes in the way shrimp aquaculture is practiced (Chapter 16; Moss et al., 2001). Biosecurity protocols are now common among industry stakeholders (Boyd, 2006) and there are continuing efforts to develop disease-resistant shrimp (Clifford and Preston, 2006). To date, shrimp breeders have focused most of their efforts on developing families of shrimp with enhanced resistance to TSV and WSSV (Argue et al., 2002; Kong et al., 2003, Jiang et al., 2004; Gitterle et al., 2005b) because these two viruses have had the greatest economic impact on the shrimp farming industry (Lightner, 2003). A noteworthy exception was reported by Tang et al. (2000) who provided unequivocal evidence that a line of P. stylirostris was selectively bred for complete resistance to Infectious hypodermal and hematopoietic necrosis virus (IHHNV). Their results indicated that IHHNV did not replicate in postlarval or juvenile P. stylirostris, although the genetic basis for IHHNV resistance was not reported. (Note that in this chapter, we use the words `resistant' and `resistance' as general terms to refer to a shrimp's ability to survive a viral infection. These terms have been adopted by many stakeholders in the shrimp farming industry, although the terms `tolerant' and `tolerance' may be more appropriate.) Selective breeding for TSV resistance began in the mid-1990s in response to a TSV epizootic that devastated populations of P. vannamei in Ecuador and the subsequent spread of TSV throughout the Americas. TSV is a single-stranded RNA virus belonging to the family Dicistroviridae (Bonami et al., 1997) and can infect juvenile shrimp within 2±4 weeks after stocking into nurseries or growout ponds. Cumulative mortalities of unselected shrimp in TSV-infected ponds were reported to be as high as 80±90% (Brock et al., 1997; Lightner et al., 1998). Revenue losses from TSV in 1993 were estimated to be US$400 million in Ecuador alone (Lightner, 1999), and this virus has since spread to and impacted major shrimp farming regions in Asia (Tu et al., 1999; Phalitakul et al., 2006). Breeding programs designed to enhance TSV resistance have generated valuable information about the quantitative genetics of disease resistance in shrimp and have highlighted some challenges associated with trying to improve resistance through selection. Unlike growth, h2 estimates for TSV resistance are considered low to moderate (h2 0:2). Argue et al. (1999) reported h2 estimates for TSV resistance in successive generations of P. vannamei with estimates ranging from ÿ0:04 0:01 (SE) to 0:31 0:07. Fjalestad et al. (1997) reported a mean h2 estimate of 0:22 0:09 for TSV resistance in P. vannamei, and Argue et al. (2002) reported a half-sib h2 estimate of 0:19 0:08 and a realized h2 estimate of 0:28 0:14 for one generation of selection in the same species. Despite low to moderate h2 for TSV resistance, significant improvements in this trait have been made through selection. Argue et al. (2002) reported an 18.4% increase in TSV survival after one generation of selection in a population of P. vannamei, compared with an unselected control population. Fjalestad et al.
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(1997) reported a selection response of 12.4% (expressed as the relative increase in TSV survival per generation) for the same species. White et al. (2002) reported an absolute increase in mean TSV survival from 24% to 37% among selected P. vannamei families over a 3-year period. In addition, there was an increase in survival from 65% to 100% among the best-performing families during the same time period. Gitterle (1999) reported that, after an initial TSV outbreak in Colombia, pond survival typically was about 45%. However, survival returned to pre-TSV levels of about 80% after just three generations of intense mass selection (selection of survivors from infected ponds). The ability to improve TSV resistance by selection (despite low to moderate h2 ) is attributed, in part, to high phenotypic/genotypic variation in TSV survival. This variation allows for a larger selection differential (and higher selection intensity) which increases the selection response (see Section 17.2.1). Argue et al. (2002) reported that TSV survival ranged from 15% to 94% among 80 P. vannamei families exposed to TSV in a per os laboratory-challenge test. Similarly, White et al. (2002) reported that TSV survival ranged from 0% to 100% among 176 families. Although large variations in TSV survival have been observed among populations of P. vannamei families, the magnitude of this variation can decline as selection progresses. For example, while mean family survival increased from 44% to 84% after five generations of selection for TSV resistance among a population of P. vannamei families at the Oceanic Institute (OI, Waimanalo, Hawaii), the coefficient of variation (CV) for TSV survival decreased from 43.3% and 13.6% (Fig. 17.2; unpublished data). This reduction in variability is expected as selection progresses and will result in progressively lower selection responses (Falconer and Mackay, 1996). In addition to developing lines of shrimp with enhanced resistance to TSV, shrimp breeders have explored the possibility of selecting shrimp for WSSV resistance. These efforts were in response to the introduction and spread of this virus throughout the Americas in the mid to late 1990s (Rosenberry, 1999, 2000). WSSV initially was identified in Taiwan in 1992 (Chou et al., 1995) and spread rapidly throughout Asia (Inouye et al., 1994; Wongteerasupaya et al., 1995; Flegel and Alday-Sanz, 1998). WSSV first appeared in the US in 1995 and was identified in many shrimp farming regions of the Americas by 1999 (Lightner, 1996, 1999; Jory and Dixon, 1999). WSSV is a double-stranded DNA virus belonging to the family Nimaviridae (Escobedo-Bonilla et al., 2008) and cumulative mortalities of shrimp in WSSV-infected ponds were reported to exceed 90% (Lightner, 1999; Gitterle et al., 2005b). Heritability estimates for WSSV resistance typically are lower than those reported for TSV. Published h2 estimates for WSSV resistance in P. vannamei range from 0.00 to 0.21, and most estimates are 100/g. The mechanism for texture toughening is not well understood; however, denaturation and aggregation of protein have been suggested as a possible explanation. This may be due to mechanical damage associated with the rate of freezing such as ice crystal formation and ion concentration effects, or interaction with chemical components such as formaldehyde or lipid oxidation products. 21.2.3 Processing crawfish The major crawfish species processed is red (Procambarus clarkii) with some white (P. acutus acutus) production. The season is limited to the spring since the carapace becomes increasingly hard as summer progresses and the animals bury in the mud during winter months. Crawfish are obtained from two sources, swamp or delta areas and rice fields. The animals are brought to the processing facility where they are traditionally cooked by boiling. The procedure is similar to that previously described for blue crabs. After cooking, the crawfish are placed under refrigeration for hand processing. Approximately 2.7±4.5 kg (6± 10 lb) of tail meat per hour can be produced by one worker. Only one product, tail meat, is produced with a yield of approximately 14±15%. Meat for sale is usually packed with adhering hepatopancreas, an important flavor ingredient in many prepared dishes. However, meat packed this way has a short shelf-life, usually less than 7 days, since the presence of hepatopancreas can cause textural problems, resulting in mushiness. The loss in texture is caused by collagenolytic and proteolytic enzymes that have been shown to degrade tail meat and collagen gels (Nip et al., 1985). Marshall et al. (1987) reported that after 20 hours of iced storage, crawfish cooked less than 7 minutes were significantly softer than crawfish cooked for 7 to 13 minutes. The longer cook time was shown to inactivate certain enzymes involved in the mushiness in crawfish meat. When the tail meat was washed and the hepatopancreas removed, the meat texture was acceptable for up to 25 days of iced storage; however, flavor was found to be slightly objectionable after 15 days (Flick et al., 1994). This compares favorably with Cox and Lovell (1973) and Gerdes et al. (1989) who found that tail meat was acceptable up to 23 days and 21 days in ice-pack storage respectively. Toughness increased significantly (P 0:05) as a result of freezing at ÿ23 ëC in a commercial still freezer. However, toughness decreased after 16 weeks of frozen storage (Godber et al., 1989). Individually quick frozen (IQF) samples were more tender than conventionally frozen samples.
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Predominate microorganisms in stored tail meat were identified by Cox and Lovell (1973) as Achromobacter, Alcaligenes, Flavobacterium, Micrococcus, Pseudomonas, and Staphylococcus. The greatest number of rapid spoilers belonged to the genus Pseudomonas with Achromobacter a distant second. Lyon and Reddmann (2000) evaluated the potential survival and outgrowth of biological hazards in both vacuum-packaged and air-permeable-packaged cooked crawfish meat stored at 4 and 10 ëC for 30 days. A total of 31 bacterial species were isolated and identified. The microorganisms were identified as belonging to the following genera: Acinetobacter, Aeromonas, Flavobacterium, Cornebacterium, Enterobacter, Enterococcus, Escherichia, Pantoea, Peptostreptococcus, Serratia, Staphylococcus, and Streptococcus. Crawfish meat was inoculated with 103 Clostridium botulinum type E spores per gram of tail meat to determine whether growth and toxin production would occur during refrigerated storage. The spore-inoculated tails were vacuum packaged in both a high-barrier film and an air-permeable bag and stored at 4 and 10 ëC for 30 days. C. botulinum toxin E was not detected in any of the containers throughout the shelf-study until day 30, by which time spoilage had occurred. The California crawfish (Pacifastacus leniusculus) is exported to Sweden since it closely resembles the species Astacus astacus, native to northern Europe. The whole live crawfish are cooked, packaged in beer brine flavored with dill spices, vacuum sealed, pasteurized, and frozen (Dehlendorf, 1981). The effects of elevated CO2 levels on the keeping quality of cooked, freshwater California crawfish was investigated by Wang and Brown (1983). An enriched atmosphere of 80% CO2: 20% air was compared with air storage at 4 ëC. Chemical and microbial changes were correlated with sensory panel evaluations of flavor, odor, and texture. After 28 days of storage, the concentrations of ammonia, trimethylamine, and total plate counts were lower in crawfish stored under enriched carbon dioxide as compared to samples stored in air. Samples stored under enriched carbon dioxide for 21 days were not significantly different from fresh cooked crawfish. Samples stored in air were found to have significantly more fishy flavor and odor after 14 days of storage. The Australian red claw crawfish (Cherax quadricarinatus) is a robust freshwater crustacean species native to the rivers and streams of northern Australia. They are commonly known as red claw because males show red marks on the outer margin of their claws. Compared with native American crawfish, red claw have several advantages, including larger potential size, high percentage of dress-out (meat), and better tolerance of crowded culture conditions (Masser and Rouse, 1997). Tail meat is approximately 22% of the total weight (Jones, 1989). The processing is similar to that of other crawfish species. Prior studies have indicated that red claw muscle is susceptible to protein denaturation and lipid oxidation at refrigerated temperatures and upon repeated freezing and thawing. Repeated freezing and thawing of the tail meat resulted in increased thiobarbituric acid reactive substances (TBARS) from 0.070 mg/kg (fresh) to 1.201 mg/kg (six freezing cycles). Concomitant with the TBARS increase was a major loss in cooking yield as well as tenderness of
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meat. To retain good qualities of the tail meat, a maximum of three freeze/thaw cycles should be employed. Red claw tail meat was dipped in water (control), tocopherols, and propyl gallate 0.06% (w/w) and stored for 1, 3, and 6 months at ÿ20 ëC (Tseng et al., 2005). The meat treated with the antioxidants showed a lower (P < 0:05) TBARS production than the control. The antioxidants did not prevent texture softening in the muscle meat during frozen storage.
