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Responsible Fisheries in the Marine Ecosystem
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Responsible Fisheries in the Marine Ecosystem
Edited by
M. Sinclair Bedford Institute of Oceanography Dartmouth Nova Scotia Canada and
G. Valdimarsson Fishery Industries Division Food and Agriculture Organization of the United Nations Rome Italy
Published by
Food and Agriculture Organization of the United Nations and
CABI Publishing
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© 2003 by FAO. All rights reserved. No part of this publication may be reproduced in any form or by any means, electronically, mechanically, by photocopying, recording or otherwise, without the prior permission of the copyright holders. A catalogue record for this book is available from the British Library, London, UK. Published jointly by: CABI Publishing CAB International Wallingford Oxon OX10 8DE UK Tel: +44 (0) 1491 832111 Fax: +44 (0) 1491 833508 E-mail:
[email protected] Website: www.cabi-publishing.org
CABI Publishing 44 Brattle Street 4th Floor Cambridge, MA 02138 USA Tel: +1 617 395 4056 Fax: +1 617 354 6875 E-mail:
[email protected]
Food and Agriculture Organization of the United Nations (FAO) Viale delle Terme di Caracalla, 00100 Rome, Italy Tel: +39 06 57051 Fax: +39 06 57053152 ISBN 0 85199 633 7 (CABI) ISBN 925 104767 7 (FAO) The designations employed and the presentation of material in this publication do not imply the expression of any opinion whatsoever on the part of the Food and Agriculture Organization of the United Nations concerning the legal status of any country, territory, city or area or of its authorities, or concerning the delimitation of its frontiers or boundaries. The designations ‘developed’ and ‘developing’ economies are intended for statistical convenience and do not necessarily express a judgement about the stage reached by a particular country, territory or area in the development process. The views expressed herein are those of the authors and do not necessarily represent those of the Food and Agriculture Organization of the United Nations. Library of Congress Cataloging-in-Publication Data Responsible fisheries in the marine ecosystem/edited by M. Sinclair and G. Valdimarsson. p. cm. Includes bibliographical references (p. ). ISBN 0-85199-633-7 (alk. paper) 1. Sustainable fisheries--Congresses. 2. Fishery management--Congresses. 3. Marine ecology--Congresses. I. Sinclair, Michael, 1944- II. Valdimarsson, G. (Grimur) III. Food and Agriculture Organization of the United Nations. SH329.S87 R47 2002 333.95´6--dc21 2002011121 Typeset by AMA DataSet Ltd, UK. Printed and bound in the UK by Biddles Ltd, Guildford and King’s Lynn.
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Contents
Contributors
ix
Foreword Director-General of FAO
xi
Foreword Minister of Fisheries, Iceland Preface Acknowledgements
xiii
xv xvii
PART I: INTRODUCTORY REVIEWS
1
1
Global Overview of Marine Fisheries Serge M. Garcia and Ignacio de Leiva Moreno
2
Obligations to Protect Marine Ecosystems under International Conventions and Other Legal Instruments Transform Aqorau
25
Incorporating Ecosystem Considerations into Fisheries Management: Large-scale Industry Perspectives Bernt O. Bodal
41
Small-scale Fisheries Perspectives on an Ecosystem-based Approach to Fisheries Management Sebastian Mathew
47
An Environmentalist’s Perspective on Responsible Fisheries: the Need for Holistic Approaches Tundi Agardy
65
3
4
5
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Contents
vi
PART II: DYNAMICS OF MARINE ECOSYSTEMS
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Ecological Geography as a Framework for a Transition Toward Responsible Fishing Daniel Pauly, Reg Watson and Villy Christensen
87
7
The Functioning of Marine Ecosystems: a Fisheries Perspective Philippe Cury, Lynne Shannon and Yunne-Jai Shin
103
8
Food Webs in the Ocean: Who Eats Whom and How Much? Andrew W. Trites
125
9
Regional Assessments of Prey Consumption and Competition by Marine Cetaceans in the World Tsutomu Tamura
10
Multi-species and Ecosystem Models in a Management Context Gunnar Stefansson
143
171
PART III: THE ROLE OF MAN IN MARINE ECOSYSTEMS
11
Multiple Uses of Marine Ecosystems Andrew A. Rosenberg
189
12
Impacts of Fishing Gear on Marine Benthic Habitats Michel J. Kaiser, Jeremy S. Collie, Stephen J. Hall, Simon Jennings and Ian R. Poiner
197
13
The Magnitude and Impact of By-catch Mortality by Fishing Gear Robin Cook
219
14
The Effects of Fishing on Species and Genetic Diversity Ellen L. Kenchington
235
15
The Effects of Fishing on Non-target Species and Ecosystem Structure and Function Henrik Gislason
16
Anthropogenically Induced Changes in the Environment: Effect on Fisheries Katherine Richardson
255
275
PART IV: INCORPORATING ECOSYSTEM CONSIDERATIONS IN FISHERIES MANAGEMENT
17
18
19
The Performance of Fisheries Management Systems and the Ecosystem Challenge Jon G. Sutinen and Mark Soboil
291
The Role of Harvest Control Laws, Risk and Uncertainty and the Precautionary Approach in Ecosystem-based Management Douglas S. Butterworth and A.E. Punt
311
Modifying Fishing Gear to Achieve Ecosystem Objectives John W. Valdemarsen and Petri Suuronen
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Contents
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Incorporating Ecosystem Objectives into Management of Sustainable Marine Fisheries, Including ‘Best Practice’ Reference Points and Use of Marine Protected Areas Keith Sainsbury and Ussif Rashid Sumaila
vii
343
21
Governance for Responsible Fisheries: an Ecosystem Approach Michael P. Sissenwine and Pamela M. Mace
363
22
Towards Ecosystem-based Fisheries Management FAO
393
Appendix 1: Industry Perspectives
405
Appendix 2: The Reykjavik Declaration on Responsible Fisheries in the Marine Ecosystem
409
Index
413
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Contributors
T. Agardy, Sound Seas, 6620 Broad Street, Bethesda, MD 20816, USA. T. Aqorau, South Pacific Forum Fisheries Agency, PO Box 629, Honiara, Solomon Islands. B.O. Bodal, American Seafoods Group LCC, 2025 1st Ave., Suite 1200, Seattle, WA 98121, USA. D.S. Butterworth, Department of Mathematics and Applied Mathematics, University of Cape Town, Rondebosch 7701, South Africa. V. Christensen, Fisheries Centre, 2204 Main Mall, University of British Columbia, Vancouver, British Columbia V6T 1Z4, Canada. J.S. Collie, Graduate School of Oceanography, University of Rhode Island, Narragansett, RI 02882, USA. R. Cook, FRS Marine Laboratory, PO Box 101, Victoria Road, Aberdeen AB11 9DB, UK. P. Cury, IRD Research Associate at UCT and MCM, Oceanography Department, University of Cape Town, Rondebosch 7701, South Africa. S.M. Garcia, Fishery Resources Division, Food and Agriculture Organization (FAO) of the United Nations, Viale delle Terme di Caracalla, 00100 Rome, Italy. H. Gislason, University of Copenhagen, c/o Danish Institute for Fisheries Research, Charlottenlund Castle, DK-2020 Charlottenlund, Denmark. S.J. Hall, Australian Institute for Marine Science, PMB 3, Townsville MC, Queensland 4810, Australia. S. Jennings, The Centre for Environment, Fisheries and Aquaculture Science, Pakefield Road, Lowestoft NR33 0HT, UK. M.J. Kaiser, Marine Benthic Ecology, School of Ocean Sciences, University of Wales-Bangor, Menai Bridge, Anglesey LL59 5EY, UK. E.L. Kenchington, Centre for Marine Biodiversity, Bedford Institute of Oceanography, PO Box 1006, Dartmouth, Nova Scotia B2Y 4A2, Canada. I. de Leiva Moreno, Fishery Resources Division, Food and Agriculture Organization (FAO) of the United Nations, Viale delle Terme di Caracalla, 00100 Rome, Italy. P.M. Mace, Northeast Fisheries Science Center, National Marine Fisheries Service, 166 Water Street, Woods Hole, MA 02543, USA. S. Mathew, International Collective in Support of Fishworkers (ICSF), 27 College Road, Chennai 600 006, India. D. Pauly, Fisheries Centre, 2204 Main Mall, University of British Columbia, Vancouver, British Columbia V6T 1Z4, Canada. I.R. Poiner, CSIRO Division of Marine Research, PO Box 120, Cleveland, Queensland 4163, Australia. ix
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Contributors
A.E. Punt, School of Aquatic and Fishery Sciences, Box 355020, University of Washington, Seattle, WA 98195-5020, USA. K. Richardson, Institute of Biological Sciences, Department of Marine Ecology, University of Aarhus, Finlandsgade 14, DK-8200 Aarhus, Denmark. A.A. Rosenberg, College of Life Sciences and Agriculture, University of New Hampshire, Hampton, NH 03824, USA. K. Sainsbury, CSIRO, Division of Marine Research, Castray Esplanade, Hobart, Tasmania 7001, Australia. L. Shannon, Marine and Coastal Management, Private Bag X2, 8012 Rogge Bay, Cape Town, South Africa. Y.-J. Shin, IRD Centre de Recherche Halieutique Méditerranéenne et Tropicale, Avenue Jean Monnet, BP 171, 34203 Sète Cedex, France. M.P. Sissenwine, Northeast Fisheries Science Center, National Marine Fisheries Service, 166 Water Street, Woods Hole, MA 02543, USA. M. Soboil, Department of Environmental and Natural Resource Economics, University of Rhode Island, Coastal Institute, Kingston, RI 02881, USA. G. Stefansson, University of Iceland, Science Institute, Dunhaga 7, 101 Reykjavik, Iceland. U.R. Sumaila, University of British Columbia, Fisheries Centre, 2204 Main Mall, Vancouver, British Columbia, V6T 1Z4, Canada. J.G. Sutinen, Department of Environmental and Natural Resource Economics, University of Rhode Island, Coastal Institute, Kingston, RI 02881, USA. P. Suuronen, Finnish Game and Fisheries Research Institute, PO Box 6, FIN-00721 Helsinki, Finland. T. Tamura, Ecosystem Section, The Institute of Cetacean Research, Tokyo Suisan Building, 4–18 Toyomi-Cho, Chuo-Ku, Toyko 104-0055, Japan. A.W. Trites, Marine Mammal Research Unit, Fisheries Centre, University of British Columbia, Vancouver, British Columbia V6T 1Z4, Canada. J.W. Valdemarsen, Fishery Technology Service, Food and Agriculture Organization (FAO) of the United Nations, Viale delle Terme di Caracalla, 00100 Rome, Italy. R. Watson, Fisheries Centre, 2204 Main Mall, University of British Columbia, Vancouver, British Columbia V6T 1Z4, Canada.
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Foreword
In recent decades, ocean capture fisheries have been expanded successfully worldwide, in both developed and developing countries. The current catch from the oceans, some 80 million t annually, has, in the view of FAO, reached the practical maximum that we can expect to take from wild fish stocks. However, this does not mean that we cannot get more food from the oceans. We can make better use of the resources by allowing overfished stocks to recuperate, by reducing wastage and, in particular, by making renewed efforts towards their better management. This means new initiatives and a new focus in developing capacity for fisheries research and management, allocating fishing rights and improving monitoring, control and surveillance, not to mention improving catch reporting. Developing countries have been able not only to take full part in exploiting their fishery resources but also, in a very significant way, to take part in international fish trade. With increasing environmental awareness, civil society everywhere is taking a more active role in conservation issues and resource management. There is concern that the exploitation of wild stocks should do the least possible damage to the environment, and that we should manage individual stocks for optimum long-term yields. Fishing activities should be managed in a way that not only takes into account the stock being targeted but also respects the balance of the ecosystem. The Reykjavik Conference on Responsible Fisheries in the Marine Ecosystem was an effort towards better understanding of how ecosystem-sensitive fisheries can be achieved. This book is a significant contribution towards better appreciation of the complexity of the biological interactions in the oceans, and how they are affected by human activities. It also addresses the legal framework for marine fisheries, and guards against unsustainable use. It was clear during the conference that ecosystems are too complex to be managed easily. Nevertheless, human activities can be controlled, particularly if the incentive regime is structured carefully. The Reykjavik Declaration signals the direction in which to proceed – and FAO is committed to doing its share in that process. I wish to express my profound gratitude to the Government of Iceland for hosting this Conference, and to the Government of Norway for providing much required support. It demonstrates the commitment of these major fishery nations towards responsible use of marine resources. Dr Jacques Diouf Director-General Food and Agriculture Organization of the United Nations xi
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Foreword
It was a great pleasure and an honour for Iceland to have the opportunity of hosting the Conference on Responsible Fisheries in the Marine Ecosystem, which took place in Reykjavík in October 2001, and we would like to express our sincere thanks to FAO for their cooperation and collaboration in bringing this project to a successful conclusion. We would also like to thank the Norwegian authorities for their substantial financial support for the Conference. Conferences such as this serve a variety of purposes, but, once concluded, it is important to consider what is to be learnt from the proceedings. Accordingly, one could ask, what stands out as most significant regarding the Reykjavík Conference? One unique feature of the Conference was the extremely wide scope of the topics discussed. By this, I refer to the fact that it was not only an international gathering of politicians and administrators, but also had considerable time devoted to technical lectures by leading scientists in the various disciplines relevant to the ocean biosphere. All the scientists made a concerted effort to present complicated questions in a manner that rendered the subjects and arguments comprehensible to more than just an inner circle. This provided a valuable opportunity for those of us involved in politics and administration to become acquainted with the results of years – and even decades – of scientific endeavour. The lessons to be learnt from the conference are thus greater than might have been expected, precisely because its perspective was so wide. In my opinion, the most important point was that the approach that had been generally accepted for so long, namely that the status of individual commercial stocks should be considered from the point of view of the performance of the individual stocks, is far from sufficient. Nevertheless, we should not forget that this singlespecies approach was a great step forward from the belief that the ocean’s living resources were practically inexhaustible. As a result of the Reykjavík Conference, it is more evident than ever before that we need to consider the marine ecosystem in its totality when taking decisions on the utilization of individual marine organisms. When viewed in this respect, no creature is superior to another except in the natural order of the established ecological equilibrium. Thanks to the untiring work of scientists throughout the world, we are steadily increasing our knowledge in this area, and with it comes the understanding that we must respect the ecological equilibrium while judiciously utilizing the ocean’s entire spectrum of natural resources. In closing, I would like to quote a traditional Icelandic adage, which succinctly sums up the understanding of our forebears in Iceland of the importance of using land with respect, i.e. to live off the land without mistreating it in any way. The saying is Land is to use and not abuse. This view is easily extended to include the ocean, and is clearly in accord with the policy of utilizing the ocean’s resources without abusing them through overexploitation, resource depletion or xiii
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Foreword
disturbance of the inherent balance of nature. Remember – as we Icelanders are so frequently forced to recall – we cannot conquer nature, but must instead learn to live in harmony with and respect Nature’s laws. Arni M. Mathiesen Minister of Fisheries, Iceland
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Preface
This book presents the edited papers from the Conference on Responsible Fisheries in the Marine Ecosystem, which was held from 1–4 October 2001 in Reykjavik, Iceland. The Conference was organized jointly by the Government of Iceland and the Food and Agriculture Organization of the United Nations (FAO), with the co-sponsorship of the Government of Norway. The objectives of the Conference were to: (i) review current knowledge on marine ecosystems of relevance to fisheries management; (ii) identify means by which ecosystem considerations can be included in fisheries management; and (iii) foresee future challenges and outline the relevant strategies to cope with them. In preparing the Conference, a Scientific Committee selected the topics to be addressed and invited experts in the respective scientific fields. Members of the Scientific Committee also undertook a peer review of all presentations. The Conference also heard some views of representatives of the fishery industry. These are summarized in Appendix 1. The Conference unanimously adopted a resolution – the Reykjavik Declaration – on how to work towards ecosystem-based fisheries management. The Declaration is presented in Appendix 2. This Declaration, having been endorsed by the FAO Conference in November 2001, has been submitted for consideration of the World Summit on Sustainable Development, to be held in Johannesburg in September 2002.
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Acknowledgements
The success of the Reykjavik Conference is a reflection of the considerable effort made by numerous persons in the preparation, implementation and follow-up phases. Sincere thanks are extended to all those who helped make the Reykjavik Conference such a success. In particular, those noted below worked hard to ensure success. The Conference Steering Committee comprised Mr Thorsteinn Geirsson, Secretary-General, Ministry of Fisheries, Iceland; Mr Sverrir H. Gunnlaugsson, Secretary-General, Ministry for Foreign Affairs, Iceland; Mr Johán Williams, Director General, Ministry of Fisheries, Norway; Dr Zbigniew S. Karnicki, Director, Fishery Policy and Planning Division, FAO; Dr Jorge Csirke, Chief, Marine Resources Service, FAO. The Conference Scientific Committee was made up of Dr Michael Sinclair, Director, Bedford Institute of Oceanography, Canada (Chair); Dr Ragnar Árnason, Professor of Economics, University of Iceland, Iceland; Dr Jóhann Sigurjónsson, Director-General, Marine Research Institute, Iceland; Dr Hein Rune Skjoldal, Professor of Marine Ecology, Institute of Marine Research, Norway; Dr Zbigniew S. Karnicki, Director, Fishery Policy and Planning Division, FAO; Dr Jorge Csirke, Chief, Marine Resources Service, FAO; supported by the Conference Secretary, Dr Grimur Valdimarsson, Director, Fishery Industries Division, FAO. The Reykjavik Executive Committee consisted of Mr Thorsteinn Geirsson, Secretary-General, Ministry of Fisheries, Iceland; Mr Sverrir H. Gunnlaugsson, Secretary-General, Ministry for Foreign Affairs, Iceland; Mr Gudmundur Helgason, Secretary-General, Ministry of Agriculture, Iceland; Mr Fridrik Pálsson, Chairman, SIF Group and Seafood Industry Consultant, Iceland; Dr Björn Sigurbjörnsson, former Secretary-General, Ministry of Agriculture, Iceland. Day-to-day management of Conference preparation and implementation activities in Iceland was in the hands of the Reykjavik Executive Officer, Dr Alda Möller, Seafood Industry Consultant, Iceland.
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Acknowledgements
In Rome, the FAO Secretariat comprised Dr Grímur Valdimarsson, Director, Fishery Industries Division (Conference Secretary); Ms Janet Webb (Meetings Officer); Ms Joanne Antonelli, Ms Nadia Brusadelli, Ms Wilma van Kessel and Ms Barbara Vermeil (secretarial and administrative assistance); and they also formed part of the FAO team in Iceland during the conference. The Secretariat benefited from the collaboration of Dr Wolfgang Krone, formerly with FAO as Director of the Fishery Industries Division and Assistant Director General (a.i.) of the Fisheries Department. The initial editing of all the Conference papers before the Conference and subsequent preparation for the publisher was by Mr Thorgeir Lawrence.