21.3
Packaging and preservation
21.3.1 Pasteurization Pasteurization processes for many crustacean species have received great attention. Shelf-lives in excess of 6 months have been obtained under refrigeration conditions (Lynt et al., 1977; Ward et al., 1984). Flick and Wall Bourne (2005) reported that a can of pasteurized blue crab meat was microbiologically acceptable after 40 years of refrigerated storage; however, a significant color change occurred in the meat. This extended shelf-life permits producers to improve their distribution and maintain their supply over the full year and also has an advantage over frozen product that customarily must be thawed prior to use. The traditional pasteurization process reported by Virginia Tech is: F185 31, Z 16. The bacteria of major concern in pasteurization are pathogens such as Listeria monocytogenes, Clostridium botulinum E, non-proteolytic Clostridium botulinum B, and non-proteolytic Clostridium botulinum F. These bacteria have the potential for growth under refrigerated conditions, 3 ëC, and are not destroyed by traditional cooking thermal processes. Ward et al. (1977) compared the flora of unpasteurized and pasteurized crab meat. They found the surviving bacteria from the process were largely lactic acid-producing, nonspore forming Gram-positive rods. They determined the decimal reduction time of these lactobacilli to be 2.5 min at 55 ëC in a phosphate buffer. Solomon et al. (1982) reported that non-proteolytic Clostridium botulinum types B and F and proteolytic type B do not grow in blue crab meat at reduced temperatures after thermal processing. Lynt et al. (1977) reported that the decimal reduction times for Clostridium botulinum type E spores were 0.74 min in Beluga blue crab and 0.51 min in Alaska blue crab, at a reference temperature of 82.2 ëC. The greatest z value for the Beluga was 8.33 ëC. Scott and Bernard (1982) compared the heat resistance of spores of non-proteolytic Clostridium botulinum type B with type E and proteolytic type B spores. Non-proteolytic type B strains had a greater thermal resistance than the proteolytic strains. Dungeness crab meat was inoculated with an equal mixture of three strains of Clostridium botulinum type B for a total of 107 spores, for calculating their thermal resistance (a 7 D process) (Peterson et al., 1997). After processing, the crab meat was transferred to enrichment medium where it was incubated anaerobically for 150 days. Process times ranged from 90 min at 88.9 ëC to 20.3 min at 94.4 ëC. D-values ranged from 12.9 for the 88.9 ëC process to 2.9 for the
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Table 21.2
Cook times for lump crab meat pasteurized at 83.3 ëC
Container type Steel can Plastic can Aluminum can Barrier pouch Non-barrier pouch 1
Contents (g)
Cook time (min)
F-value1 (min)
453.6 226.8 226.8 226.8 226.8
163 130 120 70 70
53.8 43.8 39.7 42.8 45.2
F85, Z 9:9
94.4 ëC process. The relative sterilization value Fo was 0.054 and the pasteurization value, F185 Z 16 was 240. The pasteurization process safely extended refrigerated shelf-life by inactivating spores of Clostridium botulinum non-proteolytic types B, E, and F and also non-sporeforming pathogens such as Listeria monocytogenes. The process did not inactivate the heat-resistant strains of C. botulinum or other more heat-resistant sporeformers. The sensory changes in pasteurized blue crab meat packaged in steel containers, aluminum cans, plastic cans, nonbarrier pouches, and barrier pouches and stored at 0 and 4 ëC (Table 21.2) was investigated by Gates et al. (1993). Meat pasteurized in plastic and aluminum cans had better sensory and microbiological quality and shelf-life than meat packed in steel cans. Oxygenbarrier pouches had the lowest quality and shortest shelf-life. Non-barrier pouches had product with quality similar to meat in steel cans, but with an extended shelf-life. No packaging materials improved the microbiological shelflife of freshly cooked meat. Vacuum skin packaging resulted in improved sensory qualities. A study on the discoloration in thermally processed crab meat was performed by Requena et al. (1999). The meat became darker with increased heating process; crab harvest location had a significant effect on the lightness of the meat; and meat located in the bottom of the can was darker than the top. Rock crab meat was placed in a polypropylene container with an aluminum pull-top lid containing 2% (w/w) brine (Ghazala and Trenholm, 1996). The slowest heating spot for a conductively heated product was found to be 3.0 cm below the lid. Pasteurization was evaluated at 81, 82, and 83 ëC to a pasteurization lethality of 40 min at 85 ëC (Z 8:89 ëC). The optimal process was found to be 83 ëC for 130 min. The shear strength for unprocessed crab leg samples was, on average, 0.33 N/mm2 compared with 0.24 N/mm2 for processed samples. This represents a 24% reduction in shear stress that is largely attributed to the processing method. 21.3.2 Irradiation While crustaceans are thermally processed to facilitate picking, microbial contamination occurs during picking and packaging. Low-dose gamma irradiation
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551
and electron beam irradiation has proven effective in reducing the introduction of pathogenic and spoilage microorganisms during processing in a variety of seafood products. The microbial and sensory quality of gamma-irradiated (cobalt-60, < 2 kGy) crab products (white lump, claw, and fingers) through 14 days of ice-storage were shown to be acceptable by Chen et al. (1996). Irradiation effectively reduced spoilage bacteria extending shelf-life by more than 3 days beyond control samples during iced storage. During storage, fresh crab odor and flavor were initially similar for treated and control samples, while off-flavors and odors developed more rapidly in controls. Overall acceptability scores for irradiated crab samples were higher than for control samples throughout 14 days of storage. 21.3.3 Frozen The sensory properties of cryogenically frozen blue crab meat, treated with polydextrose, a blend of sucrose/sorbitol/phosphate, or water were compared with pasteurized meat using both trained and consumer panels (Henry et al., 1995). The samples were vacuum sealed and stored for 32 weeks at ÿ29 ëC and evaluated for aroma, flavor, and texture. Polydextrose and sucrose/sorbitol/ phosphate treatments were closest in sensory attributes to fresh crab meat in that they had fewer adverse changes in quality than the water, reference, or pasteurized treatments. Panelists detected more sour, rancid, and ammonia notes and less crab flavor in the pasteurized sample. All the cryogenically frozen samples were closer to fresh crab meat, in terms of quality attributes than pasteurized meat. Whole blue crabs were frozen while alive in polyethylene polyamide (PE/ PA) pouches, shrunk, and stored for 10 months at ÿ18 ëC (Yerlikaya and Gokoglu, 2004). The crabs were withdrawn from frozen storage and sampled monthly. Changes that occurred were investigated by means of sensory assessments (odor, appearance), chemical analyses total volatile bases (TVBN), trimethylamine (TMA-N), and physical measurements (pH). TVN-B, TMAN values, and pH were significantly different (P 0:05) following frozen storage. TMA-N increased from 0.2 mg/100 g to 25 mg/100 g while TVB-N increased from 1 mg/100 g to 25 mg/100 g. Odor scores were significantly different (P 0:05) (ranging from 9 to 7 over the storage period in a 10-point evaluation score) whereas the appearance scores were not (ranged from 9 to 8). There was only a weak or negative correlation between physical and chemical analyses and sensory attributes. 21.3.4 High hydrostatic pressure processing The effect of high-pressure and high-temperature treatments at various process times on the inactivation of spores of Clostridium botulinum non-proteolytic type B strains, 2-B, 17-B, KAP8-B, KAP9-B, suspended in phosphate buffer and a pasteurized crabmeat blend was investigated by Reddy et al. (2006). Spores of KAP8-B were less resistant to high-pressure treatment than spores of 2-B, 17-B,
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Shellfish safety and quality
and KAP9-B in both phosphate buffer and crabmeat blend. No survivors of initial counts (> 4.3 log units) of KAP8-B spores were detected after processing at 827 MPa and 60 ëC for 10 min. The level of inactivation for 2-B, 17-B, and KAP9-B spores in phosphate buffer and crabmeat blend increased with the increase in processing time from 10 to 30 min at 827 MPa at 75 ëC, a reduction of > 6 log units of 2-B, 17-B, and KAP9-B spores in phosphate buffer and crabmeat blend was observed for processing times between 20 and 30 min. The crabmeat blend as a suspension menstrum provided no protection against inactivation of spores of 2-B, 17-B, and KAP9-B by high-pressure processing. High-temperature (> 95 ëC) and lower-pressure (620 MPa) treatments for up to 10 min were also found to inactivate 17-B spores in phosphate buffer.
21.4
Contaminants
Residues or unwanted contaminants in crustaceans can be loosely classified into four groups: (1) antibiotics and other drug residues, (2) pesticides, (3) industrially generated persistent organic pollutants (POPs), such as dioxins, and (4) heavy metals. Antibiotics and pesticides are more likely to be of concern in aquaculture products than wild-caught products, especially in aquaculture products produced in countries with less restrictive drug and pesticide use policies than those found in the US. Heavy metals and POPs tend to bioaccumulate in fatty tissues. These contaminants have been observed in wildcaught crustaceans but, again, may be more prevalent in farm-raised products because the high-protein feed used in aquaculture is made from fish oils and fish meal, which can concentrate toxic contaminants. For wild-caught product, environmental chemical contaminants and pesticides are most likely to be of concern for harvests from fresh water, estuaries, and near-shore coastal waters (areas subject to shore-side contaminant discharges) as opposed to harvests from open ocean. Since 1997, the FDA has required US processors of fish and fishery products to develop and implement hazard analysis critical control point (HACCP) systems for their operations. The Fish and Fisheries Products Hazards and Controls Guidance manual (Department of Health and Human Services et al., 2001) includes a table of potential health-related safety hazards listed by species. While the document indicates that the list of hazards may not be exhaustive and that new and emerging hazards may not be included in the table, it is a reasonable starting point for identifying potential areas of concern. Based on this reference, chemical residues (pesticides, POPs, and heavy metals) are potentially of concern for blue crab, Dungeness crab, swimming crab, crawfish (both aquacultured and wild caught), and aquacultured shrimp (produced in either fresh or saltwater). Drug residues are listed as a potential concern for all aquacultured crustacean products. Antibiotics are used to decrease incidences of disease in aquaculture. Unfortunately, residue of antibiotics can remain in the product. Concerns for
Slaughter, storage, transport, and packaging of crustaceans
553
antibiotic use in aquaculture include both the potential to develop antibioticresistant bacteria, and possible toxic effects on humans consuming the product. Two notable examples of antibiotics that have been banned for use with foodproducing animals and fish by both the US and the EU are chloramphenicol and nitrofurans (Public Citizen's Food Program, 2004). Shakila et al. (2006) report that chloramphenicol residues in shrimp declined when the shrimp was subjected to either retorting (at 121 ëC) or cooking (100 ëC). Retorting for 10 and 15 min resulted in loss of 9% and 16% respectively while cooking for 10, 20, and 30 min resulted in loss of 6%, 12%, and 29% respectively. Numerous papers have been published on specific chemical residues observed in fish harvested from specific waters (Hauge et al., 1994; Jop and Hoberg, 1995; Guhathakurta et al., 2000; Burgos-Hernandez et al., 2005, 2006; Fabris et al., 2006). It is impossible to list all areas of concern for hemical residues. In general, warnings are published or restrictions are placed regarding consumption of fish harvested from areas of known contamination. 21.4.1 Organic The biological assimilation of polychlorinated biphenyls (PCBs) in crabs has been reported by several researchers. The levels of PCBs in California crabs was 0.19 ppm (Young, 1982); 0.09±0.11 ppm in rock crabs harvested off the northern New Jersey coast (Reid et al., 1980); 0.03±0.38 ppm in sand crabs caught in California (Wenner, 1986); and crabs harvested off the New Jersey coast contained mean PCB levels of 0.33 ppm (Hauge et al., 1990). Research published by Zabik et al. (1992) showed that both steaming and boiling blue crabs reduced PCB concentrations in both body muscle and claw. The authors recommended that the medium used for cooking blue crabs should be discarded and not used for preparing other foods. Although cooking does reduce the residues, this advantage is lost if the cooking broth is further used to prepare soups, sauces, flavorings, and other food products. The actual average parts per million level of the cooking medium was found to range from 0.031 to 0.039 depending on the preparation of the crabs prior to cooking. Zabik et al. (1992) quantified changes of distribution of PCBs in blue crab caused by boiling and steaming. They concluded that all cooking procedures reduced PCBs by greater than 20%. Removing the hepatopancreas increased PCB loss from body muscle of boiled crab. Most of the PCBs lost from the crabs were transferred to the cooking water. 21.4.2 Inorganic Blue crabs were harvested by Ward et al. (1979) from an area known to have a high level of mercury in the sediment. The edible body meat was extracted and analyzed for both total and methylmercury concentrations. A highly significant correlation was observed (0.957) with respect to the amount of total mercury