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1
Global Overview of Marine Fisheries Serge M. Garcia and Ignacio de Leiva Moreno Fishery Resources Division, FAO, Rome, Italy
Abstract The latest FAO review of the state of the world’s marine fishery resources confirms that, from the information available, about 50% of global resources are fully exploited, 25% are overexploited and about 25% could apparently support higher exploitation rates. While the proportion of overfished stocks seems to have increased much less than in the past, reaching an asymptote in the 1990s, the historical trend towards more overfishing observed since the early 1970s has not yet reversed. The fishery sector expanded greatly during the 1970s and 1980s. The size of the industrial fleet has now stabilized and is evolving slowly, with a possible downwards trend. However, the number of people involved seems to be still growing, probably in the small-scale sector. Overall, the fisheries contribution to economic development and food security is very significant, but overcapacity seems pervasive and is jeopardizing the economic and social performance of the sector, as well as its sustainability in a number of cases. The ocean ecosystem, under high fisheries pressure, is suffering from pollution from sea-based and (mostly) land-based activities and coastal degradation. Critical coastal habitats, such as sea-grass beds, coral reefs and mangroves, as well as estuaries and lagoons, are strongly affected by coastal developments including aquaculture and pollution. Global climate change, a particularly threatening manifestation of pollution, is affecting critical resources (e.g. coral bleaching) through excessively high temperatures and high UV radiations. The institutional basis for ecosystem-based governance of fisheries is building up rapidly following the entry into force (in 1994) of the 1982 Convention on the Law of the Sea, the adoption of the UN Fish Stocks Agreement and the FAO Code of Conduct (1995), and a number of non-fisheries instruments with significant implications for fisheries. This basis has already been adopted at the highest government levels, but implementation at the lower levels (national and local) is still slow or absent in many cases. Insufficient capacity is a problem in many developing countries. Socio-economic and political short-term costs – and resultant political reluctance – are a problem everywhere. Coordination, participation and transparency represent real opportunities for positive change as well as major difficulties.
Introduction After 50 years of particularly rapid geographical expansion and technical advances, and a several-fold increase in annual harvest, marine fisheries are at a crossroads. They have evolved significantly during the last 50 years, facing the hard challenges of
sustainability, resources discovery, difficult working conditions, competition, changing demand, unpredictable ecosystems and uncertain political, social and economic environments. The process of elaboration of a world charter for fisheries, which started in 1958 (with UNCLOS I) led, in 1982, to the adoption of the UN Convention on the Law
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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S.M. Garcia and I. de Leiva Moreno
of the Sea, which entered into force in 1994. Social and institutional awareness of human impact on the ecosystem has been developing since the early 1970s with the work of the World Conference on Human Environment (Stockholm, 1972) and the UN Conference for Environment and Development (UNCED) (Rio de Janeiro, 1992), followed by the recent work of the UN Commission on Sustainable Development (CSD). Nearly 10 years after UNCED, and more than half a century after the first technical meeting on fisheries of the FAO, in 1946, and a few months before the UNCED + 10 Summit in Johannesburg, the FAO Iceland Conference on Responsible Fisheries in the Marine Ecosystem offers the very first opportunity to discuss, at world level, the cross-implications for the fishery sector of the two processes mentioned above. This chapter presents a global overview of the historical trends and of the present situation of the sector, as a contribution to the Reykjavik Conference debate. It is intended by design to be mainly descriptive and not prescriptive. It offers, first, a global picture of the state of the biological resources, followed by a brief review of the fishery system (fleet, people, technology, production and trade, and its contribution to food security) and the evolution of governance, with its approaches, performance and trend towards ecosystembased fisheries management (EBFM). The conclusion stresses inter alia the quality of the data available, the validity and potential biases in the paper’s conclusions, and the need for a significant improvement in the monitoring of the sector.
The State of the Resources1 Global situation Following its first global review of marine fish stocks (Gulland, 1970, 1971), the FAO Fisheries Department has been monitoring the state of these stocks, describing trends in an intermittent publication, The State of World
1
Fishery Resources, Marine Fisheries. Below is a summary of the latest analysis, building on individual, national and regional reports on the state of resources, accumulated over the period 1974–1998, the last year for which information was available. If all ‘stock’ items (590 in all in the last review) for which FAO had obtained some data are considered together to give a global view of the situation in 1999, 149 appeared to be in an unknown state. Among the 441 for which status information was available, 4% appeared underexploited, 21% moderately exploited, 47% fully exploited, 18% overfished, 9% depleted and 1% recovering (Fig. 1.1). In this chapter, we have used the term ‘overfished’ for stocks simply exploited beyond the level of maximum biological productivity, and ‘depleted’ for those that have been driven to extremely low levels. The former usually still support very active fisheries. The latter are closer to being ‘economically extinct’, hardly supporting a direct fishery. The most pressured resources are redfish, hake, cod, Antarctic cod, lobster, prawns and shrimps. In contrast, the least pressured species are mackerel, bivalves, tuna, cephalopods and horse mackerel (Garcia and Newton, 1997).
Global trends The process of development of the ocean’s resources by fisheries during the last two centuries, and particularly since the Second World War and since the mid-1970s during the period of establishment of the exclusive economic zones, appears as an exponential phenomenon of ‘colonization’ and utilization of potentially available resources that, obviously, cannot continue unabated (Fig. 1.2). The most likely upper limit estimated by the FAO (Gulland, 1971) for world sustainable production is reached if discards and unreported catches are taken into account. The data available in FAO reviews since 1974 (Fig. 1.3) show the impact during the last quarter century. It indicates that, in
This section draws heavily on a preceding paper prepared by the authors (FAO, 2000a).
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Fig. 1.1. State of world stocks in 1999. Stocks tagged as underexploited and moderately exploited are believed to be able to produce more under increased fishing pressure, but this does not imply any recommendation to do so. Stocks tagged as fully exploited are considered as being exploited close to their maximum sustainable yield (MSY) or maximum long-term average yield, and could be slightly under or above this level because of uncertainties in the data and in stock assessments. These stocks are in need of (and in some cases already have) effective control on fishing capacity. Stocks tagged as overexploited or depleted are clearly exploited beyond the MSY level and in need of effective strategies for capacity reduction and stock rebuilding. Stocks tagged as recovering are usually very low compared with historical levels. Directed fishing pressure on these stocks may have been reduced (by management or lack of profitability) but, depending on specific situations, these stocks may nevertheless still be under excessive fishing pressure. In some cases, their indirect exploitation as by-catch in another fishery might be enough to keep them in a depressed state despite reduced direct fishing pressure.
Fig. 1.2. Trends in world production during the last two centuries (modified from Hilborn, 1990). EEZs, exclusive economic zones.
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S.M. Garcia and I. de Leiva Moreno
Fig. 1.3. Global trends in the state of world stocks since 1974. (Key: o = fully exploited; ¡ = underexploited or moderately fished; r = overexploited.)
proportion, stocks at the level of the maximum sustainable yield (MSY) have decreased slightly since 1974 (Fig. 1.3, top line), while stocks offering potential for expansion (Fig. 1.3, middle line) have decreased steadily. As would be expected from these trends, the proportion of overexploited stocks (Fig. 1.3, bottom line) have increased, from about 10% in the mid-1970s to close to 30% in the late 1990s. The number of ‘stocks’ for which information is available has also increased during the same period, from 120 to 454. These trends reflect and confirm the conclusions of an independent global analysis by the FAO of its statistical database using different data, at global or country level (Fig. 1.4). A brief discussion on the validity of the assessment is offered in the last section of this chapter.
Regional perspective As stocks produce less when overexploited systematically, the comparison between present and historical landings in a given region provides an initial crude qualitative assessment of the state of its stocks. The data available for 1999 for the 16 FAO statistical regions
(taken from FAO, 2000b) of the world’s oceans indicate that four of them (25%) are at their maximum historical level of production, eight (50%) are slightly below it and four (25%) are well below it (Fig. 1.5). While this might result partly from natural oscillations in productivity from year to year, comparison with more limited conventional stock assessment data indicate that, in most areas, overfishing is responsible or co-responsible for the decline, together with environmental change. Considering the Pacific and Atlantic Oceans separately, total catches from the Northwest and the Southeast Atlantic are levelling off after reaching their maximum levels a decade or two ago. In the Eastern Central Atlantic and the Northwest Pacific, total catches are increasing again, after a short decline following their maximum production levels of a decade ago. Most of these changes result from increases in landings of small pelagic species. In the Northeast Atlantic, the Western Central Atlantic, the Northeast Pacific, the Mediterranean and Black Seas, the Eastern Central Pacific and the Southwest Pacific, annual catches have stabilized or are declining slightly, having reached their maximum potential a few years ago. In the Southwest Atlantic and the Southeast Pacific
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Global Overview of Marine Fisheries
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Fig. 1.4. Percentage of the minor marine fish resources in various phases of development in the world (Grainger and Garcia, 1996) and in Cuba (Baisre, 2000).
Fig. 1.5.
Ratio between recent (1998) and maximal historical production in FAO statistical areas.
(where the interpretation of historical trends is complicated by the impact of El Niño), total annual catches have declined sharply only a few years after reaching their all-time highs.
These last two areas have been seriously affected by the decline and, in some cases, the serious depletion, of important stocks (e.g. Argentine shortfin squid and Argentine hake
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in the Southwest Atlantic, and anchoveta and horse mackerel in the Southeast Pacific). The areas where total catches are still tending to increase, and where – at least in principle – there is the highest potential for growth, are the Eastern and Western Indian Ocean and the Western Central Pacific. These areas tend to have a lower incidence of fully exploited, overexploited, depleted or recovering fish stocks, and a prevalence of underexploited or moderately exploited stocks, although they also have the highest incidence of stocks whose state of exploitation is unknown or uncertain and for which overall production estimates consequently are less reliable. For a more detailed and analytical diagnosis, the stock assessment data used for Figs 1.1 and 1.3 have been disaggregated by FAO region. The percentage of stocks exploited at or beyond MSY, reflecting the need to control or reduce fishing capacity, ranges from 41% for the Eastern Central Pacific, to 95% in the Western Central Atlantic Ocean. Overall, in most regions, at least 70% of the stocks are already either fully fished or overfished. The percentage of stocks exploited at or below levels of exploitation corresponding to MSY ranges from 43% in the Southeast Pacific to 100% in the Southwest Pacific and Western Indian Oceans. The proportion of stocks that
are exploited beyond the MSY level of exploitation, reflecting straight overfishing, ranges from 0% in the Southwest Pacific and Western Indian Ocean to 57% in the Southeast Pacific Ocean. In the North Atlantic and Pacific, an increasing proportion of stocks were exploited beyond MSY level until the late 1980s or early 1990s (Fig. 1.6). In the North Atlantic, the situation seemed to improve and stabilize in the 1990s, while in the North Pacific the situation seems to remain unstable. The percentage of stocks beyond MSY has been increasing in central and southern parts of the Pacific and Atlantic Oceans since the late 1970s (Fig. 1.7). The increase might be reaching an asymptote in the Atlantic, but this does not seem to be the case yet in the Pacific, and the proportion of stocks affected appears higher in the Atlantic. In the Antarctic, the situation indeed appears more serious, but improving. A further insight in the North Atlantic has been provided by a preliminary analysis of the trends in the state of stocks in the International Council for the Exploration of the Sea (ICES) area since 1970, following the recent introduction of the precautionary approach in the scientific advice framework (Garcia and de Leiva Moreno, 2001). In ICES, the state of stocks is defined in relation to stock-specific
Fig. 1.6. Percentage of stocks exploited beyond maximum sustainable yield in the northern (N) Atlantic and Pacific Oceans.
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Fig. 1.7. Percentage of stocks exploited beyond maximum sustainable yield in the central (C) and southern (S) Atlantic and Pacific Oceans, and in the Antarctic.
precautionary levels of mortality and spawning biomass (Fpa and Bpa) as well as minimum safe levels of these indicators (Flim, Blim). Historical trajectories in relation to F and B are drawn for each stock. The framework has been normalized by Garcia and de Leiva Moreno (2001) to allow joint graphic representation of the position of all stocks in a given year as well as the determination of an annual central position (or ‘centre of gravity’) for all stocks concerned. The updated results of a more complete analysis are given in Figs 1.8 and 1.9 below. Figure 1.8 shows the historical trajectory of the centre of gravity with a worsening of the situation in terms of both F and B from 1970 to 1989 and an apparent improvement in terms of F since then. As a convention, a stock status can be labelled ‘good’ when both indicators of spawning biomass and fishing mortality are better than the precautionary targets, ‘bad’ when both indicators are worse than precautionary limits, and in the ‘buffer’ area when only one of the above indicators is adequate (Fig. 1.9). Figure 1.9, using the same data as Fig. 1.8, indicates a clear worsening of the state of stocks until 1990, and an apparent relative improvement of the situation afterwards, with a significant reduction in ‘bad’ situations, a significant increase in stocks in the
Fig. 1.8. Shift in the ‘centre of gravity’ of the ICES stocks, 1970–1998.
buffer zone, but still extremely few stocks in the ‘good’ area.
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Fig. 1.9.
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Evolution of the proportion of ICES stocks in the various states, 1970–1998.
The Fishing Industry2 The fishing fleet There are no totally reliable or comprehensive data on global fishing power or even of fleet size, while data for small-scale fisheries are scanty. The FAO analyses are usually based on the Lloyds Maritime Information Services and the FAO Bulletin of Fishing Fleet Statistics. Using these sources, and with some caveats, Garcia and Newton (1997) constructed a time series for 1970–1990 of total nominal gross registered tonnage (GRT) of the world fishing fleet, and GRT corrected for technological progress.3 Figure 1.10 presents a qualitative extrapolation of this data to illustrate the fact that the global fishing pressure on the ocean’s ecosystems increased extremely rapidly between the 1950s and
the 1990s through both geographical extension of the fleet operating range (from 1950 to 1970) and adoption of new technologies. Technical improvements are continuing, increasing the fishing capacity of individual vessels, even though the total fleet size shows signs of stabilizing and, perhaps, even decline. During the last few years, the numbers of vessels have tended to decrease in developed countries and to increase in some developing ones. The data for the last few years reflect that, after years of fast growth in the 1960s and 1970s, the total fleet of large fishing vessels has tended to stabilize. The reality of this representation rests heavily on the validity of the correction factors taken from Fitzpatrick (1996) to take account of the effect of technology improvements on fishing capacity. None the less, a characteristic
2
This section draws heavily on FAO (2000b). The correction was based on estimates of improvement in catching efficiency in the main types of vessels and technologies, provided by Fitzpatrick (1996). 3
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Fig. 1.10. Likely trends in world fishing fleet size without (black squares) or with (white squares) correction for technological progress (extrapolated from Garcia and Newton, 1997).
of many fisheries today, and more generally of the fishery sector, is the existence of significant overcapacity, roughly estimated by Garcia and Newton, globally, from 30% (in relation to MSY) to at least 50% (in relation to maximum economic yield, MEY).
The fishers Statistics are scarce, incomplete and not harmonized in concept and coverage, and the following is only intended to give some rough orders of magnitude. According to data provided by FAO and estimates made by this Organization (FAO, 1999b), employment in the primary capture fisheries and aquaculture production sectors, marine and inland, in 1998 was estimated to have been about 36 million people, comprising about 15 million full-time, 13 million part-time and 8 million occasional workers.4 For the first time since the early 1970s, there is indication that growth in employment in the primary sectors of fisheries and aquaculture may be slowing
down significantly (Fig. 1.11). It is estimated that about 60% of the total (22 million people), are employed primarily in marine fisheries, the large majority of them in the small-scale subsector. There can be 4–9 persons per household (Groenevold, 2000) and more than one fisher in each of them.5,6 As a consequence, more than 100 million people would be directly dependent upon fisheries for food, income and livelihood, with many additional people depending on associated sectors, such as marketing, boat building, gear making and bait. Hard data to describe trends of fishers’ living and working conditions are not readily available. These would probably vary greatly between the developed and developing world and between small-scale and large-scale fisheries. Reporting on the perceptions of the communities themselves, in selected countries, Tietze et al. (FAO, 2000c) indicate that facts often contradict widely held preconceptions. In the areas covered by their study, fishing no longer seems to be a ‘last resort’ employment. In many countries, employment has started to stagnate or decline (with the notable
4
Berkes et al. (2001) refer to 51 million fishers in the world, of which 99% are small-scale fishermen and 50 million are in the developing world. Kurien (personal communication) indicates 6.6 persons per household in the Kerala State of India and a dependency rate of 4.4. 6 ICLARM (1999) uses a dependency rate of 5. 5
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Fig. 1.11. 2000a).
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World fishers and fish farmers, including part-time and occasional workers (from FAO,
exceptions of India and Bangladesh) as more rewarding jobs are found outside the fisheries sector. The aquatic resources generally are perceived as seriously degraded by overfishing and pollution. None the less, the annual household income was found to be higher than in the surrounding agricultural households. A general improvement in socioeconomic conditions is perceived in some countries (e.g. the Philippines, Malaysia and India) but not in others (e.g. Senegal, Tanzania and India).
The technology The fishing technologies used to catch fish, as well as handling, preserving and processing fish, have evolved dramatically since the early 1950s as the sector adopted technical innovations coming from other industries, including military technology. In the 1950s, the introduction of synthetic fibres, such as polyamide, polyester and polypropylene, for fishing gear improved fishing capacity
significantly. More recently, the introduction of the Dynema fibre – a polyethylene of ultra-high molecular weight – has marked a new major improvement, reducing gear weight and drag, and thus fuel consumption, allowing an increase in gear size, improving gear life and allowing for more effective exploitation of scattered fish concentrations (e.g. krill or mesopelagic fish), using ‘large mouth’ trawls. The towing of two or more trawls simultaneously (multi-rig trawling), first used in the Gulf of Mexico shrimp fisheries in the early 1970s, was introduced successfully at the end of the 1990s into European fisheries for Norway lobsters (Nephrops sp.), deepwater shrimp and, to some extent, flatfishes. This technology was one of the factors in the rapid extension of tropical shrimp fisheries, increasing efficiency by 50–100%. The introduction of electronic aids for navigation and fish detection, such as Global Positioning Systems (GPS), colour echo-sounders and multi-beam sonars, has greatly increased fishing capacity. In a similar fashion, satellite communications have made a significant impact on fisheries policing (Monitoring,
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Control and Surveillance (MCS)) and to safety on board in the form of Vessel Monitoring Systems (VMS) and the Global Maritime Distress Safety System (GMDSS). It is expected that, in the future, existing technologies, such as voyage data recorders (similar to aircraft flight recorders or ‘black boxes’) and Automatic Identification of Ships (AIS), will further improve MCS. These improvements have increased the capacity to catch fish, farther away from the home ports and to process and preserve it on board for long journeys. Technological improvements have also improved safety on board, reducing the probability of casualties at sea in one of the most dangerous types of employment on earth, with more than 25,000 fatalities each year (FAO, 2000a). Finally, some of the improvements (e.g. selective grids) have also reduced the environmental impacts of fishing. Overall, however, the increase in fishing capacity and the spread of destructive fishing practices probably has significantly increased environmental damage.