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Shellfish safety and quality
present as methylmercury. Analysis indicated that approximately 35% of the total mercury was present as methylmercury. The US Environmental Protection Agency (EPA) (Government Accounting Office, 2002) estimated the average daily exposure of an American adult to dioxins from ten food categories, including both freshwater fish and shellfish and marine fish and shellfish. Exposure estimates were based upon both concentrations of dioxins observed in food samples and estimations of daily consumptions. Freshwater fish and shellfish were estimated to be the secondhighest contributing group (contributing just over 20% of daily intake), exceeded only by beef. Marine fish and shellfish were the fifth highest contributor, at just under 8% of daily intake. It should be noted that the number of samples analyzed for dioxin were limited, and all samples were from North American sources. 21.4.3 Biological/Listeria monocytogenes In a survey of frozen seafood products including, shrimp, crabmeat and lobster tails, 15 out of 57 were found to contain L. monocytogenes. L. monocytogenes has also been isolated from cooked crab meat (Weagant et al., 1988; Farber, 1991) and is able to grow on inoculated refrigerated crab meat as well as on cooked shrimp, cooked crawfish tail meat and canned lobster meat stored at 4± 5 ëC. An increase between 2 to 3 logs in 1±2 weeks was observed (Brackett and Beuchat, 1990; Farber, 1991; Dorsa et al., 1992). Weagant et al. (1988) reported 35 of 57 samples (61%) tested positive for Listeria spp. and 15 of 57 (26%) tested positive for L. monocytogenes. Listeria can grow in the pH range of 5.0 to 9.5 in a good growth medium. Listeria is salt tolerant, surviving 100 days in concentrations of salt as high as 30.5% at 4 ëC. Listeria has been reported to survive for 3 months in dry fodder and more than 6 months in dry straw (Lovett and Twedt, 1988). Commerical pasteurization of crabmeat was found to inactivate Listeria (Dillon and Patel, 1991). Noah et al. (1991) isolated Listeria from naturally contaminated frozen seafood consisting of lobster tails, shrimp, prawn, and breaded shrimp. Buchanan et al. (1989) found seafood to have a 28% incidence rate of Listeria, second only to meat. Several fish products have been recalled due to Listeria contamination: frozen canned cooked crabmeat (FDA, 1991), cooked shrimp (Anonymous, 1988a), frozen cooked shrimp, canned frozen, fresh and imitation crabmeat, and frozen lobster (Anonymous, 1988b; Farber and Peterkin, 1991). Seafoods were implicated as the leading food in the transmission of foodborne illness from 1977 to 1983, 24.8% of all foodborne illness. Again between 1984 and 1987 seafood was indicated in 22.4% of foodborne illness. The estimated cost for foodborne illnesses for 1987 was US$4.8 billion. L. monocytogenes ranked fifth on the descending cost scale with 1581 recorded cases resulting in a US$213 million loss. Industry costs for contamination of a product with L. monocytogenes can include the cost of recalling and destroying,
Slaughter, storage, transport, and packaging of crustaceans
555
reduced consumer demand, investigating the source of contamination, clean-up, changes in production, liability suits, product spoilage, disrupted work schedules, and final closure (Miller et al., 1990). Heat resistance of L. monocytogenes Heat resistance of L. monocytogenes is influenced by many factors such as strain variation, previous growth conditions, exposure to heat shock, acid and other stresses, and composition of the heating menstruum (Doyle et al., 2000). Heat resistance data for different L. monocytogenes strains have shown that under similar conditions some strains are 2.5 to 3 times more resistant than others. Cells in the stationary phase appear to be more resistant to thermal stress. Temperature and the composition of growth medium affect rates of growth and the synthesis of cellular constituents that determine the thermal tolerance. Heat resistance also varies depending on the food, L. monocytogenes was found to be more heat resistant in beef than in chicken or carrot homogenates (Gaze et al., 1989). Heat shock results in increased heat resistance (Farber, 1989; Fedio and Jackson, 1989; Bunning et al., 1990, 1992; Knabel et al., 1990; Linton et al., 1992; Carlier et al., 1996; Augustin et al., 1998; CDC, 1999; Jùrgensen et al., 1999). The degree of thermal resistance is strain dependent and varies with the length of heat shock, the pH of medium, and the growth phase of the cells. Several factors were found to significantly affect thermal destruction of L. monocytogenes in shellfish, including presence of salt, smoke, or liquid smoke, and whether there was a cover during cooking. Published heat-resistant data for L. monocytogenes reports that cooking food to an internal temperature of 70 ëC for 2 min is adequate to ensure destruction of L. monocytogenes (MacKay and Bratchell, 1989). Crab Rawles et al. (1995a) reported 13 out of 126 samples of blue crab tested positive for Listeria, with 10 being L. monocytogenes, and 3 L. innocua. Levels found in the meat were less than 100 cfu/g except one, which was 1100 cfu/g, using the MPN method. L. monocytogenes had an increased growth rate as the storage temperature increased, with approximately a 7 log10 increase in population at 5 ëC and only a 2.5 log10 increase at 1.1 ëC after 21 days. The suggestion of the authors is to maintain storage temperatures at or below 1.1 ëC for fresh blue crab. Rawles et al. (1995b) also reported 35 out of 57 (61%) frozen seafood samples, including raw, cooked and peeled shrimp, cooked crab meat, and raw lobster tails, contained Listeria spp., 15 (26%) of these samples contained L. monocytogenes. Brackett and Beuchat (1990) demonstrated that L. monocytogenes grows well at refrigerated temperatures on crabmeat, with a 2±5 log10 increase over a 2-week storage period at 5 and 10 ëC. Crawfish Gray and Killinger (1966) report that L. monocytogenes flourishes in aquatic environments of decaying vegetative matter, where crawfish typically are found.
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Shellfish safety and quality
Consequently, crawfish often enter processing plants carrying L. monocytogenes. Dorsa et al. (1992) state that the initial population of L. monocytogenes on inoculated crawfish stored at 0 and 6 ëC was approximately 104/g. At 0 ëC, less than 1 log10 growth was observed for the entire storage time of 20 days. Generation time of L. monocytogenes at 0 ëC was 72.2 h, much longer than at 6 or 12 ëC, 17.0 and 6.9 h respectively. L. monocytogenes grows poorly at 0 ëC, so crawfish should be held at a constant temperature of 0 ëC. At 6 ëC, L. monocytogenes began exponential growth immediately with no lag phase. A 1 log10 increase per 2-day period was observed until day 10, when a stationary phase was reached. This temperature is important when determining the growth rates of L. monocytogenes in crawfish tail meat since 6 ëC is close to the temperature of many retail and home refrigerators. At 12 ëC, the initial growth of L. monocytogenes (105/g) underwent rapid exponential growth for 3 days, when a stationary phase was reached. 12 ëC is not considered refrigeration; however, it is possible to reach this temperature if product mishandling occurs during transportation or storage. In the seafood industry, a 7±10-day shelf-life is typical for fresh, iced seafood products. Observed D values were 10.23, 1.98, and 0.19 min at 55, 60, and 65 ëC, respectively, with a z value of 5.5 ëC. Harrison and Huang (1990) reported a D value of 12.00 min for 55 ëC and 2.61 min for 60 ëC for crab meat, and Budu-Amoako et al. (1992) determined D values of 2.39 min at 60 ëC for lobster. Commercially processed crawfish are boiled for 5±10 min before hand peeling. This should be sufficient to destroy any L. monocytogenes before hand peeling and packaging. The occurrence of L. monocytogenes in packaged crawfish is most likely due to cross-contamination during peeling and packaging. Crawfish processing at 100 ëC for 5±10 min prior to picking of the crawfish meat is sufficient to kill L. monocytogenes and many other bacteria (Dorsa et al., 1992). L. monocytogenes has been reported to grow on inoculated cooked lobster, shrimp, crab, and crawfish, even at refrigerated temperatures (Dorsa et al., 1992; McLauchlin, 1997; Oh and Marshall, 1995). Thimothe et al. (2002) found 23 of 78 samples tested positive for Listeria, 3 were positive for L. monocytogenes. For environmental samples (floors, drains, doors, tables, etc.), 8 out of 181 samples tested positive for Listeria, with one positive for L. monocytogenes, none were from food contact surfaces. No Listeria was found on processed crawfish samples. Listeria monocytogenes was recovered from 3% of whole boiled market crawfish samples and 17% of frozen vacuum-packed partially cooked crawfish tail meat (McCarthy, 1996). McCarthy's results suggest that the survival and growth characteristics of L. monocytogenes are dependent on storage time and temperature and the nature of the seafood product. The incidence of L. monocytogenes contamination of imported and domestic seafood in the US is 5± 6% (Kvenberg, 1988). McCarthy (1996) found that storing cooked crawfish at 22 ëC for extended periods results in an increase in L. monocytogenes, causing a potential risk increase to consumers. Exoskeletons of whole boiled crawfish were inoculated with 3.0 log10 L. monocytogenes cells per gram and stored at 6
Slaughter, storage, transport, and packaging of crustaceans
557
or ÿ20 ëC. The Listeria survived but did not grow under these storage conditions. However, Dorsa et al. (1992) report growth of Listeria at 6 ëC. The psychrotrophic characteristics of Listeria are of great concern to seafood processors; its growth should be prevented owing to the lack of known infectious dose. Lobster Research found L. monocytogenes to be approximately 1 log10 cfu more on the shell than the flesh of lobster (Dykes et al., 2003). This was consistent for all their samples regardless of inoculation protocol used or peeling procedure. Their findings suggest that the shell only slightly protects the flesh, and that contamination of the flesh probably occurs before peeling. McCarthy (1992) suggests that chitin, the major component of prawn shells, enhances the attachment and growth of L. monocytogenes. Dykes et al. (2003) suggests that contamination of L. monocytogenes is not `on' the shell as much as `in' the crevasses and channels between the shell segments that open to the interior flesh. They conclude that only sufficient heating before consumption will effectively eliminate the L. monocytogenes. MacKay and Bratchell (1989) published heat-resistant data concluding that cooking foods to an internal temperature of 70 ëC for 2 min is adequate to ensure the destruction of L. monocytogenes. Budo-Amoako et al. (1992) investigated the thermal death time for L. monocytogenes in lobster meat and reported D values at 51.6, 54.4, 57.2, 60.0, and 62.7 ëC were 97.9, 55.0, 8.30, 2.39, and 1.06 min, respectively. The presence of L. monocytogenes in processed lobster may be due to undercooking or post-processing contamination from plant environment. Good personal and environmental hygiene can help minimize postprocessing contamination. Adequate cooking is required to eliminate naturally occurring L. monocytogenes. Budo-Amoako et al. (1992) speculate that the presence of L. monocytogenes in processed lobster meat is due to undercooking. Budo-Amoako et al. (1999) combined nisin and moderate heat and looked at the effects on killing L. monocytogenes. Cold pack lobster is currently pasteurized at 65.5 ëC for 13 and 18 min for 320 g and 906 g (11.3 and 32 oz) cans, respectively. This treatment causes product shrinkage and thermal breakdown of the lobster, shrinkage is 5% for the 320 g (11.3 oz) and 9% for the 906 g (32 oz) size. Quality factors affected are texture and color, owing to the heat. Some studies indicate that nisin has bactericidal and bacteriostatic properties against L. monocytogenes (Benkerroum and Sandine, 1988; Mohamed et al., 1984). Kalchayanand et al. (1992) demonstrated that a combination of nisin and moderate heat has a greater bactericidal effect than either one alone. Henning et al. (1986) suggests that moderate heat allows the nisin to penetrate the cell wall and allows easier access to the cytoplasmic membrane. Residual nisin in the meat was higher than that found in the brine. A temperature of 60 ëC for 5 min or 65 ëC for 2 min resulted in 3 to 5 log10 reductions of L. monocytogenes. pH adjustments of the brine from 8 to 5 resulted in a slightly higher reduction in L. monocytogenes. Product shrinkage was improved with the
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Shellfish safety and quality