Production and trade Reported global production of marine capture fisheries has increased from 19 million t in 1950 to about 80 million t in the mid-1980s,
Fig. 1.12.
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oscillating since then between 78 and 86 million t, excluding discards (Fig. 1.12, higher curve) and represented 67–73% of the overall fisheries production of 112–126 million t, including aquaculture. The relative stabilization of capture fisheries production results in part from the large production reported by China. If this contribution is excluded, the marine capture production of all countries except China has indeed decreased by about 10% during the 1990s (Fig. 1.12, lower curve). The current stagnation in the world marine production is illustrated by the trends in the annual rate of increase of marine catches since 1950 (Fig. 1.13) and shows that it decreased from about 6–9% per year in the late 1950s and early 1960s, to almost zero in the 1990s, crossing the zero relative rate in around 1995. This would indicate that, on average, the world oceans reached their maximal production under the present fishing regime at about that date (for an earlier analysis, see Garcia and Newton, 1997). If the data from China are excluded, the intersection occurs in around 1990 instead of 1995. Regarding the trends in species composition of the landings, Garcia and Newton (1997) underlined the large increase in the proportion of ‘miscellaneous marine fish’ between the 1970s and the 1990s, reflecting ‘the trend in many fisheries towards
Marine capture fisheries production with and without China. Source: FAO-FIDI.
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Fig. 1.13. Relative rate of increase of global marine capture production with China included (thick lines) or excluded (thin lines).
large quantities of unidentified mixtures of fish with low economic value (sometimes called ‘trash fish’) as a result of overfishing and reduction in the size of fish’. They also stressed the significant loss of economic importance of many high value species, such as Atlantic cod, hake and haddock. Total fish trade increased from US$2500–3400 million in 1969–1971, to US$53,000 million in 1999 (an increase from ~ 5 to 9% of total agricultural trade). With time, the growth in trade slowed from 19% annually in 1969–1978 to 9% annually in 1979–1990, and to 4% year−1 in 1991–1999. The contribution of the developing world to such trade has increased regularly since the 1970s. Their share in worldwide exports increased from 32% in 1969–1971 to 44% in 1990 (Garcia and Newton, 1997) and to around 50% (or more in some years) in the 1990s. The lion’s share of this trade is from marine capture fisheries. According to the statistics available in FAO, during the second half of the 1990s, the part of the harvest from capture fisheries internationally traded represented around half of the total capture fisheries production. The value of the exports from marine capture fisheries for the same period was around US$40,000–42,000 million.
Contribution to food security The oceans’ ecosystems contribute substantially to human food security. Coastal ecosystems are the source of more than 90% of the food provided by the marine ecosystem. Coral reefs, for instance, produce 10–12% of the fish caught in tropical countries, and 20–25% of the fish caught by developing nations. As much as 90% of the animal protein consumed in many Pacific Island countries is of marine origin. Part of the production is used directly as human food and part is reduced to fish meal and oil used for raising cattle, poultry and fish. The reported production used for direct human food has increased steadily with time, reaching about 55–57 million t during recent years. When China is excluded, production appears to have stagnated since the mid-1980s (Fig. 1.14). All countries together, the production of food fish per caput appears practically to have doubled since 1950, and has been fairly stable between 9.0 and 10 kg of fish per caput since the early 1970s, despite world population growth. When the data from China are excluded, however, the per caput supply from marine capture fisheries appears to have declined from 11.8 to 9.3 (less 20%) since the mid-1980s (Fig. 1.15).
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Fig. 1.14.
Marine capture fisheries production for human consumption.
Fig. 1.15.
Marine capture fisheries production for human consumption (on a per caput basis).
The proportion of the reported marine capture fisheries production that has been used directly for human food has declined from about 80% in the 1950s and 1960s to about 65% since the early 1970s (Fig. 1.16). In the future, considering that marine capture production cannot increase very much beyond present levels, while the world population will continue to grow (albeit at a slower rate), the per caput supply from marine capture fisheries can only decrease further. Maintaining the fish supply will
require more effective capture fisheries management and substantial development of aquaculture.
Governance Fisheries management is hindered by a number of governance-related factors including, inter alia, unclear objectives, inadequate policies, weak national and regional institutions,
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Fig. 1.16.
S.M. Garcia and I. de Leiva Moreno
Percentage of the marine capture fisheries production used for human food.
poor enforcement, ineffective conventional management measures, lack of participation by stakeholders and, above all, the lack of clear and defendable user rights. The problems, the main issues and the available solutions and pathways to improved sustainability are, however, too complex to be addressed properly in the context of this chapter, which offers below a very broadly drawn description of the situation.
Management approaches Contemporary fisheries governance has developed mainly in the northern hemisphere and spread to the south, with mixed results in all areas. There is no complete global inventory of management systems and approaches having or being implemented, by either countries, stocks or fisheries, and the trends differ between regions. When looking at bio-ecological or socioeconomic performance, it is clear that there is no universally successful management system,7 even though some key principles
and factors of success or failure have emerged and some approaches (i.e. fishing rights) have been more effective than others. In the large majority of FAO members, access to resources has been open (only subject to registration) and free (except for nominal registration fees). With the increased awareness of the impact of overcapacity, many countries have tried to restrict somewhat the flow of operators and capital in the fisheries during the last decade. A number of countries have established limited-entry systems which, in most cases, have failed to impede the build-up of excessive fishing capacity. A growing number of countries, including in the developing world, have experimented with the use of fishing rights, including individual transferable quotas (ITQs) and the approach is gaining attention and support.8 The main difficulty in implementing fishing rights is in the decisions regarding the allocation of such rights, the selection of the right-holders and the modalities of the right (price, duration, transfer, etc.). These necessary decisions, with significant long-term benefits for the State, the right-holders and the consumer,
7 Even the widely acclaimed property rights and ITQs have met with variable success and are not applicable everywhere. 8 According to R. Arnason (personal communication), about 10% of the world catch currently is taken under ITQ systems. In addition, many small-scale fisheries may still be functioning (formally or informally) under traditional systems of territorial use rights.
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have non-negligible short-term economic and socio-political costs, which many politicians still find hard to face. In addition, several technical measures have been tested, including: (i) modified gear design or introduction of selective grids, panels and square meshes to reduce by-catch and discards;9 (ii) bans on discards; (iii) flexible exclusion zones to protect juveniles; (iv) limits on the number of authorizations to fish; (v) reduction or suppression of subsidies; (vi) attempts to deal with fisheries within coastal area management plans; (vii) reserved areas for small-scale fisheries; and (viii) artificial reefs as enhancement and anti-trawl devices. In general, in the absence of improved systems of user rights, these measures have been ineffective in the long run. Zoning, including through marine protected areas (MPAs), has been used conventionally, for example, to protect biodiversity and to keep trawlers away from vulnerable coastal habitats and small-scale fisheries. In general, zoning, alone, will not be effective in a context of overcapacity. MPAs have become fashionable, particularly as a biodiversity conservation device, and if properly enforced may be more effective, particularly if fishing capacity could be controlled or reduced. Control and surveillance has been improved with the use of on-board observers and the development of electronic (satellite-based) VMS.
Management performance Altogether, the paradigm and the tools available have evolved positively during the last decades. Overall, however, and despite some success stories, fisheries governance generally has failed to maintain stocks at their level of maximum biological productivity (i.e. MSY) or above, as shown earlier in this chapter. In addition, large amounts of economic rent have been dissipated (of the order of tens or hundreds of millions of US dollars for most large fisheries). Large economic benefits have been forgone, fisheries collapsed,
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and many coastal communities have been pushed into deeper levels of poverty. The major challenges facing fisheries today include: (i) overfishing, with the related issues of resource collapse and endangered species; (ii) overcapacity, with the related issue of subsidies; (iii) environmental impact of fishing; (iv) illegal, unregulated and unreported fishing (IUU); (v) poor selectivity and discarding; (vi) the environmental state of the coastal zone; (vii) the integration of fisheries management into coastal zone management; (viii) fish trade and eco-labelling; (ix) the interface between fisheries management bodies and CITES; and (x) the collaboration between regional fishery bodies and regional environmental conventions. The requested management paradigm shift to EBFM aims at specifically confronting some of these issues by making them explicitly part of the management problem in addition to the basic and historical problems related to market failures, overcapacity, allocation and rights.
Implementation problems Implementation of the agreed instruments and strategies and fulfilment of the high level commitments already made require guidance, political will and resources. The literature contains enough theoretical principles. The series of technical guidelines produced by FAO in support of the implementation of the Code of Conduct provide ample guidance usable for implementation of EBFM in areas such as: the application of the precautionary approach to capture fisheries (FAO, 1996a), the inclusion of fisheries in integrated coastal zones management (ICZM) (FAO, 1996b) and the use of indicators for the sustainable development of fisheries (FAO, 1999a). Important concerns remain regarding practical implementation of all these instruments and commitments, namely: (i) the lack of institutional capacity in the developing world, as well as in fishing communities
9 An International Plan of Action to reduce by-catch of marine birds in longline fisheries was adopted by FAO in 1999.
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facing decentralization; (ii) the impact of globalization on the fisheries environment and management; (iii) the equity implications (e.g. between poor and rich, or between developed and developing countries) of new developments, such as fishing rights or eco-labelling; (iv) the mismatch of boundaries between the ecosystem and the existing management jurisdictions in exclusive economic zones or regional fishery bodies; and (v) the amount and type of science required as a basis for decision making. Much still has to be done before all the necessary fishery management authorities are in a position to implement EBFM in practice. Changes are required regarding objectives, resource allocation, decision-making processes, enforcement, participation, decentralization, transparency, etc. To improve the situation and allow performance appraisal, the setting of sustainable development reference systems (for systems of indicators of sustainability, see Garcia and Staples, 2000) with appropriate indicators and reference points will be needed. In addition, the limits of the areas of competence of fishery bodies will need to be reconsidered to match better the ecosystems limits (Garcia and Hayashi, 2000), and agreements will need to be elaborated between fishery bodies (e.g. to deal with anadromous or highly migratory species), as well as between coastal countries (e.g. to deal with shared ecosystems).
Regional fishery bodies There are 31 regional fishery bodies operating worldwide, of which nine were established under the FAO Constitution and 24 under international agreements between three or more contracting parties. Their mandates, membership and participation, decision-making procedures, modes of operation and outcomes were the subject of discussion in a recent meeting (11–12 February, 1999) convened by FAO in Rome, in which seven FAO and ten non-FAO organizations participated. Some of the factors hindering progress in the effectiveness of regional fishery bodies are: the failure by some States to
accept and implement relevant international instruments; a lack of willingness by some States to delegate sufficient responsibility to regional bodies; and a lack of enforcement of management measures at both national and regional levels. In the developing world, there is a lack of resources and capacity. Decisions are usually made by consensus, typically engendering ‘too little and too late’ decisions. A number of the regional bodies refer to the precautionary approach, and some of them (e.g. NAFO and ICCAT) have formally started considering the practical means and implications of implementing it. The International Council for the Exploration of the Sea (ICES), responsible for the assessment of the North Atlantic’s resources and management advice, has been implementing it in practice since 1998.
Improved framework In many respects, the context for management has improved a great deal over recent years. Overfishing has been recognized, widely and formally, as a fact and as a problem calling for solutions. New and better policy frameworks have been agreed, and the Law of the Sea, despite its limitations, is an achievement with no equivalent. Since 1990, fisheries frameworks have greatly improved through a range of initiatives at global, regional and national levels: UNCED (Brazil, 1992); the International Conference on Responsible Fishing (Mexico, 1992); the 1993–1995 United Nations Conference on Straddling Fish Stocks and Highly Migratory Fish Stocks (UN Fish Stocks Conference), which led to the opening for signature in December 1995 of the UN Fish Stocks Agreement; the 1992–1993 negotiation of the legally binding Compliance Agreement, which was adopted in November 1993 by the Twenty-seventh Session of the FAO Conference; and the 1993–1995 negotiation of the Code of Conduct for Responsible Fisheries, and its adoption by consensus in October 1995 by the FAO Conference. Since then, the Code has been complemented with a series of implementation technical guidelines,
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including those on integration in coastal areas management, use of sustainability indicators, and the precautionary approach. The Code’s implementation has been strengthened through adoption of four International Plans of Action (IPOAs) to: (i) manage, control and reduce fishing capacity; (ii) manage shark fisheries; (iii) reduce incidental mortality of marine birds in longline fisheries; and (iv) deter, reduce and eliminate illegal, unregulated and unreported (IUU) fishing. A draft fifth IPOA aiming at improving global monitoring of fisheries status and trends will be considered by the FAO Committee on Fisheries (COFI) at its next session. In addition, the broader biodiversity and habitat considerations are being addressed, and the need to protect the ecosystem is broadly accepted as both an ethical principle and a fundamental need. New solutions are being tested, in many cases with success: on the one hand, to control the amount of fishing (through fishing rights and capacity control) and, on the other hand, to preserve critical habitats and biodiversity (through MPAs). Participatory approaches, where fishing communities are involved in the planning, implementation and evaluation of management systems, are receiving increasing support and are being tested in many countries.
Ecosystemic considerations The concept of fisheries operating in an environment or an ecosystem is not new, but pressure is building up to make fisheries and fisheries management more ‘ecosystemconscious’. As this is the main subject of the Conference for which this chapter has been prepared, this section does not pretend to be analytical or prescriptive. As part of the overview, it will only briefly describe how far the sector is prepared to move towards EBFM. Conventional fisheries management, as it developed during the last century, is firmly based on quantitative ecology and ecosystem considerations. At its foundation is the commitment to maintain stocks at their highest level of productivity, with the principle of rebuilding them as a priority when they are
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depleted accidentally – a hard sustainability principle. As economically imperfect as it is, the fishery sector, with public support from financial, scientific and management institutions, discovered the resources, invented or successfully adapted the technology required to catch and utilize them, developed a very dynamic trade, maintained the fisheries’ terms of exchange in the developing world, improved and later maintained food supply per caput despite population growth (with the contribution from aquaculture), and provided livelihoods to more than 100 million people. It collectively failed, however, to maintain the resource base quality, allowing a degradation of the species composition and commercial value of critical habitats, and most probably a modification of the genetic composition. As negative feedback from the ecosystem to the industry and the consumers, resources and vessel performance declined, prices and costs went up, seafood quality and safety decreased, and the death toll among fishermen remained high. Awareness has been growing in various (mainly developed) countries and regions since the Second World War, with a strong acceleration since UNCED in 1992. Most fishery commissions and arrangements largely ignored ecosystem concerns when they were established (with CCAMLR as a notable exception) and remain slow to adjust their agendas, objectives and instruments. However, since the early 1990s, a number of global initiatives of importance for an ecosystembased approach to ocean fisheries management have been undertaken following UNCED, including the Global Plan of Action for the Protection of the Marine Environment (GPA, 1995); the Convention on Biological Diversity (CBD, 1992); the Jakarta Mandate on Marine and Coastal Biodiversity (CBD-JM, 1995); the FAO Commission on Genetic Resources for Food and Agriculture (CGRFA, 1995), which has broadened its mandate to cover aquatic resources; the International Coral Reef Initiative (ICRI), and its three operational units, the Global Coral Reef Monitoring Network, the International Coral Reef Information Network and the International Coral Reef Action Network (ICRAN); the Global Ocean Observing System (GOOS);
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the Marine Protected Areas initiative; and the Large Marine Ecosystems (LME) concept and projects. In addition, the pressure for more EBFM echoes the global consensus developed at and since UNCED, towards the sustainable development of the oceans, as development is considered sustainable (ecologically and socially) only if both human and environmental well-being are ensured, recognizing explicitly the link between the human and environmental elements of the ecosystem and the need for an acceptable balance between them. It is probably fair to assume that the two apparently independent processes (on indicators and ecosystem management) and the process for implementation of the precautionary approach will combine their effects towards the emergence of EBFM. In general, however, the coordination between the environment and fisheries ministries generally is less than optimal, and the implications of these new arrangements and institutions for fisheries are not yet fully understood.
The FAO Code of Conduct While there is not yet any specific global framework for EBFM, the existing fisheries frameworks already contain provisions and guidance related to sustainable development and, more specifically, to ecosystems. The FAO Code of Conduct intends explicitly to conserve aquatic ecosystems (Article 6.1), promote ‘protection of living aquatic resources and their environments and coastal areas’ (Article 2) and respect biological diversity (Code Introduction). The protection and conservation of the ecosystems are objectives of the FAO Code of Conduct (Article 2(g): to ‘. . . promote protection of living aquatic resources and their environments and coastal areas’) and are reflected in its General Principles (Article 6.1: ‘States and users of living aquatic resources should conserve aquatic ecosystems’). Appendix 1.1 lists some of the main ecosystemic considerations in the Code.
Conclusions and Discussion The resources Altogether, the latest information available on the resources and on the fisheries, by fishery, region or globally, tends to confirm the earlier FAO estimates (Gulland, 1971) of a potential for marine fisheries of about 100 million t, of which only 80 million t can probably be harvested for practical reasons, including the difficulty of optimizing the use of every wild stock. It also confirms that a large proportion of the resources are now highly stressed. The statistics on the state of stocks published by FAO often have been interpreted differently by different interest groups following, apparently, the half-full or half-empty bottle parable. The overall perception that one has on the state of world resources depends on whether one views MSY as a target to reach (a conventional view of the fisheries development phase) or a limit to be avoided (a more modern and precautionary view developed during the UN Fish Stock Conference). If ‘fully fished’ stocks, exploited close to MSY, are considered as ‘in trouble’, because, without appropriate management, they are the probable candidates for overfishing in the near future, then a majority (75%) of the stocks appear either fully exploited or overexploited and require either strict capacity and effort control to stabilize levels of exploitation (to MSY) or effort reductions to rebuild stocks (to at or above MSY level) (see Fig. 1.1). If, on the contrary, they are considered as ‘in good shape’ because after all they comply with the Law of the Sea Convention requirement of being at or above the MSY level of abundance, a majority (72%) of the stocks appear in ‘good shape’ and show no sign of overfishing. The same occurs when dis-aggregating the information by FAO region (Fig. 1.17). The relative importance of the dark (i.e. ‘bad’) and the clear (i.e. ‘good’) areas depends on one’s view of fishing around MSY as a desirable goal (Fig. 1.17A) or an undesirable one (Fig. 1.17B). Independently of the point of view, however, 28% of the world stocks appear to
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Fig. 1.17. Proportion of stocks that are in a ‘good’ or ‘bad’ state by FAO region depending on whether fishing around MSY is considered advisable or not.
have been overfished (Fig. 1.3), needing urgent action for rebuilding, and the number of stocks in this poor state have steadily increased between 1974 and 1999, pointing to a failure of management to cope with fishing capacity. The bottle should therefore rather be considered ‘half-empty’ and emptying, and this should be a source of serious concern. When the information is stratified by large oceanic region, the North Atlantic and North Pacific show a continuous aggravation of the situation until the 1980s or early 1990s, with possible stabilization thereafter, particularly in the North Atlantic. In the tropical and southern regions of these oceans, the situation seems to be still exacerbating, except perhaps in the tropical Atlantic, where stabilization and possibly some reversal might have started.
Validity of the assessment Being based on a sample of the world stocks, severely constrained by availability of information to FAO staff, the conclusions have to be considered with caution. A key question
is: to what extent does the information available to FAO reflect reality? There are many more stocks in the world than those to which FAO refers. In addition, some of the elements of the world resources referred to by FAO as ‘stocks’ are indeed conglomerate stocks (and often multi-species). One should therefore ask what validity a statement made for the conglomerate has for individual stocks (sensu stricto). We are generally confident that the global trends we observe in landings reflect trends in the monitored stocks, because the general trends are in agreement with detailed analytical reports and from similar studies conducted at a ‘lower’ level, usually based on more insight and detailed data. As an example, an analysis made by Baisre (2000) on Cuba’s fisheries, using the same approach as Garcia and Grainger (1997) for the whole world, led to surprisingly similar conclusions (Fig. 1.4), using less coarse aggregations, even longer time series, and with more possibility of ‘double-checking’ the conclusions against conventional stock assessment results. There is, of course, the possibility that stocks become ‘noticed’ by scientists, become documented and appear in the FAO information base only when they start getting into
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trouble and scientists, having accumulated enough data, start dealing with them, generating reports of intense fishing or overfishing that FAO can access. This could explain the increase in the proportion of stocks exploited beyond MSY since 1974. This assumption, however, does not hold, for at least two reasons:
•
•
The number of ‘stocks items’ identified by FAO but for which there is not enough information has also increased significantly with time, from seven in 1974 to 149 in 1999, clearly showing that new entries in the system are not limited to ‘sick’ fisheries. From the 1980s, based on the recognition of the uncertainties behind identification of the MSY level, and recognizing also the declines due to decadal natural fluctuations, scientists have become more and more reluctant to classify stocks definitely as ‘overfished’. The apparent ‘plateauing’ of the proportion of stocks with excessive exploitation in the northern regions of the World Ocean may in part be due to this trend.
Natural variability is an important potential source of bias in assessments. During the last two decades, the existence of natural oscillations in marine ecosystems’ composition and productivity, independent of fishing, but probably modified by it, has been definitively recognized. The FAO Expert Consultation to Examine Changes in Abundance and Species Composition of Neritic Fish Resources (San Jose, Costa Rica, 1983) (Csirke and Sharp, 1984) was an important step in that direction following the recognition of synchronous changes of abundance in a number of important sardine stocks (Kawasaki, 1983). The report produced more recently for FAO by Klyashtoryn (2001) illustrates very clearly the fact that the historical trends in world catches of important variable marine resources of the northern hemisphere (mainly small pelagic and semi-pelagic species, but including cod) are closely related to natural long-term climate oscillations. The extent to which these oscillations, now recognized as more general than previously thought, may have affected
some of the scientific assessments on which this review relies, is not known. One criticism of modern assessments is that the data available for most stocks underestimate grossly and greatly the ‘pristine’ levels of abundance and hence the present degree of degradation of fishery resources (Pauly, 1995). More recent papers proposing the same view also argue that overfishing is the main cause of degradation of the state of world stocks, above pollution and all other forms of human intervention. There certainly will be a heated debate around these assertions, and efforts are being made to collect as much historical data as possible to confirm or confute them. If they are correct, however, they would imply that the present assessment might still be too optimistic. The high pressure exerted on stocks has ecological effects, which have been stressed repeatedly and are demonstrated by the change in quality (species composition, size, commercial value, trophic level) observed in the landings, largely reflecting changes in the resource base (Lock, 1986; Caddy, 1993; Garcia and Newton, 1997; Pauly et al., 1998; Caddy and Garibaldi, 2000; Pauly and Palomares, 2000). Few resources could support higher fishing pressure, and these tend to be prey species for which an increase in exploitation may lead to questionable ecological consequences, including for predator stocks that society would like to see rehabilitated. There is practically no other fishing area or resource of significance to be discovered. The world’s oceans are exploited from the poles to the tropics, the littoral to the open sea, and the surface to the deep bottom. Deep-sea resources on slopes and sea mounts are already under heavy pressure (and possibly overfished in many areas), and their low natural productivity and resilience put them in serious danger. There is hardly anywhere else to go to employ existing excess fishing capacity.
The fishing industry Despite the obvious problems in the resource base, marine fisheries have become an
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important source of economic and social development in the coastal areas, where they are, however, competing with other activities for resources and for space. They provide:
•
•
• • •
Food: total production and production per caput have been maintained at the cost, however, of a decrease in the quality of the harvest and with growing support from mariculture and coastal aquaculture. Economic benefits: fisheries are a source of benefits, profits and revenue for hundreds of millions of people, a large proportion of them poor or very poor, and are an important source of foreign exchange for many developing countries. Employment and livelihood: particularly in poor coastal areas and for the poorest strata of the population. Recreation: sport fisheries provide a significant contribution to recreation and tourism. Data: despite the ongoing debate on the quality of the fishery data, fisheries have contributed a quantity of information that has hardly any equivalent in any other sector and is extremely valuable for the monitoring of the sector and the resources upon which it depends.
Technical progress continues to improve safety on board, as well as capacity to fish and, in the absence of effective management mechanisms, continues to fuel overcapacity and environmental damage. Similarly, in the absence of effective management, the potential benefits of globalization and free trade may be missed, leading to further stress. Governments have started grappling with the issue, and an International Plan of Action for the Management of Fishing Capacity was adopted at FAO in 1999. The issue of subsidies to the sector and their impact on capacity and sustainability have become an important and sensitive issue. Illegal fishing is a significant component of the overfishing and overcapacity problem, and an International Plan of Action to Prevent, Deter and Eliminate Illegal, Unreported and Unregulated Fishing was also adopted in FAO, in 2001.
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It is not possible in such a short chapter to do justice to the complex changes occurring in the sector. It is also not easy to summarize the sector’s evolution as it is rather heterogenous and its trends depend on the regions, and sometimes on the resource types concerned. The relative political support that the smallscale and industrial sectors enjoy depends on countries. In general, however, the sector (particularly the small-scale subsector) is dispersed and weakly organized. The capacity of industry to influence government decisions is highly variable and, in the absence of formal participatory decision making, is essentially non-transparent. This capacity could become critical at a time when governments will have to decide how to ‘allocate’ to the various sectors the coming ‘environmental bill’. Large and vertically integrated food companies and major supermarkets are playing a growing regulatory role in supplies and prices. The sector has developed an awareness of the environmental issues and is getting more involved in the international debate, e.g. attending more regularly the FAO COFI meetings or being deeply involved in the debate about the pros and cons of eco-labelling. In the coming evolutionary process of fisheries governance towards EBFM, a much greater involvement of fisheries will be essential.
Governance While generalization is always dangerous, the management approaches currently in use reflect largely the body of knowledge and theories developed between the 1940s and 1960s, with improvements due to scientific and technical progress. It is generally recognized that the main problem for fisheries management is the inter-related inadequacies of policies, regulations, enforcement and systems of rights, agggravated by insufficient knowledge and inadequate science. A number of measures have been tested, but there is, as yet, no general consensus even though the need for fishing rights emerges among the scholars and some industry leaders as a winning solution.
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Generally viable solutions have still to be found for small-scale fisheries, and developing countries need collaboration aiming at faster capacity building. The clearly demanded shift to EBFM and wider application of the precautionary approach, supported by generalized use of sustainability indicators systems, requires more investments in governance, better science, more efficient decision making, more deterrent enforcement, higher levels of participation, decentralization, transparency, as well as a better matching between jurisdictions and ecosystem boundaries. There is no doubt that management must change, and there is agreement on the general long-term goal as well as on potentially winning solutions. However, solutions must be tailored to particular socio-economic situations, identifying effective and affordable pathways. The Law of the Sea is the foundation on which to build the new system of governance, and the FAO Code of Conduct is recognized, generally, as the operational instrument for its practical application within the UNCED principles. The regional fishery management organizations and arrangements have been recognized as the central institutions for fisheries governance, but they will have to improve their performance to reduce potential duplication and conflicts with environmental institutions and organizations.
Need for improved information It is clear that the monitoring of the state of fisheries and their resources (and environment) needs to be strengthened substantially in the interest of the sector itself, for better informed and improved governance as well as for more transparency and better public information. No matter how and how much the baselines might need to be ‘corrected’ to gauge the system better, the fisheries management dashboard should be better able to reflect the state of its main components. The FAO data are usually taken as the reference source for global information, but they have their shortcomings. Efforts are being made to
improve the data, assisting individual countries in revising their data collection systems, elaborating manuals and providing training courses. In addition, and in order to improve quality, timeliness and transparency, FAO has started the development of a global, cooperative, Fisheries Resources Monitoring System (FIRMS), connected to the FAO Fisheries Global Information System (FIGIS), with a view to mobilize better the competences and information available in the regional fishery commissions and national centres of excellence.
Aknowledgements We wish to thank all of those colleagues who contributed knowledge and time for this chapter, providing ideas, data or constructive comments, and in particular A. Crispoldi, L. Garibaldi, R. Grainger, A. Smith, S. Vannuccini and R. Willmann. We also gratefully acknowledge the critical and constructive contribution of the reviewers, M. Sinclair and R. Arnason, to the clarity of the chapter. As usual, we remain solely responsible for any remaining inaccuracy.
References Baisre, J.A. (2000) Chronicles of Cuban marine fisheries (1935–1995): trend analysis and fisheries potential. FAO Fisheries Technical Paper No. 394. Berkes, F., Mahon, R., McConney, P., Pollnac, R. and Pomeroy, R. (2001) Managing Small-scale Fisheries. Alternative Directions and Methods. International Development Research Centre, Ottawa, Canada. Caddy, J.F. (1993) Towards a comparative evaluation of human impacts on fisheries ecosystems of enclosed and semi-enclosed seas. Reviews in Fisheries Science 1, 57–95. Caddy, J.F. and Garibaldi, L. (2000) Apparent changes in the trophic composition of world marine harvests: the perspective from the FAO capture database. Ocean and Coastal Management 43 (8–9), 615–655. Csirke, J. and Sharp, G.D. (1984) Proceedings of the Expert Consultation to examine changes in abundance and species of neritic fish
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resources. San José, Costa Rica. 18–29 April 1983. FAO Fisheries Report 291(2). FAO (1996a) Precautionary approach to capture fisheries and species introductions. FAO Technical Guidelines for Responsible Fisheries No. 2. FAO (1996b) Integration of fisheries into coastal areas management. FAO Technical Guidelines for Responsible Fisheries No. 3. FAO (1999a) Indicators for sustainable development of marine capture fisheries. FAO Technical Guidelines for Responsible Fisheries No. 8. FAO (1999b) Numbers of fishers 1970–1997. Nombre de pêcheurs 1970–1997. Nùmero de pescadores 1970–1997. FAO Fisheries Circular No. 929, Rev. 2. FAO (2000a) The State of World Fisheries and Aquaculture 2000. FAO (2000b) Fishstat Plus: universal software for fishery statistical time series, Version 2.3. FAO, Rome. FAO (2000c) Demographic change in coastal fishing communities and its implications for the coastal environment. In: Tjetze, U., Groenevold, G. and Marcoux, A. (eds) FAO Fisheries Technical Paper No. 403. Fitzpatrick, J. (1996) Technology and fisheries legislation. In: Precautionary approach to fisheries. Part 2, Scientific papers. FAO Fisheries Technical Paper No. 291(1). Garcia, S.M. (1997) Indicators of sustainable development of fisheries. Land quality indicators and their use in sustainable agriculture and rural development. FAO Land and Water Bulletin No. 5, pp. 131–162. Garcia, S.M. and de Leiva Moreno, I. (2001) Marine fisheries resources: global state. In: Steele, J., Thorpe, S. and Turekian K. (eds) Encyclopaedia of Ocean Sciences. Academic Press, pp. 1584–1589. Garcia, S.M. and Grainger, R.J.R. (1997) Fisheries management and sustainability: a new perspective or an old problem? In: Hancock, D.A., Smith, D.C., Grant, A. and Beumer, J.P. (eds) Developing and Sustaining World Fisheries Resources: the State of Science and Management. Proceedings of the 2nd World Fisheries Congress, CSIRO, Australia, pp. 631–654. Garcia, S.M. and Hayashi, M. (2000) Division of the oceans and ecosystem management: a contrastive spatial evolution of marine fisheries governance. Ocean and Coastal Management 43, 445–474. Garcia, S.M. and Newton, C. (1997) Current situation, trends and prospects in world capture fisheries. In: Pikitch, E.L., Huppert, D.D. and Sissenwine, M.P. (eds) Global Trends: Fisheries Management. America
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Fisheries Society Symposium, 20, Bethesda, USA, pp. 3–27. Garcia, S.M. and Staples, D. (2000) Sustainability reference systems and indicators for responsible marine capture fisheries: a review of concepts and elements for a set of guidelines. Sustainability indicators in marine capture fisheries. Special Issue. Marine Fisheries Research 51, 385–426. Grainger, R.J.R. and Garcia, S.M. (1996) Chronicles of marine fishery landings (1950–1994). Trend analysis and fisheries potential. FAO Fisheries Technical Paper No. 359. Groenevold, G. (2000) Sociodemographic characteristics and change in coastal fishing communities. In: Tietze, U., Groenevold, G. and Marcoux, A. (eds) FAO Fisheries Technical Paper No. 403, pp. 60–77. Gulland, J.A. (1970) The state of world resources. FAO Fisheries Technical Paper No. 97. Gulland, J.A. (1971) The Fish Resources of the Ocean. Fishing News Books (Intl.), Oxford. Hilborn, R. (1990) Marine Biota. In: Turner, B.L. III (ed.) The Earth as Transformed by Human Action. Cambridge University Press, Cambridge, pp. 371–386. ICLARM (International Centre for Living Aquatic Resources Management) (1999) ICLARM’s Strategic Plan 2000–2020. ICLARM, Manila, Philippines. Kawasaki, T. (1983) Why do some pelagic fishes have wide fluctuations in their numbers? – biological basis of fluctuation from the viewpoint of evolutionary ecology. In: Sharp, G.D. and Csirke, J. (eds) Reports of the Expert Consultation to Examine Changes in Abundance and Species Composition of Neritic Fish Resources. FAO Fisheries Report 291(3), pp. 1065–1080. Klyashtorin, L. (2001) Climate change and longterm fluctuations of catches: the possibility of forecasting. FAO Fisheries Technical Paper No. 410. Lock, J.M. (1986) Effects of fishing pressure on the fish resources of the Port Moresby barrier and fringing reefs. Department of Primary Industries. Fisheries Division. Port Moresby (Papua New Guinea) Technical Report 86(3), 1–31. Pauly, D. (1995) Anecdotes and the shifting baseline syndrome of fisheries. Tree 10, 430. Pauly, D. and Palomares, M.L. (2000) Approaches for dealing with three sources of bias when studying the fishing down the food web phenomenon. Fishing down the Mediterranean Food Webs? CIESM Workshop Series No. 12, pp. 61–66. Pauly, D., Christensen, V., Dalsgaard, J., Froese, R. and Torres, F., Jr (1998) Fishing down marine food webs. Science 279, 860–863.
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Appendix 1.1: Ecosystem-related provisions of the FAO Code of Conduct for Responsible Fisheries The Code provides for:
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• •
•
• • •
assessment of impacts on target stocks, associated or dependent species (Articles 7.2.3; 12), including before introducing any new fishing method or operation in an area (Articles 8.4.7; 12.11); monitoring of the environment and of the impacts on it (Articles 8.4.7; 10.2.4; 12.11); reduction and minimization of environmental impact (pollution, discards, ghost fishing) on target and associated, dependent or endangered species (Articles 7.2.2; 7.6.9); protection and restoration of critical habitats such as wetlands, mangroves, reefs, lagoons, nursery and spawning areas from degradation, destruction, pollution, etc., from human activities (Articles 6.8; 7.6.10); prohibition of destructive fishing (Article 8.4.2); allocation of rights subject to ecosystem conservation (Article 6.1); maintenance of the quality, diversity and availability of resources (Article 6.2);
• • • • • • • • • • • • •
restoration/rehabilitation of populations and stocks (Articles 6.3; 7.2.1); assessment of relationships among the populations in the ecosystem (Articles 7.2.3; 12); improvement of selectivity (Articles 8.5.3; 12.10); reduction of impacts on target and non-target stocks (Articles 6.2; 12.10); conservation of biodiversity and population structure (Articles 6.6; 7.2.2); prevention of overfishing and overcapacity (Article 6.3); protection of endangered species (Article 7.2.2); assessing of gear impact on biodiversity and coastal communities (Articles 8.4.8; 10.2.4; 12.5); assessing impact on non-fishing activities (Article 12.5); assessing impact of climate change (Article 12.5); adopting measures to maintain or restore stocks at levels capable of producing maximum sustainable yield (Article 7.2.1); apply widely the precautionary approach (Article 7.5.1); and ensuring a level of fishing commensurate with the state of fisheries resources (Article 7.6.1).
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Obligations to Protect Marine Ecosystems under International Conventions and Other Legal Instruments Transform Aqorau South Pacific Forum Fisheries Agency, Solomon Islands
Abstract Traditional approaches to fisheries management, which have been singular, species-based and nonsectoral, have failed to protect the world’s fisheries resources. This has resulted in the overexploitation of fish stocks, displacement of fishing fleets and dislocation of fishing communities. The first attempts at international regulation of fisheries were simple, but premised on the notion that the ocean’s resources were inexhaustible. This belief influenced attitudes towards exploitation of the fisheries resources in particular, and conservation and management of those resources in general. Improved understanding of the oceans and the exhaustibility of fisheries resources has resulted in a change both in the approach towards fisheries management and in the kind of responses developed by the international community. It is now accepted that it is not possible simply to manage a fishery by merely controlling the quantity of fish taken out of the water. Account must be taken of all factors affecting the resource, including the impact of human activities from land-based sources. New, more modern, comprehensive and holistic approaches have been developed in recent years to address fishery problems. One of these is the ecosystem management approach. This chapter examines a number of key international instruments that demonstrate the extent to which ecosystem management has been applied to conservation of fishery resources. It begins by outlining the major obligations of instruments of global applications, which include, inter alia, the 1982 United Nations Convention on the Law of the Sea (LOSC), and the 1992 Convention on Biological Diversity (CBD). The chapter also discusses a number of regional and national initiatives towards ecosystem approaches to fishery conservation and management, and highlights the problems impinging on the effective implementation of these initiatives. It concludes by pointing out that the most notable strengths of the international instruments studied for this chapter are the instruments themselves, as they attempt to establish a global framework for conservation and management of marine environments and resources. The inclusion of the ecosystem approach is a positive element, as it moves away from the traditional species and stock focuses.