use of nisin with moderate heat, 1 and 1.28% for 302 g and 906 g (11.3 and 32 oz) cans, respectively, compared with 5 and 9%. Texture and color attributes were not discussed. Shrimp The highest incidence of Listeria spp. from water and shrimp occurred when the water temperature was 20 ëC (Motes, 1990). Motes also reported that salinity had little effect on the recovery of Listeria spp. from shrimp. The frequency of recovery (11%) of L. monocytogenes from unprocessed shrimp is low compared with levels reported by others. Weagant et al. (1988) isolated L. monocytogenes (26%) and L. innocua (46%) from 57 frozen seafood products. L. monocytogenes was recovered from 28% of raw frozen shrimp products. Fuchs and Surendran (1989) isolated L. innocua from 30% of fresh fish and fishery products. The higher incidence of Listeria spp. in processed foods may come from post-process contamination. Listeria spp. were found in freshly caught seafood and their harvest waters. The occurrence was not related to fecal coliform levels or salinity levels, but higher incidence occurred when water temperatures were 20 ëC. Hartemink and Georgsson (1991) found L. monocytogenes in 9% of raw frozen shrimp, Lovett et al. (1990) report an increase in the load of L. monocytogenes in shrimp heads at 7 ëC, and Farber (1991) showed that artificial inoculation of L. monocytogenes growth on shrimp stored at 4 ëC resulted in growth of 2±3 logs10 within 7 days. Harrison et al. (1991) studied the fate of L. monocytogenes on frozen shrimp and found less than 1 log10 decrease after 3 months at ÿ20 ëC. Jeyasekaran et al. (2002c) reported that L. monocytogenes populations remain constant for the first 30 days of storage at ÿ18 to ÿ20 ëC; after 60 days the count was double; at 90 days the count was back down to the 30day level. The conclusion is that L. monocytogenes is not influenced by freezing. Ben Embarek (1994) suggests that Listeria spp. are likely to occur naturally on freshwater fish but not likely to naturally occur on fish reared in clean seawater. Motes (1990) isolated L. monocytogenes from live shrimp, but attributes contamination in processing plants and not the shrimp themselves as the most likely source. Raw shrimp have been found to contain L. monocytogenes between 1.5 and 20% of the time (Ben Embarek, 1994). The concern is that the same rate of contamination can also be found in shrimp processing plants. Jeyasekaran et al. (2002b) found the incidence of Listeria spp. in shrimp to be as high as 14.3%. Headless shrimp were found to contain Listeria spp. 44.4% of the time, and peeled and unveined shrimp had a 23.1% incidence rate. Peeled and deveined shrimp had no Listeria in the samples tested. Farber (1991) states that retail shrimp are often found to be positive for L. monocytogenes. L. monocytogenes was shown to grow on cooked lobster, shrimp, and crab, increasing 2±3 logs10 in 7 days at 4 ëC. L. monocytogenes grown at room temperature, to simulate temperature abuse, on shrimp, crab, and lobster increased 1.0, 1.0, and 0 to 1.0 logs10 respectively. Using limited samples, Farber (1991) found that ready-to-eat shrimp and crab were positive for
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L. monocytogenes. Farber reports that the level of L. monocytogenes, although present, was < 10 MPN/g. Even with growth Farber suggests that this starting level does not represent a serious health risk. Equipment Beresford et al. (2001) have shown that L. monocytogenes, which forms biofilms that are much more resistant to disinfectants and sanitizers than planktonic cells, adheres to many materials found in the food processing industry. The materials tested included stainless steel, aluminum, polycarbonate, polypropylene, polyurethane, polyvinylchloride, poly(ethylene-terephthalate glycol) (PETG), poly(tetrafluoroethene) (PTFE), Lexan, nitryl rubber, silicone rubber, natural white rubber and ethylene propylene diene M-class (EPDM) rubber. Their results showed that none of the surfaces had significantly more or less adhering cells. All the surfaces tested had adhering cells as well as shed cells from washing. Blackman and Frank (1996) looked at TeflonÕ, nylon, and polyester floor sealant. They found all three able to support Listeria biofilms. Mafu et al. (1990) also found that the pathogen attaches to glass, polypropylene, and rubber. Frank and Koffi (1990), Lee and Frank (1991), and Ronner and Wong (1993), reported production of a sanitizer-resistant biofilm on glass, stainless steel, and Buna-N rubber. These biofilms were found to be resistant to chlorine, iodine, acid anionic, and quaternary ammonium sanitizers. Arizcun et al. (1997) looked at effective ways to remove Listeria biofilms from processing plant surfaces (glass was used). The most effective temperature found was 63 ëC. L. monocytogenes biofilms were found to be not very susceptible to high osmolarity (10.5% NaCl), and the interaction of sodium chloride and acid did not seem to have important effects in inactivating these bacteria (1.3±1.9 log10 reduction). The combination of NaOH (pH 10.5; 100 mM) and acetic acid (pH 5.4; 76.7 mM) applied sequentially at 55 ëC for 5 min was shown to be the most effective treatment to remove L. monocytogenes biofilms (4.5±5.0 log10 reduction). Recontamination from the processing environment is the principal source of Listeria contamination of processed ready-to-eat foods (Tompkin, 2002). Some harsh conditions in the factory environment (e.g. acidity) can result in crossprotection of Listeria to other stresses (e.g. heat) (Lou and Yousef, 1999). The probability of ready-to-eat product contamination is affected by a number of variables, including but not limited to (a) proximity of microbial growth niches to the product stream, (b) number of growth niches, (c) spatial relationship of niches to product stream, (d) microbial populations in niches, (e) extent of niche disruption, and (f) exposure of product stream to the environment (Faust and Gabis, 1988). Product handlers Jeyasekaran et al. (2002a) reported a 4% incidence of positive L. monocytogenes on the hands of seafood handlers (2 out of 50). From two processing plants there was also a 15% (4 of 59) incidence of Listeria on tables, floors, drains,
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containers, and equipment. Their conclusion was that since the handlers did not have infections in their hands, the Listeria was probably due to contamination. Listeria found in the plate freezer, on processing tables and in floor drains suggest contamination with raw product and then insufficient cleaning procedures. Once Listeria adheres to a surface, regular cleaning and sanitizers will not be effective in removing Listeria biofilms. Kerr et al. (1993) reported a 7% incidence of L. monocytogenes in food production workers and 12% positive for Listeria spp. Of the 99 workers from 44 establishments tested, 6 were positive for Listeria, but they were able to remove the contaminant with hand-washing. In 4 others, Listeria numbers were reduced but not eliminated. In two cases hand-washing increased the level of Listeria. The hand-washing technique for the workers was deemed inadequate in 10 out of 12 (1 was not retested). There was failure to use soap and/or other washing agent, short duration for washing (less than 5 s), and the use of visibly dirty towels was noted. The importance of proper hand-washing techniques cannot be stressed enough, especially where there are both raw and ready-to-eat food products. Disinfectants Cetylpyridinium chloride (CPC) is a cationic surfactant belonging to the group of quaternary ammonium compounds (QACs), which are successful antiseptics and disinfectants. The mechanism of action of QACs on bacteria has been proposed to include the sequence of (1) adsorption and penetration of porous cell wall; (2) interaction with cytoplasmic membrane (lipid±protein) followed by membrane disorganization; (3) leakage of intracellular low molecular weight constituents, such as amino acids, nucleotides, ions; (4) degradation of proteins and nucleic acids; and (5) lysis due to wall-degrading autolytic enzymes (Salton, 1968). Although CPC is not an FDA-approved antimicrobial agent for seafood, it has been shown to reduce L. monocytogenes (between 2.5 and 7.0 logs10) on inoculated raw, cooked shell-on, and peeled shrimp at lower levels than approved for poultry (Dupard et al., 2006).
21.5
Conclusions
The cooking and processing of most crustacean species have not significantly changed over time. The processing of blue crabs (Callinectes sapidus) was first commercialized in the late 1800s and the same hand-picking procedures are employed in the twenty-first century. The animals are placed either in boiling water or in a steam vessel and the meat removed primarily by hand with some mechanical processing. Many of the recent processing procedures were developed to maximize meat recovery from parts that were previously discarded, such as legs and small claws. The meat obtained from these processes results in the production of products having low economic value. However, byproduct recovery is becoming more important as an ancillary process operation
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to reduce solid and liquid wastes while also providing some additional revenue. Two new methods of processing, high hydrostatic pressure processing and pasteurization, are receiving greater importance. High hydrostatic pressure has been used to remove the meat from lobsters without the application of a thermal process. This process provides greater yield and also produces a new product into the marketplace, uncooked lobster meat. Pasteurization is able to extend the shelf-life of crustacean meat for up to several years, provided adequate storage conditions are employed. The process has recently created an opportunity for crab meat to become internationally distributed. These two new processes also reduce product losses to the processor, retailer, and consumer through their extended shelf-life. An emerging food safety concern is the presence of chemical contaminants in crustacean meat. Antibiotic and other drug residues are an increasing concern since their effects on the body have not been fully defined. Their presence will continue to increase and impact producers, importers, distributors, and consumers as new analytical methods for their detection are developed. Chemical residues will continue and increase in severity since many of the pollutants have become widely dispersed in the environment and no cost-effective methods exist to reduce their presence. Since most crustacean species are given a thermal process to reduce pathogenic microorganisms and to facilitate the processing process, the potential for foodborne illness through the presence and proliferation of Clostridium botulinum and Listeria monocytogenes continues to be a major concern. Time±temperature indicators may be required of all ready-to-eat crustacean products so that any temperature abuse can be readily identified.
21.6
References
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and BERNARD D T (1982) Heat resistance of spores of nonproteolytic type B Clostridium botulinum, J Food Protect. 45 909±912. SHAKILA R J, VYLA S A P, KUMAR R S, JEYASEKARAN G and JASMINE G I (2006) Stability of chlorampenicol residues in shrimp subjected to heat processing treatments, Food Microbiol. 23 (1) 47±51. SOLOMON H M, KAUTTER D A and LYNT R A (1982) Effect of low temperatures on growth of nonproteolytic Clostridium botulinum types B and F and proteolytic type G in crabmeat and broth, J Food Protect. 45 516±518. THIMOTHE J, WALKER J, SUVANICH V, GALL K L, MOODY M W and WIEDMANN M (2002) Detection of Listeria in crawfish processing plants and in raw, whole crawfish and processed crawfish (Procambarus spp.), J Food Protect. 65 (11) 1735±1739. TOMPKIN B A (2002) Control of Listeria in the food-processing environment, J Food Protect. 65 709±723. TSENG Y-C, XIONG Y L and WEBSTER C D (2005) The preservation of the quality of the muscle in frozen Australian red claw crayfish (Cherax quadricarinatus) by prestorage and antioxidant dipping treatments, Int J Food Sci Technol. 40 841±848. WANG M Y and BROWN W D (1983) Effects of elevated CO2 atmosphere on storage of freshwater crayfish (Pacifastacus leniusculus), J Food Sci. 48 (1) 158±162. WARD D R, PIERSON M D and VAN TASSELL K R (1977) The microflora of unpasteurized and pasteurized crabmeat, J Food Sci. 42 597±600, 614. WARD D R, NICKELSON II R and FINNE G (1979) Relationship between methylmercury and total mercury in blue crabs (Callinectes sapidus), J Food Sci. 44 920±921. WARD D R, NICKELSON R, FINNE G and HOPSON D J (1983) Processing technologies and their effects on microbiological properties, thermal processing efficiency, and yield of blue crab, Mar Fish Rev. 48 (7,8,9) 91±102. WARD D R, PIERSON M D and MINNICK M S (1984) Determination of equivalent processes for the pasteurization of crabmeat in cans and flexible pouches, J Food Sci. 49 1003± 1004, 1017. SCOTT V N
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and THAYER C F (1988) The incidence of Listeria species in frozen seafood products, J Food Protect. 51 655±657. WENNER A M (1986) Crustaceans and other invertebrates as indicators of beach pollution. In: Saue D F and Kleppe G S, Marine Organisms as Indicator, New York, Springer-Verlag, p. 209. WENTZ B A, DURAN A P, SWARTZENTRUBER A, SCHWAB A H, MCCLURE F D, ARCHER D and READ R B (1985) Microbiological quality of crabmeat during processing, J Food Protect. 48 (1) 44±49. YERLIKAYA P and GOKOGLU N (2004) Quality changes of blue crabs (Callinectes sapidus) meat during frozen storage, J Food Qual. 27 83±89. YOUNG D R (1982) Chlorinated hydrocarbon contaminants in the Southern California and New York bights. In: Mayer, G F, Ecological Stress and the New York bight: Science and Management, Estuarine Research Foundation, Columbia, SC, p. 271. ZABIK M E, HARTE J B, ZABIK M J and DICKMANN G (1992) Effect of preparation and cooking on contaminant distributions in crustaceans: PCBs in blue crab, J. Agric. Food Chem. 40 1197±1203.