Introduction The crisis in international fisheries has been caused largely by overexploitation of fisheries resources fuelled by the perception that fishery resources are inexhaustible and are
common property. This has led to overexploitation. Improvements in fishing technology increased fishing effort, resulting in declines in stocks in many parts of the world (UNGA, 2001). FAO has warned that most of the world’s fish stocks are fully
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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exploited or are nearing levels of overexploitation. Because of their importance to world food supply, it is in the interest of the world community to ensure that the world’s fish stocks are managed sustainably. Government attitudes towards proper management of fish stocks have not helped. A report highlighting the failure of governments to embrace holistic approaches noted that: Governments have traditionally addressed human activities on a piecemeal basis, separating decision-making on environmental quality from decision-making on natural resource management or on social or economic issues. Even within the environmental field, agencies have traditionally managed air issues separately from those dealing with water, land or wildlife. An ecosystem approach to management is a holistic approach that recognises the interconnectedness of and addresses the linkages occurring among, air, water, land and living things. (Mehan, no date)
By the late 1980s, the international community began to take serious action to address the crisis in international fisheries. These actions, however, were a little too late to save communities dependent on fishing. For instance, by 1992, the fishing community along the east coast of Canada, whose livelihood depended on the cod fishery, was disrupted by the collapse of the cod fishery, displacing thousands of fishers. This was a wake-up call, albeit too late, for the international community. The piecemeal approach to fishery management had failed. In an attempt to avoid such problems, international instruments developed over the past two decades have included new obligations for management activities regulating uses of the oceans. The conventions (and codes) make explicit reference to protection of ecosystem features. The overarching convention in this respect is the 1992 Convention on Biological Diversity (CBD). Other international legal instruments include the UN Agreement for the Implementation of the Provisions of the United Nations Convention on the Law of the Sea of 10 December 1982 concerning the Conservation and Management of Straddling and Highly Migratory Fish
Stocks (‘UN Fish Stocks Agreement’) and the FAO Code of Conduct for Responsible Fisheries (FAO, 1995) (‘Code of Conduct’). At the regional and national level, legislation and policies have been put in place to incorporate ecosystem considerations more explicitly within national ocean management regimes. It has been argued that a ‘healthy ecosystem is good for fisheries and good for the environment and contributes to quality of life’ (Fluharty, 2001). This chapter examines the provisions of selected international instruments to demonstrate the extent to which ecosystem management has been incorporated in the instruments, using the Australian Ocean Policy and the Canadian Oceans Act, amongst others, as examples. It also examines the implementation of ecosystem management principles at the domestic level.
International Instruments That Apply the Ecosystems Approach to Fisheries Management The international instruments discussed are:
• • • • • •
• •
1982 United Nations Convention on the Law of the Sea (LOSC) 1992 Convention on Biological Diversity (CBD) Jakarta Ministerial Statement on the Implementation of the Convention on Biological Diversity 1995 UN Fish Stocks Agreement (Fish Stocks Agreement) 1995 FAO Code of Conduct for Responsible Fisheries (Code of Conduct) 2000 Convention for the Conservation and Management of Highly Migratory Fish Stocks in the Western and Central Pacific Ocean Forum Fisheries Agency (WCPT Convention) (FFA, 2000) 2001 Convention on the Conservation and Management of Fishery Resources in the South East Atlantic Ocean (SEAFO Convention) 1976 Convention on Conservation of Nature in the South Pacific (SPREP, 2000)
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Agreement for the Establishment of the Regional Commission for Fisheries (RECOFI) (FAO, 2001b) Convention on the Conservation of Antarctic Marine Living Resources (CCAMLR) (1982) 1994 Convention for the Establishment of the Lake Victoria Fisheries Organization (LVFO) 1949 (with subsequent amendments, last 1997) Agreement for the Establishment of the General Fisheries Commission for the Mediterranean (GFCM) 2000 Framework Agreement for the Conservation of Living Marine Resources on the High Seas of the South Pacific (Galapagos Agreement) Washington Declaration on Protection of the Marine Environment from Landbased Activities (1995).
The agreements and instruments discussed in this chapter do not comprise an exhaustive list of applicable conventions. Their inclusion does not imply that they are the only instruments that give rise to obligations to apply the ecosystems approach to fisheries management. They are merely indicative of trends emerging in various parts of the world, and exemplify the shifts in approaches towards fisheries management evident in different parts of the world. Table 2.1 describes the status of the various agreements discussed in this chapter. The national policies examined include Australia’s Oceans Policy, the EU Green Paper and the Canadian Oceans Act 1996.
The United Nations Convention on the Law of the Sea LOSC was opened for signature on 10 December, 1982, and entered into force on 16 November, 1994. LOSC is the basic legal framework that governs the uses of the oceans and seas. LOSC also establishes a framework for the development of conservation and management measures concerning marine resources and scientific research within the exclusive economic zone (EEZ) of a State, as well as on the high seas. It is
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accepted that most provisions of LOSC reflect customary international law.
Protection and preservation of the marine environment Fundamental to the management and conservation of fisheries resources is the protection and preservation of the marine environment. Part 12 of LOSC outlines provisions for the protection and preservation of marine ecosystems. These provisions are very broad and so are applicable to fisheries on a global scale. All States have a duty to undertake measures to protect the marine environment and to control, reduce and manage pollution of the sea (§192 and 194). Although the provisions in this part of the Convention do not refer specifically to fisheries, they are relevant in the sense that they urge States to prevent, reduce and control pollution of marine ecosystems through any source (§194(1)), and this could include debris and waste from fisheries operations. The provisions relating to the protection and preservation of the marine environment emphasize the importance of cooperation between States and the need for States to undertake surveillance of activities that they permit or engage in, in order to determine whether these activities are likely to have significant adverse impacts on the marine ecosystem and its various components (§204(2)).
Conservation of the living resources within the EEZ With respect to the living resources, Parties are required to establish measures for the conservation and management of marine living resources in their EEZs. These measures must take into account, inter alia, the effects of harvesting target species on species that are associated with or dependent upon the harvested species, whilst ensuring that living resources are not endangered by overexploitation (§61(2) and (4)). In addition, LOSC addresses highly migratory species, marine mammals, and anadromous and catadromous stocks to ensure that these species are conserved and managed in their State of origin and external areas (§64–67).
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Table 2.1.
Status of agreements.
Treaty
Area of application
United Nations Convention on the Global Law of the Sea Global Convention on Biological Diversity Global UN Fish Stocks Agreement Convention for the Conservation and Management of Highly Migratory Fish Stocks in the Western and Central Pacific Ocean Convention on the Conservation and Management of Fishery Resources in the South East Atlantic Ocean Convention on Conservation of Nature in the South Pacific Agreement for the Establishment of the Regional Commission for Fisheries (RECOFI) Convention on the Conservation of Antarctic Marine Living Resources Convention for the Establishment of the Lake Victoria Fisheries Organization Agreement for the Establishment of the General Fisheries Commission for the Mediterranean Framework Agreement for the Conservation of Living Marine Resources on the High Seas of the South Pacific, ‘Galapagos Agreement’
Regional (Western and Central Pacific)
Date of adoption 10 December, 1982 22 May, 1992 5 September, 1995 4 September, 2000
Entry into force
Contracting States
16 November, 1994 23 December, 1993 11 December, 2001 Not in force
138
Nil
182 31 3
Regional (South East Atlantic
20 April, 2001
Not in force
Regional (South Pacific) Regional (Gulf States)
12 June, 1976
28 June, 1990
6
11 November, 1999
26 February, 2001
5
Regional (Antarctic and Southern Oceans) Regional (Southern Africa)
20 May, 1980
7 April, 1982
27
30 June, 1994
24 May, 1996
3
Regional (Mediterranean)
24 September, 1949
3 December, 1963
21
Regional (South Pacific)
14 August, 2000
Not in force
Source: http://fao.org/legal/TREATIES
Conservation of the living resources of the high seas LOSC provides that all States have the right for their nationals to engage in fishing on the high seas provided that they do not contravene LOSC objectives and are consistent with Articles 63(2) and 64–67, as well as with provisions dealing with the high seas (§116). States are obliged to undertake measures to conserve the living resources of the high seas and, in doing so, must cooperate with each other and establish regional or subregional
fisheries organizations as appropriate, to promote this objective (§118).
The Convention on Biological Diversity CBD was signed on 5 June, 1992, and entered into force on 23 December, 1993. CBD provides an international framework for the conservation and ecologically sustainable development and use of biodiversity. The Convention does not address fisheries
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specifically. However, it applies to all terrestrial and marine biodiversity, and thus affects fisheries. The Convention outlines, inter alia, measures for conserving biodiversity which include general, in situ1 and ex situ2 conservation measures. General measures for conserving biodiversity and ensuring ecologically sustainable development include developing national policies, strategies and programmes that should, inter alia, reflect the principles espoused in the CBD (§6(a)). The CBD also urges Parties to integrate biodiversity conservation policies and strategies with crosssectoral plans (§6(b)). In situ conservation Measures outlined for in situ conservation of biodiversity encompass certain key issues. These include, inter alia protected areas, ecosystems and habitats. With respect to protected areas and ecosystems, the CBD imposes the following obligations on all Contracting Parties:
•
•
•
Protected areas – establish a system of protected areas for conserving biodiversity; and – develop guidelines for the selection, establishment and maintenance of protected areas. Biological resources – regulate and manage biological resources that are important for conserving biodiversity within protected areas and in ex situ circumstances; and – promote ecologically sustainable development in areas adjacent to protected areas with a view to protecting these areas to complement protected areas. Ecosystems and habitats – rehabilitate and restore degraded ecosystems, inter alia, through the
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development and implementation of management plans and strategies; – promote the (in situ) protection of ecosystems, natural habitats and the maintenance of viable populations of species; and – eradicate, prevent or control and manage alien species that threaten native habitats and species (§8 (a)–(f), (h)). The Parties are obliged to regulate and manage threatening processes affecting or likely to affect biodiversity in an adverse manner (§8(l)). Additionally, Parties must develop and implement measures to control and manage the risks associated with potentially threatening activities, such as the use and release into the environment of organisms that have been modified through biotechnology (§8(g)). Ex situ conservation In addition to outlining measures for in situ conservation of biodiversity, the CBD recommends that all Parties undertake activities to ensure the protection of biodiversity in ex situ circumstances, although such activities should complement the in situ conservation measures articulated in the Convention (§9). The CBD requires Parties to undertake the following ex situ biodiversity conservation measures:
• •
•
1
Establish and maintain facilities for ex situ conservation of, and research into, biodiversity in the country of origin of the biodiversity in question. Adopt measures to ensure the recovery and rehabilitation of threatened species, and the re-introduction of such species into their natural habitats under appropriate conditions. Regulate and manage the collection of biological resources from habitats to
CBD, §2: ‘in situ conservation’ means the conservation of ecosystems and natural habitats and the maintenance and recovery of viable populations of species in their natural surroundings and, in the case of domesticated or cultivated species, in the surroundings where they have developed their distinctive properties. 2 CBD, §2: ‘ex situ conservation’ means the conservation of components of biological diversity outside their natural habitat.
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ensure that the survival of in situ species, populations and ecosystems is not threatened. Cooperate in providing financial and other support for ex situ conservation measures, particularly to developing nations (§9(a)–(e)).
reduction in overcapacity and subsidies to fisheries (Tsamenyi and McIlgorm, 1999). This resolution was endorsed at COP 2 (Tsamenyi and McIlgorm, 1999). Furthermore, the SBSTTA was also responsible for the following:
• •
The Jakarta Ministerial Statement on the Implementation of the Convention on Biological Diversity The Jakarta Ministerial Statement on the Implementation of the Convention on Biological Diversity (Jakarta Mandate on Coastal and Marine Biodiversity) was issued during the second meeting of the Conference of Parties to the CBD (COP 2), held in Jakarta in November 1995, as a result of the Conference of Parties (COP) identifying marine and coastal biodiversity as a high priority issue. The Mandate essentially re-affirms the importance of the conservation and ecologically sustainable use of coastal and marine biodiversity and urges the COP to initiate the immediate development and implementation of actions concerning this issue. The Mandate specifically links conservation, the use of biodiversity and fishing activities, and establishes a new global consensus on the importance of marine and coastal biodiversity. The Mandate identifies the following areas as being of critical importance (Tsamenyi and McIlgorm, 1999):
• • • • •
Integrated management of marine and coastal areas. Marine and coastal protected areas. Ecologically sustainable use of marine and coastal living resources. Mariculture. Alien species.
A Subsidiary Body on Scientific, Technical and Technological Advice (SBSTTA) was established following the Ministerial Declaration, with marine and coastal biodiversity appointed as the first key sector to be investigated by the SBSTTA (Tsamenyi and McIlgorm, 1999). The SBSTTA was responsible for, inter alia, a resolution calling for a
•
Establishing a roster of Experts on Marine and Coastal Biological Diversity. Applying the precautionary approach to biodiversity. Implementing integrated marine and coastal area management (Tsamenyi and McIlgorm, 1999).
The UN Fish Stocks Agreement The failure of LOSC to prevent overexploitation of fish stocks, especially highly migratory and straddling fish stocks on the high seas, led to negotiations that resulted in the conclusion of the UN Fish Stocks Agreement. The UN Fish Stocks Agreement provides a level of detail not found in LOSC, for the management and conservation of highly migratory and straddling fish stocks. The fundamental objective of the UN Fish Stocks Agreement is to ensure the long-term conservation and sustainable use of straddling fish stocks and highly migratory fish stocks through effective implementation of relevant LOSC provisions. The UN Fish Stocks Agreement imposes certain obligations on Parties with regard to the protection of the marine environment. In general, the Agreement requires that States ensure the sustainable utilization of fish stocks, and that they assess the impacts of fishing on the marine environment. For instance, Parties must assess the impacts of fishing, other human activities and environmental factors on target species, species that are part of the same ecosystem, and species that are associated with or dependent upon target species (§5(d)). In doing so, Parties must take into account the precautionary principle and uncertainties relating to data used in the development of conservation and management measures (§6(c)). Furthermore, data collection and research programmes must be established for assessing the impacts of
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fishing on non-target (fish and non-fish) species (§6(e)). Parties must adopt appropriate conservation and management measures to maintain or restore populations of species that are a part of the same ecosystem as target species, or are associated with or dependent upon target species (§5(e)). Parties must also establish conservation and management measures for habitats of special concern (§6(d)). Parties must minimize discards, waste and by-catch of target and non-target species through various measures, including the development and use of selective fishing gear and techniques (§6(f)). Where stock populations of target species and populations of non-target species are of concern, Parties must enhance monitoring of those species and review their management and conservation status. Parties are also obliged to collect and share all relevant and up to date fisheries data (§5(j)). Annex I of the UN Fish Stocks Agreement provides standard requirements for the collection and sharing of data. Data that can be collected include information on vessel position, catch and yield statistics, composition of catch, including target and non-target species (Annex I, §3), fishing gear description, etc. States are also required to establish mechanisms for verifying fisheries data; mechanisms include scientific observer programmes for monitoring details of fishing operations such as catch composition (target and nontarget species) (Annex I, §6). Management strategies aimed at restoring or maintaining populations of species associated with or dependent upon target species must do so at levels consistent with precautionary reference points (Annex II, §4). Flag States are obliged to record and report catch and vessel information (Annex II, §18(e)). Flagged vessels must also have their catch of target and non-target species verified through measures such as observer programmes and inspection schemes (Annex II, §18(f)). The fishing activities of flagged vessels must be regulated to ensure compliance with subregional, regional or global by-catch reduction measures (Annex II, §18(l)). The UN Fish Stocks Agreement is a departure from the traditional, species-based approach
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to fisheries management. Its implementation will strengthen global application of ecosystem-based fisheries management.
FAO Code of Conduct for Responsible Fisheries The Code of Conduct for Responsible Fisheries addresses specific impacts of fisheries on the marine and aquatic environment, including by-catch, and marine resource protection. The Code is not legally binding, but links other international fisheries obligations, including those established under LOSC. The general principles of the Code suggest that fisheries management measures should ensure the protection not only of target species but also of non-target, associated or dependent species (§6.2). Under the Code, States are urged to apply the precautionary principle in conserving, managing and exploiting fisheries resources (§6.5). States are to ensure, inter alia, the use of selective fishing gear, and reduce waste, discards and catch of non-target species (fish and non-fish) (§6.6). Furthermore, States are encouraged to reduce the impacts of fisheries on species associated with or dependent upon target species (§6.6). The provisions have the scope to provide effective protection of marine ecosystems by protecting target and non-target species and the ecosystems associated with those species. In addition, the Code requires States to implement appropriate measures (within the precautionary principle framework) so as to minimize waste, discards, ghost-fishing, by-catch and negative impacts of fishing on associated or dependent species (§7.22, 7.52 and 7.69). The principles also require fisheries management authorities to promote the development and use of selective gear and efficient operational methods as part of their overall effort to conserve the marine environment (§7.69). States must ensure that regulations related to measures for the reduction of waste, discards and by-catch are not circumvented by technical devices (§8.51). The Code suggests that reduction of waste and by-catch may be achieved by technical
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measures, such as modifying gear to prevent smaller, unwanted species or individuals being trapped in the net (§8.51). States are also required to improve their understanding of the status of fisheries by collecting appropriate data and exchanging information with all relevant groups (§12.4). The Code is a part of a new generation of treaties that provide a higher benchmark for fisheries management.
Convention for the Conservation and Management of Highly Migratory Fish Stocks in the Western and Central Pacific Ocean The WCPT Convention was adopted on 4 September, 2000, and is one of the first agreements to be developed following the conclusion of the UN Fish Stocks Agreement in 1995. The objective of the WCPT Convention is to ensure the long-term and effective conservation and sustainable use of highly migratory fish stocks in the western and central Pacific Ocean, in accordance with LOSC and the UN Fish Stocks Agreement.
Conservation of the marine environment The WCPT Convention deals specifically with highly migratory fish stocks in the western and central Pacific Ocean. However, it outlines some broad provisions that can be applied for the protection of marine ecosystems. These include, inter alia, measures to ensure the long-term sustainability of highly migratory fish stocks; minimization of wastes, discards and other impacts associated with fishing; applying the precautionary principle in implementing the WCPT Convention; protecting marine biodiversity; preventing or eliminating overexploitation of fish stocks; and enforcing conservation measures through effective monitoring, control and surveillance (§5). The WCPT Convention establishes a Commission for the Conservation and
Management of Highly Migratory Fish Stocks in the Western and Central Pacific Ocean to assist in the implementation of the Convention (§9). The functions of the Commission include, inter alia, promoting the sustainable utilization of highly migratory fish stocks in the Pacific; adopting measures for the conservation and management of highly migratory fish stocks, other species and the marine environment in general; and adopting measures to promote responsible fishing in the western and central Pacific (§10).