22 Packaging, storage and transport of molluscan shellfish V. Garrido, Institute of Food and Agricultural Sciences, USA and G. E. Rodrick, University of Florida, USA
Abstract: The product specifications and packaging is of major importance in commerce of molluscan shellfish. All molluscan shellfish processors and buyers must have well-defined specifications for each molluscan shellfish product they handle. Product description for molluscan shellfish should include: (1) a general description of the product; (2) the product source; (3) the product form and packaging style; (4) master packaging and shipping format; (5) the product weight (master and individual containers); (6) labeling; (7) storage and shipping temperatures; and (8) microbial specifications and maximum allowable levels. Key words: molluscan shellfish packaging, product specification, labeling, storage and shipping temperatures, product form and packaging styles.
22.1
Introduction
This chapter will present the different requirements and recommendations for product specification and packaging used by the molluscan shellfish industry. An emphasis on the United States market, either produced or exported to the US, will be discussed.
22.2
Product specification
All processors and buyers must have a clearly defined set of specifications on the products they handle. Written product specification must be developed internally or in conjunction between processors and buyers to define the quality
Packaging, storage and transport of molluscan shellfish
569
and safety parameters of the product. This document will facilitate communications between sellers and buyers by defining each other's expectations. The product specification document should be at least as stringent as the regulations and should always comply with the minimum requirements from the buyers. Product specification document should include the following: 1. 2. 3. 4. 5. 6. 7. 8.
General description of the product. Product source. Product form and packaging style. Master packaging and shipping format. Product weight (master and individual containers). Labeling. Storage and shipping temperatures. Microbial specifications. a. Maximum allowable levels. b. Type of microorganism monitored. 9. Product flavor and odor profile. 10. Food additives/preservatives added. 11. Grading system and workmanship.
22.3
Packaging formats and materials
The molluscan shellfish industry has traditionally used bags of many different materials, such as burlap (see Fig. 22.1), plastic mesh and others, to store and transport live shellfish (shellstock). Sometimes the industry use large plastic
Fig. 22.1
Typical burlap bags used for transporting oysters.
570
Shellfish safety and quality
Fig. 22.2
Plastic trays used for frozen shellfish commerce ± whole and shucked.
totes to transport bulk shipments of shellfish primarily oysters or plastic baskets in the case of clams and mussels. These bags, totes or baskets are mostly used to transport product from the harvest areas to the primary or secondary processing plants. Mussels and clams, when fresh, use bags at the point of sales. Oyster shellstock, on the other hand, are received at the processing plant in burlap bags or bulk totes from the harvest areas but when processed they are transferred to waxed boxes or trays depending on the final product and intended use (see Fig. 22.2). The traditional shellstock boxes are presented in four different formats, 50 lbs, 30 lbs, 20 lbs and 10 lbs (22, 13, 9, 4 kg; see Fig. 22.3) and are commonly referred by the number of units in each container i.e. 200, 300 or 400 counts. These boxes are typically made from waxed carton with perforations at the bottom of the box to be able to withstand the humidity and drain the melted ice. Oysters, mussels and clams are also found in the market as frozen shellstock packed in plastic trays with or without flexible films. These are intended for institutional or retail distribution and are presented in a 12-tray box with 12 units per tray. If the finished product is shucked meats, then there are two main presentations in the market; fresh/refrigerated meats packed into rigid plastic containers in 1 gallon (3.79 litres), 1 pint (0.24 litres) or 1 quart (0.95 litres) format as seen in Fig. 22.4 or frozen meats packed in flexible plastic pouches referred to as pillow-pouch. This format is normally packed as 2.5±3 lbs (1.1± 1.3 kg) units.
Packaging, storage and transport of molluscan shellfish
Fig. 22.3
571
Waxed carton boxes used for transport and sale of shellstock.
Fig. 22.4 Rigid plastic containers used to pack and sell shellfish meats.
22.4
Product labeling and tagging
Regulations in the US (FDA, 2005) require that all shellfish in commerce bears a label or tag where essential traceability data must be displayed. All shellfish are tagged initially at the time of harvest indicating the harvest area (official and common names), harvest date and in most cases the time of harvest. The name of the harvester and its ID number are also recorded, these tags are called harvester's tag. Harvesters are allowed to tag each individual bag of shellfish harvested or tag the entire load with a bulk tag indicating the size of the load. Once the product reaches the primary processor, the harvest information will be transferred to the dealer's tag which also includes the processor's name, address and official designation or certification and in some cases an expiration date as seen in Fig. 22.5. Processors are also required to include an advisory indicating that there is a
572
Shellfish safety and quality
Fig. 22.5
Typical molluscan shellfish dealer's tag.
risk associated with the consumption of raw animal proteins such as shellfish (FDA, 2005). Figure 22.6 shows a typical consumer information statement that is found on the back of the dealer's tag or in the label of any shellfish container. This consumer information statement is required by state law in the states of Florida, Louisiana and Texas. Shellfish that have gone through a post-harvest processing (PHP) to reduce Vibrio bacteria or shucked meats (fresh or frozen) are regularly marketed with labels instead of tags as shown in Fig. 22.7. These labels contain portions of the tag information in addition to a unique lot number which could be used to traceback the product in case of an illness investigation or a recall. Additionally, shellfish packaging labels have to comply with other US regulations such as Country of Origin labeling (USDA, 2004) where the product has to be clearly marked as per origin (wild or farm-raised) and the country where the product originates (i.e. product of the US, China or Canada) as well as nutritional panel requirements if the product is packed to be sold in the retail market (FDA, 2001).
Fig. 22.6 Consumer information statement on back of dealer's tag.
Fig. 22.7
Packaging label for post-harvest processed oysters.
574
22.5
Shellfish safety and quality
Product size standards
The shellfish industry has used grading standards for many years to be able to communicate with buyers. These standards have no regulatory basis but are defined by commerce. Table 22.1 shows the relation between the various standard names and the actual product size or counts. In addition to the product size and counts, the industry has to define the quality standards accepted by the buyers. These could be defined as per allowable defects of color, fragments, and appearance, among others. Table 22.1
Quality standards for oysters ± shellstock and shucked meats
Product form
Grade
Count
Shellstock
Standards Selects Extra selects
200±250 per bushel 100±200 per bushel Less than 100 per bushel
Shucked meats
Very small Standards Selects Extra selects
Over 500 per gallon 300±500 per gallon 210±300 per gallon 160±210 per gallon
Conversion factors: 1 bushel 27.22 kg; 1 gallon 3.785 liters.
22.6
Accepting shellfish shipments
Shellfish shipments are considered acceptable when: (1) shipments are properly identified with tags and/or labels and shipping documents; (2) shellstock is alive and cooled to an internal shellstock body temperature of 50 ëF (10 ëC) or less; (3) shucked or post-harvest processed shellfish are cooled to a temperature of 45 ëF (7.2 ëC) or less; and (4) the time±temperature indicating device shows that the ambient air temperature has exceeded 45 ëF (7.2 ëC). Shellfish should be rejected when: (1) shellfish are not properly identified with tags or shipping documents; (2) the internal shellstock body temperature exceeds 60 ëF (15.6 ëC) unless the harvest initiation time can be documented and indicates that the time from harvest has not exceeded established Interstate Shellfish Sanitation Conference (ISSC) guidelines; (3) shucked shellfish exceeds 45 ëF (7 ëC) or (4) if the shellfish are unwholesome or unsafe for human consumption.
22.7
Conclusions
In summary, the packaging of shellfish is a critical part of shellfish commerce. Proper packaging of shellfish can facilitate product specifications, such as weight, temperature, size, location, traceability, consumer information, maximum allowable temperature and bacterial loads, and insure product safety.
Packaging, storage and transport of molluscan shellfish
22.8
575
References
(2001). Food labeling regulation (21 CFR part 101) http://www.fda.gov (2005). National Shellfish Sanitation Program (NSSP) Guide for the Control of Molluscan Shellfish. Chapters VIII and X. http://www.cfsan.fda.gov/~ear/nss3toc.html USDA (2004). Mandatory Country of Origen Labeling of Fish and Shellfish; Interim Rule (7 CFR Part 60). http://www.ams.usda.gov/AMSv1.0/getfile?dDocName= STELDEV3103356 FDA FDA
INDEX
Index Terms
Links
A abalone
276
Aber disease
276
absolute growth rate
432
absorption efficiency
206
accumulation of biotoxins
53
see also toxin accumulation modelling action plans
163
active management
488
active suspension ponds (ASPs)
344
adaptive immune system
369
additive genetic variance
428
adductor muscle
213
adenoviruses aeration aerosols, toxic
164
84 346
514
67
age and biofouling
320
selective breeding for growth
432
air drying Alexandrium monilatum
328
329
46
50
This page has been reformatted by Knovel to provide easier navigation.
Index Terms algal toxins
Links 43
129
60
142
67
153
economic impacts
66
176
future trends
67
human health impacts
57
detection and analysis new approaches
major algal toxins and their sources management responses
456
130 59
modelling see toxin accumulation modelling monitoring
47
60
origins of phycotoxins
45
47
processing for detoxification
63
61
regulation see regulation risk management
478
in crustaceans
498
in gastropods
499
post-harvest processing
490
risk assessment matrices
493
trophic dynamics
482
498
52
biotransformation
55
natural shellfish detoxification
56
toxin accumulation
53
toxin uptake
52
This page has been reformatted by Knovel to provide easier navigation.
66
Index Terms
Links
algal toxins (Cont.) see also amnesic shellfish poisoning (ASP); azaspiracid shellfish poisoning (AZP); diarrhetic shellfish poisoning (DSP); harmful algal blooms (HABs); neurotoxic shellfish poisoning (NSP); paralytic shellfish poisoning (PSP) algicides
187
‘all or none’ response
141
allergies, shellfish
458
aluminium sulphate (alum)
347
ammonia
188
ammonium sulphate
188
amnesic shellfish poisoning (ASP)
44
45
163
166
58
137
reduction in king scallops by shucking
485
487
491
risk matrix
494 47
49
131
146
149
effect on humans
toxins analytical techniques
58
see also domoic acid (DA) amoebic gill disease (AGD)
325
Amoebophrya spp.
185
analysis
145
see also detection
This page has been reformatted by Knovel to provide easier navigation.
162
137
Index Terms
Links
animals animal viruses and human disease
96
sources of microbial contamination
21
sources of viruses
25
552
561
110
antibiotics
345
antibody-based methods
279
algal toxins
22
61
143
integrated antibody-based screening systems shrimp disease antibody capture RT-PCR
154 408
410
95
antifouling coatings
327
331
Approved areas/waters
258
462
Aquasonic
189
Argonaute protein
366
Asia, Southeast
535
526
Asia Pacific Economic Cooperation (APEC) HAB monitoring assays
65 168 143
see also detection astroviruses
84
atherosclerosis
249
Aureococcus anophagefferens
184
Australia
535
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
Australian red claw crawfish
548
avian influenza H5N1 virus
97
azaspiracid shellfish poisoning (AZP)
44
46
47
50
131
138
151
162
131
138
166 effect on humans azaspiracids (AZA)
58
138
50
59
B bacteria
456
biological control of HABs
184
crustaceans preservation
554 549
faecal-borne see faecal-borne bacteria molluscan shellfish-borne diseases
108
management
258
bacteriophages bags
249
92 568
ballast water
26
barley straw
189
baskets
568
beach sands benzo[a]pyrene (BaP)
26 236
best management practices pond culture of crustaceans
339
SPF shrimp
414
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
between-family selection
430
Bilbao estuary
117
bioactivity of toxins
144
bioavailability of metals
242
biochemical fractionation techniques
233
biodeposits
514
biofilms
559
bio-flocs
344
biofouling
317
Plate I affected bivalves
318
current removal/treatment methods
325
future trends
331
problems and benefits
320
costs
323
disease
325
handling and packaging issues
325
shell quality
322
shellfish growth and predation issues
321
type and extent of fouling biokinetic model
319 231
234
biological control biofouling
330
HABs
183
191
biological treatment
114
115
bioluminescence
129
117
This page has been reformatted by Knovel to provide easier navigation.