Convention for the Conservation and Management of Fishery Resources in the South East Atlantic Ocean The SEAFO Convention is a regional agreement covering the southeast Atlantic region. It was concluded on 21 April, 2001, and, like the WCPT Convention, is one of the first post-UN Fish Stocks Agreement treaties. The objective of the SEAFO Convention is to ensure the long-term conservation and sustainable use of the fishery resources in the Convention Area. The SEAFO Convention prescribes principles for conservation and management, which the Contracting Parties are obliged to apply either directly or through the Organization to be established by the Parties. The Parties have a duty to apply the precautionary approach to fisheries management, and adopt, where necessary, conservation and management measures for species belonging to the same ecosystem (§3(d)). Further, Parties must ensure that fishing practices and management measures take due account of the need to minimize harmful impacts on living marine resources as a whole (§3(e)), and protect biodiversity in the marine environment (§3(f)). The SEAFO Convention imposes an obligation on the Commission to apply the precautionary approach widely to conservation management and exploitation of fishery resources to protect the fishery resources and preserve the marine environment (§7(l)).
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Convention on Conservation of Nature in the South Pacific The Convention on Conservation of Nature in the South Pacific (Apia Convention) was developed in Apia, Western Samoa, in June 1976, with the fundamental objective of conserving, utilizing and developing the natural resources of the South Pacific Region through careful planning and management for the benefit of present and future generations. There are certain core provisions in the Apia Convention that address matters such as protected areas, conservation of indigenous species under threat of extinction, customary use of species and areas, and research.
Protected areas The Apia Convention requires Parties to establish protected areas to safeguard, inter alia, representative samples of natural ecosystems and endangered species (§2(1)). Parties are required not to alter the boundaries of national parks within their jurisdiction to either decrease the size of such areas, or to allow the commercial exploitation, collection or hunting of resources contained therein, without first conducting a full investigation (§3(1)–(3)). Additionally, national reserves must be maintained inviolate to the greatest extent possible, although permission for scientific research in reserves may be granted if the purpose of such research is consistent with the purposes for which the reserves were established (§4).
Conservation of indigenous species Parties must not only protect indigenous species in general, but also give particular attention to indigenous and migratory species being exploited in an unsustainable manner or under threat of extinction (§5(1)). In order to achieve the latter, each Party must develop and maintain a list of indigenous species that are threatened with extinction (§5(2)). Species listed accordingly must be protected to the greatest extent possible, and permission to collect, capture or hunt such species may be
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granted only under circumstances that will improve the conservation status of the listed species and their ecosystems (5(3)). Parties must bear in mind the traditions of indigenous communities and make special provisions to enable such communities to use species and areas in accordance with their customs (§6).
Research The Apia Convention requires all Parties to initiate research relating to nature conservation and the management of protected areas and species (§7(2)). Furthermore, Parties are obliged to cooperate in exchanging information and results relating to such research and in interchanging and training personnel for nature conservation objectives (§7(2)).
Agreement for the Establishment of the Regional Commission for Fisheries (RECOFI) The Agreement is a subregional fisheries agreement amongst a number of Arab Gulf States, which establishes the Regional Commission for Fisheries (RECOFI). In the preamble to the Agreement, the Parties note the objectives and purposes stated in Chapter 17 of Agenda 21, and in the FAO Code of Conduct for Responsible Fisheries. The Parties also note other international instruments that have been negotiated concerning conservation and management of certain fish stocks. The functions of the Commission are to promote the development, conservation, rational management and best utilization of living resources (§III(1)). This is to be achieved by keeping under review the state of the resources, including their abundance and the level of their exploitation. The Commission is required to recommend appropriate measures for the conservation and rational management of living marine resources, including regulating fishing methods and fishing gear, prescribing the minimum size for individuals of specified species, and establish open and closed fishing seasons and areas (§III(1)(a) and (b)).
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The Commission is required to apply the precautionary approach to conservation and management decisions, and take into account the best scientific evidence available and the need to promote the development and proper utilization of the living marine resources (§III(2)).
Convention on the Conservation of Antarctic Marine Living Resources The Convention on the Conservation of Antarctic Marine Living Resources (CCAMLR) recognizes the importance of the ecosystems approach to fisheries management. In the preamble, the Parties recognize the importance of safeguarding the environment and protecting the integrity of the ecosystems of the seas surrounding the Antarctic. CCAMLR provides a management system that both protects the ecosystem and allows fishing activities in the southern oceans. It is the first international convention to address ecosystem management goals. The objective of CCAMLR is the conservation of Antarctic marine living resources (§II). Three principles of conservation underpin the objective of CCAMLR. These are to prevent the decrease in the size of harvested populations below unsustainable levels; maintain the ecological relationships between harvested, dependent and related populations of Antarctic marine living resources; and prevent changes in the marine ecosystem that are not potentially reversible over two or three decades (§II(a)–(c)). The functions of the Commission, established under CCAMLR, is to give effect to the objective and principles of the Convention. Conservation and management measures the Commission may adopt include designation of quantities of species that may be harvested; designation of protected species; and the determination of the size, age and sex of species that may be harvested (§IX). The Commission may also take measures to regulate the effects of harvesting and associated activities on components of the marine ecosystem other than the harvested populations (§IX).
Convention for the Establishment of the Lake Victoria Fisheries Organization The Parties to the Convention for the Establishment of the Lake Victoria Fisheries Organization (FAO, 2001c) recognize the continuing need to increase scientific understanding of Lake Victoria, its living resources, its ecosystem, and the impact on those resources of climate, human populations and settlement, non-indigenous wildlife and industrialization. The Convention establishes the Lake Victoria Fisheries Organization (§II). The objective of the organization is to harmonize national measures for the sustainable utilization of the living resources of the lake, and to develop and adopt conservation and management measures. To achieve these objectives, the organization is required to promote the proper management and optimum utilization of the fisheries and other resources of the lake, advise on the effects of the introduction of non-indigenous aquatic animals or plants into the waters of the lake, and to adopt measures regarding the introduction, monitoring, control and elimination of such animals or plants (§II).
Agreement for the Establishment of the General Fisheries Commission for the Mediterranean The Parties to the Agreement for the Establishment of the General Fisheries Commission for the Mediterranean (FAO, 2001d) have a mutual interest in the development and proper utilization of the living marine resources in the Mediterranean and the Black Sea. While the Agreement does not provide specifically for the ecosystems approach to fisheries management, the Commission has an indirect responsibility, which may entail the application of such an approach. The Commission is required to keep under review the state of the resources, including their abundance and the level of their exploitation (§III(1)(a)). Measures which the Commission may recommend for conservation and management include regulating
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fishing methods and fishing gear, prescribing the minimum size for individuals of specified species, establishing closed and open seasons, and regulating the amount of total catch and fishing effort and their allocation among members (§III(1)(b)).
Framework Agreement for the Conservation of Living Marine Resources on the High Seas of the South Pacific – the Galapagos Agreement The objective of the Framework Agreement for the Conservation of Living Marine Resources on the High Seas of the South Pacific – the Galapagos Agreement – is the conservation of living marine resources in the high seas zones of the southeast Pacific, with special reference to straddling and highly migratory fish populations (§2). The Galapagos Agreement specifies a number of conservation principles that have an impact on an ecosystems approach to fisheries management (§3). §5(1)(c) of the Galapagos Agreement provides that in the establishment of conservation measures for regulated species, the effects of fishing for specific fish stocks on the populations of associated or dependent species, as well as on the marine ecosystem as a whole, shall be taken into account. Further, the effects of environmental changes and other phenomena that might affect the marine ecosystem, along with the direct or indirect effects of capture, shall be taken into account, in order to reduce or prevent the risk of potentially irreversible alterations.
The Washington Declaration on Protection of the Marine Environment from Land-based Activities The Washington Declaration on Protection of the Marine Environment from Land-based Activities (the Washington Declaration) was developed in Washington in November 1995 as part of the UN Global Programme of Action for the Protection of the Environment. The primary objective of the Washington
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Declaration is to protect the marine environment from the impacts of land-based activities, and in particular from:
• • • • • • • • •
sewage; persistent organic pollutants; radioactive substances; heavy metals; oils (hydrocarbons); nutrients; sediment mobilization; litter; and physical alteration and destruction of habitats (§1).
The Programme recommends, inter alia, the following measures for protecting the marine environment from the impacts of land-based sources:
•
Reviewing national action programmes within a few years and implementing these programmes in accordance with national capacities and priorities. Cooperating to undertake capacity building activities and mobilizing resources for developing and implementing the programme, in particular, for countries in need of assistance. Undertaking measures to prevent and/or mitigate impacts on the marine environment resulting from land-based activities. Promoting access to knowledge, expertise and cleaner technologies to address land-based activities that impact upon the marine environment. Promoting measures to address the consequences of sea-based activities that require national and/or regional land-based actions, such as recycling facilities (§2–6 and 14).
•
• • •
National Policies for Protecting the Marine Environment International efforts to apply the ecosystems approach to fisheries management are complemented by national policies. The following section discusses some of the more recent national policies adopted by a number of countries.
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Australia’s Oceans Policy Australia’s Oceans Policy (AOP) was released in December 1998 by the Federal Government of Australia (Environment Australia, 1998a,b) as an initiative to promote and facilitate the development of an integrated and ecosystem-based approach to ocean management and conservation. The AOP has several objectives, including the protection of Australia’s marine biodiversity and the ocean environment, and to ensure that the uses of oceanic resources are ecologically sustainable (AOP, p. 4). Developing and establishing regional marine planning for large marine ecosystems is a key aspect of the AOP’s focus on facilitating an integrated and ecosystem-based oceans planning and management framework. The first Regional Marine Plan (RMP) will be developed for the southeastern region of Australia’s EEZ. The AOP establishes the following arrangements to assist with the implementation of the Policy:
• • • •
A National Oceans Ministerial Board, comprising key Commonwealth Ministers, to be chaired by the Minister for the Environment. A National Oceans Advisory Group, comprising industry, community and government stakeholders. Regional Marine Planning Steering Committees, comprising regional stakeholders. A National Oceans Office within Environment Australia, to provide secretariat and technical support for oceans policy initiatives (AOP, p.15).
There are two sections to the AOP. The first, and main part, of the policy document outlines broad goals and planning and management principles to guide the development of an integrated oceans management framework. This part also provides a brief overview of issues and measures related to different uses of oceans, including fisheries management, by-catch and environmental impact assessment. The AOP identifies, inter alia, overcapacity and excess effort in fisheries as major cause of overfishing and in reducing
the viability of marine species populations (AOP, p. 25). Moreover, the Policy states that measures to reduce overcapacity in domestic fisheries will be pursued to ensure that fishing efforts do not exceed ecologically sustainable levels(AOP, p. 25). The second part of the Policy outlines specific measures on a sectoral basis, including fisheries. This section of the Policy identifies key issues and challenges for the fisheries industry, including:
• • • • • •
Fisheries management. Ecologically sustainable fisheries practices. Economic and regulatory instruments (to improve the sustainability and efficiency of fisheries operations). Structural adjustments (to make the fisheries industry more viable while protecting the marine environment). Recreational and charter fishing management. Illegal fishing and compliance with conventions.
The Policy essentially states that the Federal Government, in collaboration and consultation with all stakeholders, will address, inter alia, the aforementioned issues in order to improve the efficiency and economic and environmental or ecological viability of Australia’s fisheries, as well as to protect the marine environment from the environmental impacts of fisheries (and other industries). Examples of measures suggested by the Policy for the fisheries industry include:
• • • •
•
Reviewing fisheries laws and regulations to streamline procedures and reduce compliance costs. Undertaking strategic environmental impact assessments of all new management plans for Commonwealth fisheries with and without management plans. Continuing to address the issue of by-catch in fisheries. Continuing to develop and implement policies for ecologically sustainable fisheries through, for example, precommercial assessments and experimental fisheries. Encouraging the adoption and use of codes of responsible fishing practices.
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Undertaking research and development measures to develop management initiatives to ensure ecologically sustainable fisheries. Continuing to develop national and international strategies to address illegal fishing, non-compliance with conventions and enforcement.
Canada’s Oceans Act The Oceans Act of Canada (OA) received Royal Assent in the House of Commons on 18 December, 1996, and entered into force on 31 January, 1997. The preamble to the OA underscores the approach Canada has adopted with regards to the protection of its marine environment. The OA is intended to ‘promote the understanding of oceans, ocean processes, marine resources and marine ecosystems to foster the sustainable development of the oceans and their resources’. Further, the OA underlines that ‘conservation based on an ecosystem approach is of fundamental importance to maintaining biological diversity and productivity in the marine environment’. OA has three core sections. The first part establishes a legal demarcation of Canada’s waters, including a Contiguous Zone (CZ) and an EEZ. The first part also outlines Canada’s rights and responsibilities with respect to its EEZ. These responsibilities essentially enable Canada to explore, exploit and conserve the natural living and non-living marine resources of the EEZ, as well as to regulate scientific research and control the construction of offshore structures in the EEZ (§13–14). The second part of the OA establishes certain obligations for the Minister of Fisheries and Oceans with respect to the management and conservation of Canadian waters (§30–35). The third part of the OA outlines the powers, duties and functions of the Minister of Fisheries and Oceans (§40–52.1).
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and protection of Canada’s waters. The Act establishes a legal framework for the development and implementation of a national strategy for the management of estuarine, coastal and marine waters within Canadian jurisdiction (§29). Under the OA, the Minister of Fisheries and Oceans (in collaboration with all relevant stakeholders) must develop and implement the national strategy, as well as facilitate the development and implementation of an integrated management plan for managing all activities and measures that affect Canada’s estuarine, coastal and marine waters (§31–32). The Act also enables the Minister to establish marine protected areas specifically for the conservation and protection of marine ecosystems and biodiversity, threatened or endangered species and their habitats (§35). The OA also enables the Minister to establish advisory or management bodies for the purposes of implementing integrated management plans for Canadian waters (§32(c)(i)). In performing the duties assigned by the OA, the Minister must consult and collaborate with all appropriate stakeholders (§3(2)).
EU Green Paper on the Future of the Common Fisheries Policy The EU Green Paper on the Future of the Common Fisheries Policy (EC, 2001a) addresses changes to approach by the European Union towards oceans management. The Policy advocates the implementation of multi-annual frameworks that integrate the precautionary approach, and the establishment of medium-term environmental and ecosystem objectives and strategies for key species and habitats (EC, 2001a, p. 22). The Common Fisheries Policy is supplemented by the Biodiversity Action Plan for Fisheries (EC, 2001b). The key areas for the purpose of this chapter identified in the Action Plan are:
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Protection of the marine environment As mentioned earlier, the second part of the OA creates obligations for the management
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The promotion of the conservation and sustainable use of fish stocks and feeding grounds through control of exploitation rates and through the establishment of technical conservation measures to
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support the conservation and sustainable use of fish stocks. The reduction of the impact of fishing activities and other human activities on non-target species and on marine and coastal ecosystems to achieve sustainable exploitation of marine and coastal biodiversity (EC, 2001b, p. 6).
The foregoing discussion illustrates that the trends at international, regional and national level point towards an increasing application of the ecosystems approach in fisheries management.
Conclusion In analysing the strengths and weaknesses of international efforts to incorporate ecosystem management principles into international instruments, the most notable strengths of the international instruments studied for this chapter are the instruments themselves, as they attempt to establish a global framework for the conservation and management of marine environments and resources. Moreover, the inclusion of ecosystem conservation is also a positive element, as it is a step away from the traditional species and stock focuses. This ecosystem-based focus also provides scope for an increased involvement of regional bodies in establishing integrated marine and coastal management measures. The past two decades has been characterized by the realization that only a holistic approach towards fisheries management can ensure the proper governance of the oceans. The implementation of the new international instruments will strengthen fisheries management and ocean governance. As one international expert said: These instruments – and particularly the Code of Conduct for Responsible Fisheries – contain a number of provisions referencing broader environmental goals, for example, the need for consideration of aquatic ecosystems, the need for an ecosystem approach to management, and the need to minimize by-catch, pollution, waste and discards. As such, they illustrate what we believe is essentially a ‘paradigm shift’ in international
fisheries, which flows from the increasing recognition of the nexus between international fisheries law and international environmental law. This paradigm shift involves growing recognition of two requirements for the sustainable conservation and management of capture fisheries. The first reflects ecosystem concerns. There must be effective steps to provide for the health, not only of populations of target species, but also of non-target species, and we must take steps to maintain relationships among species. Second, there must be effective steps to protect fisheries habitat. In particular, we must protect fisheries habitat threatened by adverse impacts stemming from human activities, including harmful fishing practices, and other activities, which are increasingly concentrated in coastal areas. In effect, fisheries managers must concern themselves with the entire marine ecosystem. (West, no date)
There are, however, several weaknesses that need to be considered. One of the major drawbacks of international instruments is that many States are not party to them, thereby limiting the extent to which these instruments are being applied. The provisions outlined in instruments are often vague and ambiguous with respect to the protection of the marine environment, and these need to be addressed in order to assert more clearly environmental protection obligations to States. Even though many of the instruments include illegal, unregulated and unreported (IUU) fishing, surveillance and enforcement as key issues to be addressed, it will be difficult, or even impossible, to control these problems through comprehensive and effective monitoring of an area so vast. Moreover, developing nations, in particular, will be hard pressed to find sufficient resources to implement many of the measures outlined in the international instruments. The international instruments studied for this chapter clearly include ecosystem management as a key element in protecting the marine environment and its resources. The effectiveness of those instruments, however, remains to be seen. It is clear that a lot of work still needs to be done. At a recent meeting of Environment Ministers from Asia and the
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Pacific, the Ministers issued the following Statement: We express our concern that the integrity of the coasts and oceans is under threat from unsustainable development and overexploitation. The discharge of hazardous and toxic wastes, land-based sources of pollution, the destruction of corals and mangroves, offshore oil drilling and mineral exploration and exploitation, oil spills, marine accidents, excessive coastal tourism and overfishing have been identified as some of the main causes of marine environmental degradation. We call for a renewed commitment to sustainable development of oceans and coastal resources through effective cooperation among national, subregional, regional and international institutions responsible for marine and ocean protection and management; the implementation of national, subregional and regional policies for enhancing sustainable management and uses of oceans and their resources; and the promotion of total ecosystem marine resources management through capacity building . . . (ESCAP, 2001)
References CCAMLR (Convention on the Conservation of Antarctic Marine Living Resources) (1980) http://www.oceanlaw.net/texts/ccamlr.htm EC (European Commission) (2001a) The Future of the Common Fisheries Policy. (20.3.2001 COM (2001) 135 final). Commission of the European Communities, Brussels. EC (2001b) Biodiversity Action Plan for Fisheries. (27.3.2001 (COM (2001) 162 final Vol. IV). Commission of the European Communities, Brussels. ESCAP (United Nations Economic and Social Commission for Asia and the Pacific) (2001) Phnom Penh Regional Platform on Sustainable Development for Asia and the Pacific. Commission on Sustainable Development, acting as the Preparatory Committee for the World Summit on Sustainable Development, 28 January–8 February 2002. Environment Australia (1998a) Australia’s Oceans Policy (AOP 1). Department of the Environment, Sport and Territories, Canberra, ACT. Environment Australia (1998b) Australia’s Oceans Policy (AOP 2). Department of the Environment, Sport and Territories, Canberra, ACT.