Index Terms biosecurity and the culture of wild seed/broodstock
Links 398
405
413
442
413
through environmental control and best management practices
414
biosecurity plan
405
biosensors
119
Bioterrorism Act 2002
356
154
biotoxin contamination see algal contamination biotransformation of toxins toxin accumulation modelling birds, wild
55 203
215
221
25
blue (swimming) crabs
544
boiling
543
Bonamia spp.
272
Bonamia exitiosa
272
273
Bonamia ostreae
272
273
Bonamia roughleyi
272
273
bonamiasis-resistant oyster breeding programme
281
bone meat separator
544
botulism
258
boxes
570
571
48
57
breeding programmes see selective breeding brevetoxins
67
135 This page has been reformatted by Knovel to provide easier navigation.
131
Index Terms broodstock SPF shrimp
Links 426 413
brown crabs
498
buoys for HAB measurement
169
burlap bags
569
C cadmium
229
safety standards
240
uptake
235
California crawfish
548
Campylobacter spp.
251
230
231
Canada cholesterol
466
depuration
537
capillary electrophoresis (CE)
61
catastrophic natural events
19
ceilings
525
cell culture-reverse transcriptasepolymerase chain reaction (CC-RT-PCR)
92
cellular debris
233
cellular toxin content
222
certification
520
cetylpyridinium chloride (CPC)
560
chemical analysis techniques for toxins chemical control of HABs
56
145
187
191
This page has been reformatted by Knovel to provide easier navigation.
233
Index Terms
Links
chill killing
351
China
470
clay flocculation
178
chloramphenicol
553
chlorine
515
cholera
254
cholesterol
466
471
67
134
ciguatoxic fish poisoning (CFP) clams
182
biofouling
323
US production
295
classification of growing waters
518
11
258
clay flocculation
177
190
applications
178
concept impacts on benthos
462
526
526
528
177 182
cleaning depuration facilities
531
shells
327
climate change
27
120
Closed areas
258
462
closed depuration systems
510
Clostridium spp.
258
closures, bed
14
65
coastal demographic increase
113
114
coastal management
114
121
Cochlodinium polykrikoides
178
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
Codex Alimentarius Commission (CODEX)
240
297
497
498
470
Codex Committee on Fish and Fishery Products (CCFFP) Codex Committee on Food Hygiene (CCFH) coliforms, faecal
497 4
7
12
249
457
528
depuration and
523
528
management of
258
maximum zero-hour level of
523
colloidal organic carbon (COC)
238
commercial fisheries, impacts of HABs on
66
communication
27
competitive algae
186
Conditionally Approved areas
258
485
526
confidence, public see public confidence confidence levels
483
construction of depuration plant
524
consumer demand
121
consumer information statement
571
529
contaminants crustaceans
552
see also algal toxins; faecal-borne bacteria; microbial contamination; vibrios; viruses
This page has been reformatted by Knovel to provide easier navigation.
88
Index Terms
Links
control strategies
176
controlled purification see depuration cooking/heat treatment algal toxins
63
crustaceans
543
vibrios
489
viruses
86
cooler temperature monitoring record
490
88
313
cooler thermometer calibration/accuracy verification log
314
cooperation
500
copper
230
copper sulphate
187
coronary heart disease (CHD)
466
coronaviruses corrective actions (CAs) HACCP
234
468
97 357 297
307
309
costs biofouling HABs
323 66
176
modelling and reducing costs of monitoring biotoxins
220
country of origin labelling
572
crabs
323
biotoxins in
498
control of biofouling
330
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
crabs (Cont.) freezing
551
irradiation
551
Listeria monocytogenes
555
pasteurisation
549
slaughter/cooking
543
crawfish Listeria monocytogenes
555
processing
547
critical control points (CCPs)
297
301
307
critical limits (CLs)
297
306
307
Cryptosporidium parvum cylindrospermopsin
465
7 67
D Danish algal toxin monitoring programme
163
dealer’s tag
571
572
decarbamoylations
216
218
Denmark
536
depth of gear
326
depuration algal toxins natural depuration
27
56
261
63 56
around the world
535
equipment construction and facility design
529
HACCP
465
521
This page has been reformatted by Knovel to provide easier navigation.
509
Index Terms
Links
depuration (Cont.) importance of seawater quality
511
national legislation
461
plant location, design and construction
524
relaying
537
risk management
488
rules and guidelines
519
source of shellfish
526
toxin accumulation modelling
202
types of depuration plant
510
types of seawater treatment
515
and viral contamination
204
206
208
95
110
120
60
67
142
89
120
499
46
162
163
detection algal toxins metals and organic contaminants
241
shrimp pathogens
407
viruses
12
diagnosis diseases of molluscan shellfish shrimp disease vs surveillance/screening diarrhetic shellfish poisoning (DSP)
279 407 411 44 166
accumulation
55
effect on humans
57
risk matrix
136
495
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
diarrhetic shellfish poisoning (DSP) (Cont.) toxins
47
49
analytical techniques
151
152
modelling toxin accumulation
217
dicer gene
366
DiCANN
170
dietary advice
465
131
136
131
136
dietary uptake organic contaminants
236
metals
234
digestive gland
213
Dinophysis acuminata
171
dinophysistoxins (DTX)
49
57
217 dioxins
554
dipping methods for cleaning
328
discharge water
341
329
disease resistance selective breeding
281
437
282
368
SPF shrimp see specific pathogen-free (SPF) shrimp transgenic animals diseases of shellfish
270
biofouling and
325
diagnostic methods
279
effects on international shellfish industry
279
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
diseases of shellfish (Cont.) future trends
283
major pathogens and diseases of molluscs reducing disease in aquaculture disinfection
271 281 27
Listeria monocytogenes
560
viral contamination
114
115
dissolved phase metal uptake
234
organic contaminant uptake
236
distribution pathogens associated with
257
259
261
safety and quality standards
352
DNA-based molecular methods
403
408
409
410
dogwhelk
330
dominance genetic variance
428 49
55
58
62
131
137
163
analytical techniques
146
149
depuration
490
shucking to reduce ASP in king scallops
491
domoic acid (DA)
see also amnesic shellfish poisoning (ASP) double-stranded RNA (dsRNA)
365
response to in shrimp
367
drain harvesting
348
Drinking Water Directive
118
dropping lines 327 This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
E Eagle Creek
117
early warning system
120
efflux rate
239
electron beam irradiation
550
electron microscopy
169
emerging diseases
97
encephalomiocarditis virus (EMCV)
94
end-product standards
531
end-product testing
492
endobionts
322
environment conditions and viral contamination
112
environmental effects on microbial contamination
15
genotype–environment (G × E) interactions
436
regulation of shellfish growing water quality
11
environmental control
414
environmental virology
85
258
462
526
147
149
150
enzyme-linked immunosorbent assay (ELISA)
143 152
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
epimerisations
216
epistatic genetic variance
428
218
equipment depuration
529
Listeria monocytogenes biofilms
559
Escherichia coli (E. coli)
249
E. coli ribotyping
23
ESP Environmental Sample Processor
170
Europe
535
European Union (EU)
88
algal toxins
64
457
118
Bathing Waters Directive
477
Framework Water Directive
118
HAB monitoring
167
HACCP
298
hygiene regulation
461
478
metals
240
241
shellfish growing water quality
463
external harvest basin
349
external variables, models with
211
212
4
6
119
F faecal-borne bacteria in beach sands disease
26 8
environmental effects on contamination
17
recovery and growth after disinfection
27
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
faecal-borne bacteria (Cont.) risk management
478
sources
20
survival capacity
20
water standards
12
479
493
6
12
faecal coliforms see coliforms, faecal faecal indicators
4
88
249 and viral contamination
109
119
faecal load, limiting
114
familial hypercholesterolaemia (FH)
467
468
family selection
430
441
131
140
153
110
111
‘fast-acting’ lipophilic toxins (macrocyclic spiroimines) feeds
344
feed-based therapeutic agents
345
storage
345
Fenneropenaeus chinensis
404
fertilisation
342
filtration rate
53
finfish
26
AGD
325
benefits of vaccines to aquaculture
362
flesh samples, examination of
168
flocculation see clay flocculation flooding
19
This page has been reformatted by Knovel to provide easier navigation.
112
Index Terms floors
Links 525
flow cytometry
92
flow rate
530
flow-through depuration plants
510
FlowCAM
170
Food and Agriculture Organization (FAO)
497
food business operators cooperation and resources shared with regulators
500
onus of risk management on
492
role in risk management
485
487
14
65
HACCP
297
309
food security plan
356
foot
213
foot-sweeping
320
formulated feeds
344
Framework Water Directive (FWD)
118
Food and Drug Administration (FDA)
France
89
free-ion activity model (FIAM)
237
FreeCalc
413
freezing
261
crab
551
HACCP and processing of oysters
300
FRNA phages
92
functional assays
67
470
536
311
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
G gamma irradiation
261
gastropods
499
gene probes
119
General Agreement on Tariffs and Trade (GATT)
489
550
64
general surveillance
411
genetic variance
427
genotype–environment (G × E) interactions germplasm protection
436 443
Giardia lamblia
7
gills and mantle
213
global warming
27
gonad goniodomin A
120
213 50
good management practice (GMP)
464
good manufacturing practices (GMP)
299
grading standards
574
grazers biological control of HABs
186
control of biofouling
330
Greenland
167
growth selective breeding for
432
shellfish growth and biofouling
321
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
growth rate
239
Gulf Shellfish Challenge
262
gymnodimines
131
140
H haemolysins
50
hand-held toxin detectors
154
hand-picking
544
560
hand-washing
526
560
handling effects of biofouling
325
food handlers and Listeria monocytogenes
559
infected food handlers
458
pathogens associated with
257
Haplosporidium costale (SSO disease)
275
Haplosporidium nelsoni (MSX disease)
274
harmful algal blooms (HABs) impacts of
44
261
62
129
175
economic impacts
66
impact on bivalves
67
mitigation of effects
259
176
175
approaches to management
176
biological control
183
191
chemical control
187
191
clay flocculation
177
190
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
harmful algal blooms (HABs) (Cont.) ethos of HAB control
190
future trends
190
ultrasonic devices
189
modelling dynamics of
169
monitoring
61
action plan design
164
future trends
170
methods and techniques
168
regulation
166
toxin accumulation modelling and
220
new approaches to detecting
162
67
see also algal toxins; toxin accumulation modelling harmonization of regulations Harris claw machine harvest water classification harvester’s tag
64 544 11
258
462
526
301
302
353
571
harvesting impact of microbial contamination
10
pond culture of crustaceans
348
safety and quality standards
350
hatchery
374
hazard analysis
297 354
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
hazard analysis and critical control point (HACCP) programmes
295
499
corrective actions
297
307
309
critical control points
297
301
307
critical limits
297
306
307
depuration
465
521
HACCP plan
299
300
353
hazard analysis
297
301
302
353
311
465
354 monitoring
297
306
prawn production
352
354
prerequisite programmes
298
299
principles
297
process flow diagram
299
processing of frozen raw oysters
298
311
records and documentation
297
307
310
verification
297
307
309
health dietary and health advisories effects of microbial contamination effects of shellfish disease impacts of algal toxin contamination public confidence in shellfish
465 7 280 57 455
public perception of benefits and risks associated with shellfish risk management and occurrence of illness shellfish-borne transmission of viral infections
469 478 85
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
health management programmes disease management strategies
283
using vaccines as part of for shrimp
372
heat resistance
555
heat-sensitive proteins (HSP)
233
239
5
85
87
88
109
253
457
471
91
94
hepatitis A virus (HAV)
detection monitoring viral contamination
108
hepatitis E virus
85
hepatopancreas
547
hepatopancreatic parvovirus (HPV)
360
heritability estimates
428
selective breeding for growth Heterosigma akashiwo
433
97
361
434
181
high hydrostatic pressure (HHP) processing
261
544
551
561
(HPLC)
61
67
145
152
domoic acid
150
HPLC-UV
61
NSP toxins
151
PSP toxins
147
high performance liquid chromatography
Hong Kong
240
241
This page has been reformatted by Knovel to provide easier navigation.