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FAO (1995) Code of Conduct for Responsible Fisheries. Rome: FAO. (Also as pp. 51–77, in: International Fisheries Instruments with Index. Division for Ocean Affairs and the Law of the Sea, Office of Legal Offices, United Nations, New York.) FAO (2001a) Convention for the Conservation and Management of Fishery Resources in the South East Atlantic Ocean. http://www.fao.org/ Legal/TREATIES/032t-e.htm FAO (2001b) Agreement for Establishment of the Regional Commission for Fisheries (RECOFI). http://www.fao.org/Legal/TREATIES/0282 -e.htm FAO (2001c) Convention for the Establishment of the Lake Victoria Fisheries Organization. http://www.fao.org/Legal/TREATIES/027t -e.htm FAO (2001d) Agreement for the Establishment of the General Fisheries Commission for the Mediterranean. http://www.fao.org./Legal /TREATIES/003t-e.htm FFA (Forum Fisheries Agency) (2000) Convention for the Conservation and Management of Highly Migratory Fish Stocks for the Western and Central Pacific Ocean. Forum Fisheries Agency, Honiara. Fluharty, D.L. (2001) Transition to Ecosystem-based Fisheries Management. White Paper prepared for Seminar on Ocean Governance, 16 November 2001, Seattle, Washington. Working Paper of the School of Marine Affairs, University of Washington (unpublished). Also included in submission to the Ocean Commission, US Congress (see proceedings, 19 November 2001). Galapogos Agreement (2000) Framework Agreement for the Conservation of Living Marine Resources on the High Seas of the South Pacific. http://www.oceanlaw.net/texts/ galapagos.htm GFCM (General Fisheries Commission for the Mediterranean). Agreement for the Establishment of the General Fisheries Commission for the Mediterranean. http://www.fao.org/Legal/ TREATIES/003t-e.htm LVFO (Lake Victoria Fisheries Organization). (1994) Convention for the Establishment of the Lake Victoria Fisheries Organization. http://www.fao.org/Legal/TREATIES/ 027t-e.htm Mehan, G.T., III. (No date) Ecosystem Management in the Great Lakes Basin. http:// www.nmu.edu/sbp/ecomangt_MG.HTML Oceans Act (1996) Oceans Act 1996. Department of Justice, Ontario. http://laws.justice.gc.ca/ en/0-2.4/78384.html
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SEAFO Convention (2001) Convention for the Conservation and Management of Fishery Resources in the South East Atlantic Ocean. http://www.fao.org/Legal/TREATIES/032t -e.htm SPREP (South Pacific Regional Environment Programme). (2000) Convention on Conservation of Nature in the South Pacific. SPREP, Samoa. Tsamenyi, M. and McIlgorm, A. (1999) International Environmental Instruments: Their Effect on the Fishing Industry. University of Wollongong and AMC Search Ltd, Wollongong, NSW. UN (United Nations) (1983) The Official Text of the United Nations Convention on the Law of the Sea with Annexes and Index. United Nations, New York.
UN (1992) Convention on Biological Diversity. United Nations, New York. UN (1998) Agreement for the Implementation of the Provisions of the United Nations Convention for the Law of the Sea of 10 December 1982 relating to the Conservation and Management of Straddling Fish Stocks and Highly Migratory Fish Stocks. In: International Fisheries Instruments with Index. Division for Ocean Affairs and the Law of the Sea, Office of Legal Offices, United Nations, New York, pp. 1–37. UNGA (United Nations General Assembly). (2001) Oceans and the Law of the Sea. UNGA Fifty-Sixth Session, 6 March 2001. West, M.B. (No date) US Deputy Assistant Secretary of State: State Department Official on Fisheries Management. http:// www.usembassy.state. gov/posts/jal/wwwt2620.text
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Incorporating Ecosystem Considerations into Fisheries Management: Large-scale Industry Perspectives Bernt O. Bodal American Seafoods Group, Seattle, Washington, USA
Abstract Internationally, the large-scale industry is a diverse group of both shore-based and at-sea harvesting and processing operations. As with all elements of the fishing industry, the performance of large-scale fisheries is controlled by various degrees of governmental and institutional constraints. The record shows that the degree of responsible fisheries practised in any sector of the fishing industry largely depends upon the level of responsibility within government and regulatory institutions, and a commitment to responsible fisheries by the fishing industry. There are a number of reasons why some fisheries have attracted larger vessels, such as remote fishing grounds, the large size of the resource, the perishable nature of the fish, the need for capital-intensive production equipment and the harsh and dangerous fishing conditions. In this production environment, only large-scale fisheries are able to deliver seafood at cost-effective prices. Without the economies of scale of the large-scale seafood industry, this healthy source of protein would either be left in the water or be affordable only to the wealthy. Greenpeace and other non-governmental organizations (NGOs) have repeatedly attacked the large-scale sector as unsustainable and of ‘strip mining’ the seas. However, in the North Pacific under USA jurisdiction, the facts dispute this notion. Fisheries in this region are widely regarded as some of the most responsibly and conservatively managed fisheries anywhere in the world. With a track record of nearly 30 years of commercial fishing activities, none of the 63 species of groundfish in the USA North Pacific are classified as overfished or even approaching the overfishing level. Bering Sea pollock, the largest fishery in the USA, currently is at a high biomass level, 10 million t. The allowable harvest rate of Bering Sea pollock in 2001 is well below the acceptable biological catch of 1.85 million t, and about half of the maximum sustainable yield (MSY) harvest rate. The primary reason that these fisheries are healthy and sustainable is due to the responsible application of the precautionary approach in the calculation of quotas and in the overall management of the fishery since the inception of the 200 mile exclusive economic zone (EEZ) in the late 1970s. In addition to precautionary levels of allowable catch, harvests are monitored closely and reported on an ongoing basis. In the Alaskan pollock and Pacific whiting fisheries, the large-scale fleets are required to have two federal fishery observers aboard at all times, who collect fishery data on 99% of all hauls. One hundred percent of all fish caught are weighed on flow scales, and catch data is reported daily to the National Marine Fisheries Service (NMFS), the agency responsible for in-season monitoring of the fishery. Both regulatory and voluntary
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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by-catch controls are important tools that have been employed for over 20 years. The large-scale fleet in the North Pacific has the ability to respond rapidly to changes in by-catch and is able to relocate to areas of lower by-catch. This is demonstrated by an overall by-catch rate of 0.6% in the pollock fishery, the lowest of the world’s major fisheries. Management in the USA North Pacific has implemented marine protected areas to protect habitat. In an effort to protect fish and crab habitat in the eastern Bering Sea, areas closed to bottom trawling encompass 30,000 square miles, or about 25% of the available fishing area. Other ecosystem principles employed include prohibitions on fishing for forage fish stocks in the North Pacific, to protect these important prey species for seabirds and marine mammals. Further, NMFS conducts research on and manages not only targeted fish stocks but also non-targeted species of fish, seabirds, and marine mammals, and takes into consideration the inter-relationships between these species and the physical and chemical forces of the marine environment. The large-scale fleets in the Pacific Northwest and Alaska have been supportive of conservative ecosystem-based management. They are all aware that their economic viability is dependent on sustainable resources, and hence they share a long-term commitment to healthy resources. Recent changes in USA law have allowed the large-scale sector to pursue new avenues, such as harvesting cooperatives, in which quotas are assigned to vessels, thus ending the race for fish. In an era when most fisheries throughout the world are heavily overcapitalized, managing harvesting effort with Olympic-style quotas, where vessels must compete against each other as frantically as possible, waste and inefficiency are all too common. In certain fisheries, harvesting cooperatives have proven to be far superior to Olympic quotas as a management tool. Cooperatives have led to reductions in by-catch, while at the same time providing increased recovery of processed seafood product: an impressive 36% increase in the pollock fishery. Harvesting cooperatives also result in spreading catch effort more evenly over space and time, decreasing the potential for localized depletion of resources. Because cooperatives allow for individual accountability, and hence a meaningful role in managing the resource, their members are willing to support, both logistically and financially, scientific research to improve resource assessments, increased monitoring and testing of innovative fishing practices. For instance, the Pollock Conservation Cooperative contributes US$1.4 million annually to fisheries research. In the USA North Pacific, the large-scale fishing industry, and American Seafoods, are very supportive of good scientific information and understand that sustainable fisheries, such as the eastern Bering Sea pollock fishery, are only possible with good data on stock status and fishery removals. Integrating additional ecosystem data into existing fishery management plans is an ongoing process and will require careful and comprehensive analysis. However, in many parts of the world, this is already being done, and these efforts should continue as long as clear, measurable benefits to the environment and stakeholders can be demonstrated. With the right incentives, the fishing industry can provide positive, creative energy for responsible management practices and fishery research.
Introduction and Background I am Chairman and Chief Executive Officer of the American Seafoods Group and have been involved in the seafood industry for the last 25 years. For 17 years, I was the captain of crab and groundfish vessels off the coast of Alaska. From 1994 to 1998, I was President of American Seafoods Company and was instrumental in helping to grow the company and its affiliates into a fleet of over 30 vessels, with offices in the USA, Argentina, Chile and Russia.
Today, the American Seafoods Group comprises the American Seafoods Company and American Seafoods International. American Seafoods Company owns and manages seven USA-flag catcher-processors that operate in the groundfish fisheries in the USA North Pacific. American Seafoods is a founder member of both the Pollock Conservation Cooperative and the Pacific Whiting Conservation Cooperative. American Seafoods International is a value-added processor of retail seafood products, which include the Frionor,
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Bayside Bistro and Arctic Cape product lines. Our customers include every segment of the food service market, from restaurants and cafeterias to schools, health care facilities, business and industry, hotels, catering, country clubs and cruise lines. What I have been asked to speak about today is the perspective of large-scale fishing interests regarding the incorporation of ecosystem considerations into fisheries management. I would like to focus on the following: 1. What the large-scale industry is, why it exists and its importance to the rest of the industry and the economy. 2. Some examples of successful fisheries management systems that already apply the precautionary approach and consider ecosystem principles, with demonstrable and impressive results. 3. Suggest harvesting cooperatives as alternatives to the traditional quota management systems that have had such a poor track record over the years. 4. Distinguish between nearshore and offshore ecosystems, and the differential impact of different types of fisheries on each. 5. The impacts of incorporating ecosystem considerations into fisheries management.
Reasons for Large-scale Fishing Operations Large-scale fishing operations play an important role in fishing communities throughout the world. They are a diverse group of both shore-based and at-sea harvesting and processing operations. There are a number of reasons why some fisheries have attracted larger vessels.
• • • • •
Remoteness of fishing grounds. Harsh and dangerous fishing conditions. The large size of the resource. The perishable nature of the catch. The need for capital-intensive production equipment.
In this production environment, only large-scale fisheries are able to deliver seafood at cost-effective prices.
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Importance of the Large-scale Seafood Industry Of all the protein produced for human consumption, seafood is recognized more and more as the best source of protein for human health. Without the economies of scale of the large-scale seafood industry, this healthy source of protein would either be left in the water or would be affordable only to the wealthy. The large-scale sector also provides employment for a significant portion of workers in the seafood industry, and is an important source of trade for countries.
The Record in the USA North Pacific As with all elements of the fishing industry, the performance of large-scale fisheries is controlled by a variety of institutional constraints. Healthy, responsibly managed fisheries in any sector of the fishing industry are dependent upon the degree of responsibility practised by government and regulatory institutions, and on the commitment of the fishing industry. Greenpeace and other non-governmental organizations (NGOs) have repeatedly attacked the large-scale sector as unsustainable and ‘strip mining’ the seas. However, at least in fisheries conducted in the USA North Pacific, the facts do not support these claims. We are extremely proud of our management system in the area. We believe it is one of the finest examples of ecosystem-based management anywhere in the world, one that is widely regarded for its responsible and conservative fisheries management. With a track record of nearly 30 years of commercial fishing activities, none of the 63 species of groundfish in the USA North Pacific are classified as overfished or even approaching the overfishing level. Bering Sea pollock, the largest fishery in the USA, currently is at a high biomass level of 10 million t. The allowable catch of Bering Sea pollock in 2001 is 1.4 million t, well below the acceptable biological catch of 1.85 million t, and about half
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the maximum sustainable yield (MSY) level of harvest.
Elements Contributing to Sustainable Fisheries in the North Pacific The primary reason that these fisheries are healthy and sustainable is the responsible application of the precautionary approach in the calculation of quotas and in the overall management of the fishery since the inception of the 200-mile exclusive economic zone (EEZ) in the late 1970s. In addition to precautionary levels of allowable catch, harvests are monitored closely and reported on an ongoing basis. In the Alaskan pollock and Pacific whiting fisheries, the large-scale fleets are required to have two federal fishery observers onboard at all times, who in 2000 collected fishery data on 99% of all hauls. One hundred per cent of all fish caught are weighed on flow scales, and catch data is reported daily to the US National Marine Fisheries Service (NMFS), the agency responsible for in-season monitoring of the fishery. To be effective, total harvest (including discarded fish) must be monitored and enforced to prevent exceeding sustainable levels. Both regulatory and voluntary by-catch controls are important tools that have been employed for over 20 years. The large-scale fleet in the North Pacific has the ability to respond rapidly to changes in by-catch and is able to relocate to areas of lower by-catch. This is demonstrated by an overall by-catch rate of 1.6% and an overall discard level of 0.6% in the pollock fishery, the lowest level recorded among the world’s major fisheries. Management in the USA North Pacific has implemented marine protected areas to protect habitats. In an effort to protect fish and crab habitats in the eastern Bering Sea, areas closed to bottom trawling encompass 30,000 square miles, or about 25% of the available fishing area. Some environmental groups have suggested that turning the entire Aleutian Islands into a marine protected area would not significantly reduce access to
fishing grounds. However, to put this in perspective, the Aleutian Islands span a distance in excess of 1600 km. This is roughly as long as the western USA coast, or the west coast of Norway. I think most people would consider such a large area significant! Other ecosystem principles employed include prohibitions on fishing for forage-fish stocks in the North Pacific, to protect these important prey species for seabirds and marine mammals. Further, NMFS conducts research on and manages not only targeted fish stocks but also non-targeted species of fish, seabirds and marine mammals, and takes into consideration the inter-relationships among these species and the physical and chemical forces of the marine environment.
Industry Support of Fishery Management The large-scale fleets in the Pacific Northwest and Alaska have been supportive of conservative, ecosystem-based management. They understand that their economic viability depends on sustainable resources, and hence share a long-term commitment to healthy resources.
Harvesting Cooperatives Recent changes in USA law have allowed the large-scale sector to pursue alternatives to Olympic quotas, such as harvesting cooperatives in which quotas are assigned to individual vessels, thus ending the race for fish. In an era when most fisheries throughout the world are heavily overcapitalized, Olympicstyle quotas force vessels to compete against each other as frantically as possible, which results in waste and inefficiency. In certain fisheries, harvesting cooperatives have proven to be far superior to Olympic quotas as a management tool. Harvesting cooperatives in the USA have resulted in voluntarily reducing the size of fishing fleets without the need for government-funded buy-back programmes. Cooperatives have
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led to reductions in by-catch, while at the same time providing increased recovery of processed seafood product, an impressive 35% increase in the catcher-processor pollock fishery. This means that, by comparison, for every 100 t of product produced in the Olympic-style fishery, now 135 t of product is produced under the cooperative system. Harvesting cooperatives also help to spread out catch effort more evenly over space and time, decreasing the potential for localized depletion of resources. Further, because cooperatives allow for individual accountability, fishermen have a stake in assuring sustainable resources. Cooperative members are willing to support, both logistically and financially, scientific research to improve resource assessments, increased monitoring and testing of innovative fishing practices. For instance, the Pollock Conservation Cooperative contributes US$1.4 million annually to fisheries research, and the Whiting Cooperative has sponsored scientific resource surveys since 1998.
Large-scale Fisheries, Nearshore Ecosytems Large-scale fisheries do not necessarily have greater impacts on the ecosystem than those of smaller scale fisheries. Some fisheries, such as pollock, typically occur farther offshore, with effort spread out over time and area. The USA pollock fishery harvests with pelagic gear only, so there are no negative impacts to the ocean floor habitat. The incidental harvest of non-target species, or by-catch, is negligible. Fisheries with fewer and larger vessels facilitates highly effective monitoring and enforcement. In contrast, many small-scale fisheries take place close to shore. Nearshore marine ecosystems are fragile and generally more easily affected than offshore ecosystems. Breeding sites for seabirds and marine mammals occur in these nearshore environments, and species interactions are often more complex than they are farther offshore. Small vessels are limited to fishing close to shore
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and, over time, could cause localized depletions and disturbance of nearshore organisms. A good example of this is the foraging habits of Steller sea lions in Alaska. Satellite monitoring data indicate that the vast majority of these foraging trips take place within 10 miles of nearshore rookeries and haul-outs in Alaskan waters. Based on these data, the nearshore, small-boat fisheries could be more of a threat to the continued survival of Steller sea lions than are the large vessel fisheries operating farther offshore. Further, small boats are less able to accommodate independent observers to monitor catches and collect scientific data. As we all know, poor data lead to poor resource management.
Impacts of Incorporating Ecosystem Considerations Management actions based upon new knowledge about the ecosystem should not be taken frivolously and should only be implemented when clear, measurable benefits to the environment can be demonstrated. The precautionary approach has two meanings. Most people think about being more cautious in the face of uncertainty, but precaution also means not taking action when we are unsure about whether or not that action could do more harm than good. Inevitably, conflicts between different components in the ecosystem will arise, and actions will affect these components differently. At some point, determining the value of, and prioritizing, these components is unavoidable, and there are no easy answers. However, there is one strategy that clearly does not work in addressing these conflicts. NGOs will need to change their relationship with industry and government agencies, from adversarial to collaborative. Lawsuits interfere with legitimate research, which can result in more harm to resources than good. If the millions of dollars each year that NGOs spent on lobbying and lawsuits were spent instead on research, we could solve many of the most serious problems threatening marine ecosystems today.