Index Terms Hurricane Katrina
Links 19
husbandry biofouling control
325
pond culture of crustaceans
339
fertilisation and semi-intensive systems
331
342
formulated feeds and intensive pond culture systems
344
harvest
348
land selection
340
substrate
345
water quality management
346
water source and conservation
341
hydrogen peroxide
188
hydrolysis
217
hygiene legislation
461
hypochlorite
509
hypothesis-testing
203
515
518
I immune system, invertebrate adaptive in situ HAB monitoring in vitro assays
364 369 169 61
143
inbreeding
431
442
inbreeding depression (IBD)
442
443
incorporation, toxin
204
206
444
This page has been reformatted by Knovel to provide easier navigation.
Index Terms individual (mass) selection
Links 430
440
360
361
385
389
392
398
403
413
infectious hypothermal and haematopoietic necrosis virus (IHHNV) selective breeding for IHHNV resistance SPF shrimp
437 387
IHHNV-resistant Litopenaeus stylirostris
390
MCE experiment
394
USMSFP
398
inhibitors
94
inorganic fertiliser
343
insoluble fraction
233
inspections of depuration plants
520
239
integrated antibody-based screening systems integrated studies for microbial sources intensive pond culture systems interim control plan
154 25 344 14
International Conference on Molluscan Shellfish Safety
27
International Council for the Exploration of the Sea (ICES) Guidelines International Mussel Watch Programme international policies
399
401
228 64
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
International Principles for Responsible Shrimp Farming
340
397
international shellfish industry effects of shellfish disease
279
economic losses
279
trade and health issues
280
effects of viral contamination
88
microbial contamination and
10
Interstate Certified Shellfish Shippers List (ICSSL)
520
Interstate Shellfish Sanitation Conference (ISSC) invertebrate immune systems adaptive
65
296
300
134
180
364 369
iodine
469
iodophors
518
irradiation
550
Italy
536
J Japan
178
K kainic acid
138
kainoids
58
Karenia brevis
67
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
kidney
213
kinetic model
231
Korea, South
178
234
L labelling
571
depurated shellfish
533
land selection
340
lateral flow immunoassay (LFI)
147
149
152
legislation environmental
462
hygiene
461
interaction of research, risk management and
497
see also regulation light microscopy
169
408
lighting
525
lipophilic toxins, ‘fast acting’
131
140
(LCMS)
145
153
domoic acid
150
DSP toxins
152
LCMS/MS
151
NSP toxins
151
PSP toxins
149
liquid chromatography mass spectrometry
153
This page has been reformatted by Knovel to provide easier navigation.
153
Index Terms Listeria monocytogenes
Links 252
crab
555
crawfish
555
disinfectants
560
equipment
559
heat resistance
555
lobster
557
product handlers
559
shrimp
558
Litopenaeus stylirostris
388
MCE experiment
394
super-intensive culture systems
391
TSV- and IHHNV-resistant
390
Litopenaeus vannamei
554
385
414
development as dominant species in the Americas
388
selective breeding for disease resistance SPF
437 398
429
IHHNV-resistant
390
402
USMSPF program
398
lobsters Listeria monocytogenes
544 557
location depuration plant
524
site selection for pond culture
340
losses, economic
279
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
M macrocyclic spiroimines
131
maintenance
531
male-specific coliphage (MSC)
13
Malthus impedance technique
460
140
153
19
marine biotoxin management plan (MBMP)
163
164
391
394
Marine Culture Enterprises (MCE) project Marteilia refringens
276
Marteilia sydneyi
276
mass (individual) selection
430
master sanitation programme
299
440
maximum zero-hour level of faecal coliforms
523
mechanical cleaning
327
Mediterranean diet
471
Memorandum of Understanding (MOU)
520
Mengo virus mercury methylmercury
94 235 235
metal complex species
237
metallothionein-like proteins (MTLP)
233
471
553
239
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
metals antifouling coatings
327
bivalve molluscs
228
concentrations
229
detection, management and risk assessment
241
exposure routes and application of kinetic model
234
future trends
242
internal speciation
233
safety standards
240
uptake and transfer
236
crustaceans
552
553
methylmercury
235
471
3
248
microbial contamination crustaceans preservation
553
554 549
effects on international shellfish industry
10
future trends
27
impacts on human health incidence major microbial contaminants microbial indicators
262
7 15 4 249
pathogens associated with handling, processing and distribution management
257 261
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
microbial contamination (Cont.) pollution-associated pathogens
249
management
258
risk management
478
post-harvest processing
479
493
488
sources of contamination and their identification
20
see also faecal-borne bacteria; vibrios; viruses microbial source tracking (MST)
4
microbiological analysis
306
microcell parasites
272
microcystins
22
51
143
Mikrocytos mackini
272
273
microsatellites
431
microscopy
169
MIST Alert lateral flow assay
147
mitigation strategies
176
408
HABs see harmful algal blooms modelling algal toxin accumulation see toxin accumulation modelling dynamics of HABs
169
managing viral contamination
117
This page has been reformatted by Knovel to provide easier navigation.
497
Index Terms molecular markers
Links 431
molecular methods diagnosis of shellfish diseases
279
shrimp disease
408
toxin-producing algae
67
viruses
89
improving detection
409
410
119
93
MONCOZE project
170
monitoring
357
464
47
60
61
61
162
algal toxins costs chemical contaminants
66 241
cooperation and shared resources between regulators and industry faecal-derived microorganisms
501 479
see also faecal indicators HABs procedures and HACCP
297
risk assessment
478
vibrios
482
viral contamination
108
306
479
conditions responsible for contamination
112
future trends
120
improving risk management
118
pollution source identification
110
strategies for reducing contamination
114
This page has been reformatted by Knovel to provide easier navigation.
220
Index Terms
Links
monodon baculovirus (MBV)
360
361
Morlaix, Bay of
115
116
most probable number (MPN) method
459
Mourilyan virus (MoV)
360
mouse bioassay MSX disease
60
143
145
149
296
478
274
MSX-resistant oyster breeding programme
281
multicompartment models
211
multiple antibody ELISA
143
multiple primer sets
91
multi-toxin screening tools
153
Mussel Watch Program
228
mussels
456
biofouling
324
modelling toxin accumulation
202
US production
295
207
N National Advisory Committee on Microbiological Criteria for Foods (NACMCF)
297
National Shellfish Sanitation Program (NSSP)
14
65
510
519
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
National Status and Trends Program
228
229
44
45
57
47
48
131
134
5
84
87
109
254
456
91
94
naturally occurring pathogens see vibrios neurotoxic shellfish poisoning (NSP) toxins analytical techniques
150
New Delhi hepatitis outbreak
85
New Zealand Marine Biotoxin Management Plan nitrofurans noroviruses (Norwalk-like viruses)
detection
535 167 553
monitoring viral contamination
108
risk management
499
nucleic acid extraction
90
nucleic acid sequence-based amplification (NASBA)
94
nursery ponds
457
375
O OIE
385 listed crustacean diseases
387
registry of assays
411
statistical table for sample size
412
okadaic acid (OA)
406
49
57
163
217
131
This page has been reformatted by Knovel to provide easier navigation.
136
Index Terms
Links
omega-6 fatty acids
468
one-compartment models
208
operational oceanography
170
ORCAS
170
209
211
organic contaminants bivalve molluscs
228
exposure
236
uptake and transfer
238
crustaceans
552
organic fertilisation
342
553
oxygen dissolved in seawater
512
uptake and metal uptake
238
oysters
6
biofouling
322
324
298
311
HACCP programme for frozen raw oysters parasites microcell parasites
272 272
production in US
295
public confidence
455
ozone chemical control of HABs
189
depuration
509
516
packaging crustaceans
549
effects of biofouling
325
This page has been reformatted by Knovel to provide easier navigation.
212
Index Terms
Links
ozone (Cont.) molluscan shellfish
568
accepting shellfish shipments
574
depurated shellfish
533
formats and materials
569
labelling and tagging
571
product size standards
574
product specification
568
P PAH
236
paralytic shellfish poisoning (PSP)
44
45
163
166
accumulation of toxin
53
effect on humans
57
heat treatment
490
risk matrix
496
129
132
toxin accumulation modelling biotransformations
216
king scallops
213
mussels
202
207
47
130
toxins analytical techniques
218
145
blooms producing
62
Paramoeba pemaquidensis
325
paramyxean parasites
276
This page has been reformatted by Knovel to provide easier navigation.
162
Index Terms
Links
parasites biological control of HABs
185
on shellfish
271
particulate uptake of metals
235
Parvilucifera infectans
186
passive surveillance
411
pasteurisation
261
pathogen-assisted molecular patterns (PAMPs)
364
pattern recognition receptors (PPRs)
364
pectenotoxins (PTX)
Penaeus monodon
277
549
561
49
55
58
139
153
385
389
429
Penaeus vannamei see Litopenaeus vannamei periostracum
319
Perkinsus spp.
277
Perkinsus marinus
277
Perkinsus olseni
277
permits
520
persistent organic pollutants (POPs)
552
553
pesticides
552
553
Pfiesteria spp.
46
51
Pfiesteria toxin (PfTX)
51
pH
346
phagocytosis
364
phenotypic variance
427
This page has been reformatted by Knovel to provide easier navigation.
131
Index Terms
Links
phlorotannins
188
phosphate clays
180
phycotoxins see algal toxins phytoplankton monitoring
61
picornaviruses
85
pinnatoxins
131
140
plant design, depuration
524
529
plant growth reduction
347
plant sterols
466
plastic trays
570
pleiotropic genes
440
Poison Cove
456
poliomyelitis
85
pollution-associated pathogens management
467
249 258
see also coliforms, faecal; Escherichia coli (E. coli); faecal-borne bacteria pollution source identification
20
110
polychlorinated biphenyls (PCBs)
236
553
Polydora
319
322
330
89
93
457
real-time PCR
93
119
RT-PCR
90
91
polymerase chain reaction (PCR)
pond culture of crustaceans
93
339
collection during harvest
358
fertilisation and semi-intensive systems
342
This page has been reformatted by Knovel to provide easier navigation.
457
Index Terms
Links
pond culture of crustaceans (Cont.) formulated feeds and intensive systems
344
harvest and post-harvest treatment
349
safety and quality standards
350
site selection
340
substrate
355
water quality management
356
water source and conservation
341
positive price elasticity post-harvest processing (PHP) crustaceans
470 27
260
349
542
collection and transport
350
contaminants
552
modification of taste
349
monitoring, verification and record keeping
357
packaging and preservation
549
regulations and temperature management
352
safety and quality standards
350
sanitation and sanitary conditions
357
slaughter
351
storage and distribution
352
traceability and food security plan
356
HACCP and frozen raw oysters
300
311
labelling
572
573
pathogens associated with
257
259
management
543
261
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
post-harvest processing (PHP) (Cont.) risk management
488
reduction of biotoxins
490
reduction of microbial pathogens
488
postlarvae (PLs) (‘seed’) wild seed
386 387
pouches, plastic
570
predation
321
control of biofouling
413
426
330
preservation methods
549
prevention strategies
176
primary septicaemia
257
primary treatment
114
191
115
processing see post-harvest processing product size standards
574
product specification
568
prorocentrolides
131
140
Prorocentrum micans
46
51
Prorocentrum minimum
46
51
protein phosphatase inhibition assay (PPIA) protozoa
59
152 7
pteriatoxins
131
public confidence
455
dietary and health advice
465
future trends
470
hygiene legislation and
461
140
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
public confidence (Cont.) perception of health benefits and risks
469
risk aversion
471
in shellfish and risk
455
public health controls public health impacts of HABs pumping rate
476 66 238
Q quality assurance
531
quality standards
350
quaternary ammonium compounds (QACs)
560
QX disease
276
QX-resistant oyster breeding programme
282
radiolabelled toxins
144
203
rainfall
112
113
Ralston Purina
388
R
rapid-cycling real-time PCR
93
raw consumption of shellfish
86
real-time PCR
93
119
recall/traceability programme
298
299
receiving record
312
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
receptor binding assays (RBAs)
144
recirculation depuration plants
510
recombinant tracers
95
147
357
depuration
534
HACCP
297
307
66
89
red crawfish Red Tide/Toxic Algae Project
311
65 458
reductions
216
regional disease surveillance programmes
284 88
algal toxins
310
547
REDRISK Project
regulation
150
97
record keeping
recreational activities
149
218
352
461
258
462
64
harmonization and international policies mandatory harmful algal monitoring
64 166
application and improving risk management strategies growing water quality
118 11
HACCP
296
limitations of regulatory approach
463
metals
240
oyster production
296
and public confidence
461
This page has been reformatted by Knovel to provide easier navigation.
526
Index Terms
Links
regulation (Cont.) risk management official and industry roles
485
487
shared resources and cooperation by regulators and industry relaying
500 27
algal toxins and
63
management of vibrios
260
for natural depuration
538
operating procedures
537
viral contaminants
497
‘residual risk’
459
response to selection
428
Restricted areas/waters
rigid plastic containers
489
95
research
reverse transcription PCR (RT-PCR)
537
461
464
469
258
462
526
528
90
91
93
457
570
571
risk limitations of regulatory approach modelling public confidence and shellfish
464 28 455
risk assessment improved
119
486
interaction of research, legislation and risk management
497
metals
241
supporting risk management
476
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
risk aversion
471
risk management
474
future trends
486
active management
488
onus on producers
492
post-harvest processing
488
improved application to microbiological and toxin problems improving risk management strategies
479 118
interaction between public health controls and industry
476
interaction of research, legislation and
497
need for
476
optimising
476
regulators’ and industry roles
485
487
shared resources and cooperation by regulators and industry risk profiling supporting risk management rivers RNA genomes RNA interference (RNAi)
500 479
112 95 365 365
response to dsRNA in shrimp
367
rotaviruses
481
476
as a defence against viral infection
RNA silencing
480
375
365 84
91
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
S salinity
16
and depuration
512
taste modification
349
salmon
362
Salmonella spp.
250
saltwater wells
514
513
374
sampling biotoxin risk management
221
monitoring HABs
168
programme and depuration
531
482
sample size and surveillance for SPF shrimp sanitary conditions/facilities Sanitary and Phytosanitary Agreement
412 357 64
sanitary surveys
259
sanitation
357
audit form
478
315
sanitation standard operating procedures (SSOPs)
298
352
sapoviruses
84
SARS
97
satellite monitoring of HABs
68
169
saxitoxins (STX)
47
131
132
This page has been reformatted by Knovel to provide easier navigation.
163
Index Terms scallops
Links 182
biofouling
318
324
320
321
322
shucking to reduce ASP
485
487
491
toxin accumulation modelling
213
by sponges
scheduled controlled purification process (SCPP) scombrotoxin Sea Fish Industry Authority (SEAFISH)
522 67 464
465
seasonality algal toxins risk management microbial contamination seaweed wrack
497 484 17
113
26
secondary treatment
114
security plan
356
selection differential
429
selection intensity
429
115
selective breeding basic concepts
427
reducing disease in molluscan shellfish
281
shrimp
425
inbreeding
431
programs for shrimp
429
selection for disease resistance
437
selection for growth
432
442
This page has been reformatted by Knovel to provide easier navigation.
Index Terms selective evisceration
Links 63
selenium
469
self-regulation
464
semi-intensive systems
342
seston depletion effect
320
sewage
456
viral contamination limiting faecal input
110
112
114
transmission of viruses
85
sewage treatment plants (STPs)
114
sexual growth dimorphism
433
shared resources
500
shell quality
322
shellfish receiving record
312
shellfish shipments, accepting/rejecting
574
shellfish tissue standards
111
11
Shigella spp.
252
shipments, accepting/rejecting
574
shrimp global production of farmed shrimp
426
Listeria monocytogenes
558
pond culture see pond culture of crustaceans selective breeding
425
SPF see specific pathogen-free (SPF) shrimp
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
shrimp virus-binding proteins
375
shrimp viruses
359
developing vaccines to manage viral disease
369
emergence and their impact
359
future trends
375
invertebrate immune system
364
adaptive
385
369
using the RNA interface to target viruses
365
using vaccines in health management strategies
372
shucking king scallops
485
487
sib selection
430
441
silver
230
232
simulation modelling site selection
491
234
25 340
slaughter chill killing
351
and cooking
543
small interfering RNA (siRNA)
365
sodium hypochlorite
187
sophorolipid
180
South Korea
178
Southeast Asia
535
Spain
536
SPATT
170
367
188
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
species-specific probes
262
specific pathogen-free (SPF) shrimp
384
429
444
adaptation of SPF concept to domesticated shrimp stocks
392
freedom from specific diseases
392
MCE experiment
391
394
398
405
biosecurity and the culture of wild seed/bloodstock
413
413
through environmental control and best management practices disease diagnosis and surveillance diagnosis vs surveillance historical perspective
414 407 411 386
International Principles for Responsible Shrimp Farming
397
lessons learned in developing SPF shrimp stocks
402
listed shrimp diseases/pathogens
405
maintenance of SPF status
405
USMSFP
393
specifically listed diseases (SLDs) spirolides sponges biofouling of scallops SPR-43
398
405 50
131
318
319
320
321
140
322
390
This page has been reformatted by Knovel to provide easier navigation.
153
Index Terms SSO disease
Links 275
standards harvesting waters
11
product size
574
safety limits for metals
240
258
462
523
531
safety and quality standards for pondcultured crustaceans virus
350 97
standpipes
348
Staphylococcus aureus
258
State Shellfish Control Agency (SSCA)
519
statins
469
steaming
543
sterols
466
Stockton Springs, Maine
467
24
storage feed and feedstuffs
345
safety and quality standards
352
storm events stormwater runoff stress
112
113
456
239
242
373
375
18 374
inbreeding and
443
in shrimp
414
Stylochus frontalis
322
subcellular pools
233
substrate
345
subunit vaccination
370
This page has been reformatted by Knovel to provide easier navigation.
526
Index Terms
Links
Super Shrimp
390
supervision
534
surface texture
319
surveillance
405
diagnosis and diagnosis vs surveillance/screening survival capacity, microbial
407 411 19
sustainable development
121
swimming (blue) crabs
544
T tagging
528
571
tanks, depuration
529
530
targeted surveillance
411
412
taste, modification of
349
Taura syndrome virus (TSV)
360
selective breeding for TSV resistance
437
TSV-resistant Litopenaeus stylirostris
390
361
385
temperature control of vibrios
15
management
353
seawater temperature and depuration
511
tertiary treatments
114
tetrodotoxin (TTX)
50
Texas A&M University pond studies
16
513
59
67
389
This page has been reformatted by Knovel to provide easier navigation.
388
Index Terms therapeutic agents, feed-based thin layer chromatography (TLC) time-temperature matrices tissue-based standards
Links 345 61 260 11
tissue samples, examination of
168
total suspended solids
514
totes
568
tourism
66
toxin absorption efficiency
206
toxin accumulation modelling
200
89
applications for improving safety and quality
220
depuration
202
future trends
221
204
206
221
208
historical use and development of models
204
models of biotransformation
203
215
models with external variables
211
212
multicompartment models
211
processes
206
rationale
201
toxin incorporation
204
206
298
299
see also algal toxins traceability
353
534
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356
Index Terms
Links
trade effects of shellfish disease viral contamination and
280 88
training programme
299
transgenic animals
282
368
transport
350
352
528
trays
570
treated water
112
triggers
164
trophic transfer
235
turbidity
514
Turkey
536
two-compartment models
208
211
212
typhoid fever
250
515
518
119
470
65
297
U ultrafiltered organic carbon (UOC)
238
ultra-high-pressure (UHP) treatment
471
ultrasonic devices
189
ultraviolet irradiation
510
United Kingdom (UK)
470
depuration
535
hygiene legislation
461
United States of America (USA)
88
clay flocculation
180
Clean Water Act
118
Food and Drug Administration
14
This page has been reformatted by Knovel to provide easier navigation.
309
Index Terms
Links
United States of America (USA) (Cont.) HACCP regulations ISSC
296 65
labelling regulations
571
NACMCF
297
NSSP
296
14
65
510
519
regulation of metals
240
regulation of shellfish growing water quality
462
shellfish-producing regions
295
296
526
United States Marine Shrimp Farming Consortium (USMSFC) SPF programme untreated water, used in cooling
393 398 475
uptake metals and transfer toxins urchins
234 236 52 330
V vaccines
369
benefits to aquaculture
362
developing to manage viral disease in shrimp
369
emergence of viral vaccines
370
history of shrimp vaccination
370
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478
Index Terms
Links
vaccines (Cont.) using as part of health management strategies validation procedures
372 309
VBNC state
19
venerupin shellfish poisoning (VSP)
51
ventilation
525
verification
357
HACCP Vibrio Management Committee
297
307
309
14
Vibrio cholerae
254
Vibrio parahaemolyticus
255
478
Vibrio vulnificus
257
260
262
469
4
5
109
254
259
296
459
8
9
vibrios
and disease outbreaks environmental effects on contamination
15
HACCP and processing of frozen raw oysters
300
311
management of
14
259
262
478
482
497
373
375
risk management post-harvest processing
489
Vietnamese HAB monitoring programme
168
viral hepatitis
254
viral vaccines
370
virus concentration methods
90
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Index Terms
Links
virus extraction methods viruses
biological control of HABs
90 4
5
252
456
83
183
depuration
95
110
120
detection
12
89
120
improving using molecular-based methods
93
disease outbreaks
7
87
effects of contamination on international shellfish industry
88
environmental effects on contamination
17
future trends in virus studies
112
120 96
human enteric viruses and their fate in the environment
83
identifying sources of pollution
110
risk management
478
improving risk management strategies
118
post-harvest processing
488
shellfish-borne transmission of virus infections
479
493
85
shrimp viruses see shrimp viruses sources strategies for reducing contamination visible implant elastomer (VIE) tags
108
22
24
114 431
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499
Index Terms
Links
W walls
525
warning system
120
water depuration
529
flow rate
530
supply
525
source and conservation for pond culture of crustaceans
341
water quality and depuration
511
environmental legislation for shellfish growing water quality
11
management in pond culture
346
MPN method
459
watershed ponds
341
waxed cartons
570
waxy coatings
327
white crawfish
547
White spot syndrome virus (WSSV)
360
selective breeding for WSSV resistance
462
526
385
388
571
361
437
viral vaccines
370
wild animals and birds
25
wild-caught shrimp
426
wild postlarvae (‘wild seed’)
387
biosecurity and the culture of
258
373
426
413
This page has been reformatted by Knovel to provide easier navigation.
Index Terms
Links
winter mortality
273
withering syndrome
276
within-family selection
430
World Health Organization (WHO)
497
World Trade Organization (WTO)
64
X Xenohaliotis californiensis (withering syndrome)
276
yellow clay dispersal
178
Yersinia enterocolitica
257
Y
yessotoxins (YTX)
182
49
58
141
151
230
231
131
140
233
469
Z zinc uptake
235
This page has been reformatted by Knovel to provide easier navigation.