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Conclusion To be successful, seafood businesses of the future must ensure that natural resources are managed to be sustainable and productive over the long term. Since our very livelihood and future is heavily dependent on healthy fishery resources, nobody has a stronger commitment to sustainable fishery management than does the fishing industry. In the USA North Pacific, the large-scale fishing industry, and American Seafoods, are supportive of good scientific information, and understand that sustainable fisheries – such as the eastern Bering Sea pollock fishery – are only possible with good data on stock status and fishery removals, and an understanding of their effects on the ecosystem. We have found that harvesting cooperatives are an effective alternative to traditional fishery management systems, allowing greater individual accountability and slowing down the frantic race for the fish. Integrating additional ecosystem data into existing fishery management plans is an ongoing process that requires careful and comprehensive analysis. Conclusions should
not be reached hastily and should be supported by scientific facts. However, in many parts of the world, this is already being done and these efforts should continue as long as clear, measurable benefits to the environment and stakeholders can be demonstrated. Remember, the precautionary principle has two meanings. Most people think about being more cautious in the face of uncertainty. Precaution also means not taking action when we are unsure whether an action could actually do more harm than good. Industry has a vital and active role in supporting research and management efforts. Indeed, without input from and cooperation with stakeholders, incorporating ecosystem principles into fishery management is doomed to fail. Fisheries scientists should seek and encourage this cooperative relationship with industry. Any fishery management system must carefully balance sound conservation objectives with the interests of fishermen and fishing communities. With the right incentives, the fishing industry can provide positive, creative energy to responsible management practices and fishery research.
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Small-scale Fisheries Perspectives on an Ecosystem-based Approach to Fisheries Management Sebastian Mathew International Collective in Support of Fishworkers (ICSF),1 Chennai, India
Humanity stands at a defining moment in history. We are confronted with a perpetuation of disparities between and within nations, a worsening of poverty, hunger, ill health and illiteracy, and the continuing deterioration of the ecosystems on which we depend for our well-being. However, integration of environment and development concerns and greater attention to them will lead to the fulfilment of basic needs, improved living standards for all, better protected and managed ecosystems and a safer, more prosperous future. No nation can achieve this on its own; but together we can – in a global partnership for sustainable development. (Para 1.1, Preamble, Chapter 1, Agenda 21 Programme of Action for Sustainable Development)
Abstract In 1992, UNCED Agenda 21 highlighted the protection and preservation of highly diverse marine ecosystems and the problems that degraded ecosystems posed to marine fishing activities. The 1995 UN Fish Stocks Agreement referred to the need to maintain the integrity of ecosystems and to consider problems posed by fishing and degrading ecosystems. Further, the 1995 FAO Code of Conduct for Responsible Fisheries gave greater significance to an ecosystem-based approach to fisheries management. Artisanal and small-scale fisheries are accorded special recognition by the Code of Conduct for Responsible Fisheries. Such fisheries contributed more than a quarter of world catch, and accounted for half of the fish used for direct human consumption. Individually, small-scale fishing units are less threatening to the marine ecosystem than are largescale ones, because they participate in a multi-species fishery with low quantities of gear that are often passive and selective, and in accordance with the fisheries resources that are seasonably accessible to their gear. With the widespread adoption of motorization, small-scale fisheries have grown significantly over the past two decades. The rapid expansion of artisanal fishing capacity under open access regimes has begun to exert overfishing pressures on coastal fisheries resources, especially in Asia and Africa. There are increasing conflicts between different gear groups as a result of increased mobility of fishing vessels, capacity expansion and overfishing pressures.
1 Disclaimer: the views expressed in this paper do not necessarily represent the official position of the International Collective in Support of Fishworkers (ICSF).
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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In the present scenario, there is an urgent need for the State to take up fisheries management measures for greater equity and sustainability through consultative mechanisms. In this context, greater recognition should be given to small-scale rather than large-scale fisheries. The emphasis has to change: away from increasing fish production, toward conservation and management goals. To initiate fisheries management measures in developing countries, a ‘crossword’ approach could be considered, i.e. filling up management niches that are relatively easy at first, and then moving to more difficult ones with the aid of early breakthroughs or solutions. There could also be global initiatives towards fisheries management in developing countries. Industrialized countries, in the first place, should not transfer their excess fishing capacity to developing countries. There is also a need to establish a well-designed, time-bound, international aid programme in exchange for a commitment to manage fisheries in a consultative, transparent and sustainable manner. For small-scale fisheries that are overcrowded in developing countries, industrialized nations could contribute to alleviating such demographic pressure in fisheries by facilitating temporary migration of surplus labour into their fisheries, particularly into fisheries that are earmarked by labour shortage. Concurrent with proposing and implementing measures that basically address the impact of fishing on fish stocks and the marine habitat, there is also need for measures to minimize the effect of pollutionrelated habitat degradation on fish stocks, and to understand better the intricacies of weather and climate factors. Programmes to conserve ‘charismatic’ species such as sea lions, dolphins and sea turtles also sometimes become counterproductive when these resources multiply in large number and compete with fishers for the quarry, without significantly contributing to the health of the marine ecosystem. Unlike the single-species model in fisheries management, which is by far the most prominent model in most parts of the world, an ecosystem-based approach to fishery management could be an effective tool in developing countries since it could take into account the complexity of the marine and coastal ecosystems. A universally acceptable definition of ecosystem-based fishery management, however, has to consider fishers as part of the ecosystem, which is an important consideration for developing countries that have 95% of the world’s fisher population and over 60% of the world’s marine fisheries resources. An ecosystem approach has to be used in a dialectical sense. It should, on the one hand, take into account the effects of fishing on fish stocks, especially the unequal impact of small- and large-scale fishing on targeted fish stocks and the marine and coastal ecosystems, undertaken under different economic, social and political milieu. On the other hand, it should also take into account the effects of marine ecosystems, and alternative livelihoods for fishers. This would be within the framework of what could be considered as an ecosystem-based approach to fisheries management indicated in Agenda 21 and the UN Fish Stocks Agreement.
Introduction As far back as half a century ago, at the United Nations Technical Conference on the Conservation of the Living Resources of the Sea, in Rome in 1955, the components of an ecosystem approach to fisheries management were already recognized. The Conference observed that conservation measures for the living marine resources were to be in accordance with the ‘maintenance of the existing ecological system in a given maritime zone’ (UN, 1955, p. 45). Further, the 1982 UN Convention on the Law of the Sea (LOSC) made a mention of
‘ecosystem’ in the context of protecting and preserving rare and fragile ecosystems (UN, 1983, Article 194(5)). The impact of degraded ecosystems on fishing was recognized internationally at the United Nations Conference on Environment and Development (UNCED). In the 1992 Agenda 21 of UNCED, protection and preservation of highly diverse marine ecosystems and the problems posed by degraded ecosystems to marine fishing activities were recognized. The impact of fishing on degrading ecosystems was highlighted at the 1995 UN Fish Stocks Agreement. It referred to the need for
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maintaining the integrity of ecosystems and to consider the problems that fishing posed in degrading ecosystems, noting the need ‘to assess the impacts of fishing . . . on target stocks and species belonging to the same ecosystem’. (UN Fish Stocks Agreement, 1995, Article 5(d)). Further, the 1995 FAO Code of Conduct for Responsible Fisheries gave great significance to an ecosystem-based approach to fisheries management. At the national level, in the aftermath of the cod crisis in the early 1990s, Canada seems to be the first country in the world to have adopted an explicit ecosystem approach to fisheries management. According to the 1997 Oceans Act, ‘. . . Canada holds that conservation, based on an ecosystem approach, is of fundamental importance to maintaining biological diversity and productivity in the marine environment’. The USA has also recently (January 2001) adopted an ecosystem-based fisheries management plan, namely a plan for the coral reef ecosystems in the Western Pacific. The EU is in the process of implementing ‘ecosystem-oriented’ management in Community waters from 2002 as part of the Common Fisheries Policy (EC, 2001). Further, several Organisation for Economic Co-operation and Development (OECD) countries seem to be in the process of adopting an integrated ecosystem approach to management of fisheries (OECD, 2001). There are, however, no developing countries that have adopted, or are in the process of adopting, such an approach. What is an ecosystem, or an ecosystembased approach to fisheries management? According to the Fisheries Resource Conservation Council of Canada (FRCC), which has been involved since 1993 in finding a solution to the dramatic declines in the Atlantic Canadian groundfish stocks, ‘the term [ecosystem approach] implies trying to manage our own participation in the system with a fuller understanding of its processes and our effects upon them’ (FRCC, 1996). It further observes, ‘In the broadest sense, ecosystem management in the ocean means managing the behaviour of people (chiefly their fishing) in order to maintain or restore desired levels of diversity, abundance and productivity in the
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ocean system . . .’ (FRCC, 1996). The FRCC definition thus seems to draw a distinction between ecosystems and the people who interact with them. The National Research Council of the USA (NRC) defines the ecosystem-based approach to fisheries management as ‘an approach that seriously takes all major ecosystem components and services – both structural and functional – into account in managing fisheries and one that is committed to understanding larger ecosystem processes for the goal of achieving sustainability in fishery management’ (NRC, 1999). It advocates recognizing humans as ‘components of the ecosystems they inhabit and use’, and cautions against dividing the world into ‘the ecosystem’ and ‘the users of the ecosystem’, as implied in the FRCC approach. A third definition is from the Report to Congress of the USA by the Ecosystem Principles Advisory Panel (EPAP), set up by the USA National Marine Fisheries Service (NMFS). An ecosystem-based approach to fishery management, according to the EPAP, should take into account four aspects: (i) the interactions of a targeted fish stock with predators, competitors and prey species; (ii) the effects of weather and climate on fisheries biology and ecology; (iii) the interactions between fish and their habitat; and (iv) the effects of fishing on fish stocks and their habitats, especially how the harvesting of one species might have an impact upon other species in the ecosystem (NMFS, 1999). None of the above definitions make a distinction between ‘small’ and ‘large’; the significance of the scale of fishing operations is not an issue that is dealt with, although the effects of fishing on fish stocks and fish habitats are bound to vary between smalland large-scale fishing units. Most small-scale fishing units can be individually less threatening to the marine ecosystems because they often participate in a fishery with low quantities (and greater diversity) of gear that are often passive and selective, and in accordance with the fisheries resources seasonably accessible to their gear. As has been pointed out by the International Conference of Fishworkers and their
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Supporters, in Rome in 1984, small-scale units are also capital and fuel-saving, and in the tropical belt they are better adapted to the aquatic ecosystems (DAGA, 1984).
Marine Fisheries in Developing Countries The contribution of developing countries to world marine fish production in 1998 was 60%. Of the top seven fish-producing countries in the world in 1998, five are developing countries (FAO, 2000a). Three of them (China, India and Indonesia) have a huge population of nearly one billion people living below the UNDP income poverty line of US$1 day−1 (UNDP, 1999). Artisanal, smallscale fisheries contributed more than 25% of the world catch, and accounted for half of the fish used for direct human consumption (FAO, 1998). What is most significant about the contribution of small-scale fisheries to world fish production is that it has been achieved in spite of receiving very few subsidies from governments, and insignificant development assistance from the international aid community. According to an FAO estimate (FAO, 2001) there are about 36 million fishers – people involved in fishing and fish farming – in the world, of which 80% are in Asia. Sixty per cent of the global population of fishers are in marine capture fisheries, 25% in inland and marine aquaculture and the remainder in inland capture fisheries. The proportion of fishers to total population is highest in Viet Nam and Indonesia – one in every 25 of the population is a fisher in Viet Nam, and one in every 44 in Indonesia. Most of them are employed in artisanal, small-scale fisheries. In absolute terms, Bangladesh, China, India, Indonesia, the Philippines and Viet Nam have the largest number of fishers in the world. For example, Chennai, the capital of the State of Tamil Nadu, India, where the International Collective in Support of Fishworkers (ICSF) is located, alone has an active fisher population of 31,000. In contrast, Iceland and New Zealand together account for less than 12,000 fishers, but their combined fish production, at
2.6 million t (1998 figures), equals the total marine fish production of India. According to FAO, while employment in agriculture in developing countries grew by 35% between 1970 and 1996, employment in fisheries doubled (FAO, 1999), but employment in fisheries in the OECD countries saw a one-third decline in the same period, with the exception of Iceland and Portugal (OECD, 2000). Small-scale fisheries, being an economic activity in the remote areas of many coastal countries, especially in areas where alternative sources of employment are scarce, seem to have played a crucial role in employment creation, income generation and poverty alleviation, arguably because of resilient coastal fisheries, to where people migrate from other, less rewarding occupations, or from occupations that cannot guarantee a basic livelihood due to factors such as drought conditions. China, India, Madagascar, Peru and Senegal provide examples of such migration. It has also been estimated by FAO that, for every full-time fisher in the smallscale subsector, additional employment for between one and three persons is generated in the fisheries sector. Since the small-scale subsector also targets fish for the international market, it contributes to foreign exchange earnings. The contribution of small-scale fisheries to foreign exchange revenue in many developing countries is significantly much higher than the contribution of small-scale farmers or peasants in agriculture. According to FAO, net foreign exchange earnings from fishery products are more important in these countries than the net earnings from exports of coffee, tea, rice and rubber combined (FAO, 2000b). Though commodity export prices for cocoa, rubber, palm oil, coffee and tea have been considerably depressed since the 1990s, prices for fish exports have remained advantageous. In several African, Caribbean and Pacific (ACP) countries, for example, fisheries exports, especially from the small-scale subsector, are now the major export earner, ahead of other exports, as in Senegal and Mauritania (SESRTCIC, 1998). Fisheries products are one of the few areas where ACP countries have seen their share of world trade increase. Between 1976
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and 1986, ACP fish exports to the EU rose from ECU 36 million to ECU 309 million, while, by 1996, the value of ACP fish exports exceeded ECU 946 million. In the 4 years from 1992 to 1996, the ACP share of total EU fish imports rose from 16.4 to 22.5%. This contrasts with general ACP trade performance, which saw the ACP share of imports into the EU decline from 6.7 to 3.4% in 1994 (CFFA, 1999). Despite being among the top ten fishproducing countries in the world, the per caput shares of marine fish production per fisher in China, Indonesia and India are quite low, at 1.7, 1 and 0.5 t, respectively (1998 figures), because of their large fisher populations. The difference is very striking when these developing countries are compared with the Nordic countries. In 1998, for example, Iceland had an annual per caput marine production per fisher of 334 t, Denmark 325 t and Norway 125 t.
Artisanal and Small-scale Fisheries in Developing Countries Artisanal and small-scale fisheries are accorded special recognition by the 1995 FAO Code of Conduct for Responsible Fisheries. Article 6.18 of the Code states: Recognizing the important contributions of artisanal and small-scale fisheries to employment, income and food security, States should appropriately protect the rights of fishers and fishworkers, particularly those engaged in subsistence, small-scale and artisanal fisheries, to a secure and just livelihood, as well as preferential access, where appropriate, to traditional fishing grounds and resources in the waters under their national jurisdiction.
What is an ‘artisanal’ or ‘small-scale’ or ‘traditional’ fishery? What exactly do we mean by terms such as ‘traditional’, ‘small-scale’ or ‘artisanal’ fisheries? These terms seem to have gained currency during the post-mechanization phase in many developing countries as a
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descriptive characteristic of those fisheries that were unmechanized. Traditional, smallscale or artisanal became the antonyms of ‘modern’, ‘large-scale’, ‘mechanized’ and ‘industrial’ fisheries. These terms had political significance in some contexts, where they became rallying points for fishers who opposed the introduction of destructive forms of bottom trawling, especially in Asia. However, the situation changed with the widespread adoption of motorization in small-scale fisheries worldwide. Traditional, artisanal or small-scale fisheries now include a range of fishing activities targeting anything from sedentary molluscs in littoral waters to highly migratory tuna stocks in distant waters. Thus, according to FAO (2001), half of the tuna production in the Indian Ocean originates from artisanal fisheries, meaning tuna that are caught by all gear other than purse-seines and industrial longlines. It includes subsistence fishers in the South Pacific as well as those fishing mainly for the export market, in Senegal and Chile. It ranges from resident women crab gleaners in the mangroves of northeastern Brazil, to Mexican longline fishers who go up to 200 nautical miles in their 7 m fibre-reinforced plastic (FRP) boats with 200 hp outboard motors in pursuit of shark, to the migrant longline fishers of Sri Lanka, who fish the farthest points of the Indian Ocean targeting tuna and shark resources. It may be an activity that is resident or migrant; occasional, seasonal, part-time or full-time. Traditional, artisanal or small-scale fisheries include, inter alia, rudimentary 3 m dugout canoes with a crew size of just one in Madagascar, as well as the 18 m pirogue of West Africa and the 16 m plywood or FRP boat of India that employ up to 40 crew members on board a single fishing trip, and the term extends further to shore-seines of Sri Lanka and India that would employ as many workers on shore to haul the net as a pirogue or a plywood boat would employ on board for purse-seine operations. Artisanal fishing thus includes highly individualized fishing operations such as cast nets and handlines; small-crew operations such as setting traps or pots in lagoons, estuaries or nearshore waters, diving for sedentary species in reefs
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and lagoons, operating a regime of gillnets and longlines; and the labour-intensive purse-seining and shore-based, beach-seining operations. The terms traditional, small-scale or artisanal could, however, have distinct connotations in different socio-economic contexts; for example, the definition of what constitutes traditional, artisanal or small-scale in an economic sense is fishing-operation-specific, although the definition of traditional fishing per se also has social overtones. Whereas in Madagascar the term artisanal refers to motorized fishing for the domestic as well as for the international market, the term traditional refers to unmotorized, kinship-based fishing for subsistence or for the local market, undertaken by fishers who respect local taboos and customs. In Fiji, the term artisanal is used to refer to fishing units harvesting for the domestic market; it is thus market-specific. In India, only the term traditional is legally recognized, but, unlike Madagascar, it denotes traditional fishing craft, i.e. a fishing craft of a type already in use before the arrival of mechanized fishing vessels. India also includes boat designs of foreign origin that were adopted during colonial times. The definition is thus craft-specific. In Indonesia and Malaysia, the term traditional is used, but, unlike India, the term is used in a gear-specific sense. All fishing units other than trawling are defined as traditional fishing units. In Peru, artisanal is the term in vogue, defined in tonnage-specific terms to indicate fishing vessels below 30 gross registered tonnage (GRT), although, according to Federacíon de Integracíon y Unificacíon de Pescadores Artesanales del Peru (FIUPAP), the organization of the artisanal fishers of Peru, about 85% of fishing vessels in Peru are below 10 GRT. In Chile, the term artisanal is used to indicate vessels below 50 GRT and less than 15 m in length, and an artisanal fisher in Chile will also hail from a particular caleta, work on a particular type of boat or in a particular sector (line fisherman, shellfish diver, seaweed harvester, etc.). Indeed, back in 1995–1996 in Chile, there was debate as to whether trawling could be considered as an artisanal gear in the hake fishery. In France,
the term used is artisanal, but the definition is length-specific. All vessels up to 25 m in length are categorized as artisanal units (Le Sann, 1999). The equivalent term in Canada is inshore fisheries, which refers to fishing vessels that are less than 20 m in length. A major distinction between the North and South is that, irrespective of the size of the unit, trawling operations, in general, are not considered small scale or artisanal in the South. There is thus no elegant definition. The problem of defining traditional, artisanal and small-scale categories has been compounded of late because of new technical changes, i.e. motorization of hitherto unpowered vessels, the use of powered gear-hauling devices, ice boxes, synthetic webbing for fishing gear, and the adoption of modern miniaturized electronic aids for navigation and fish detection. Artisanal and small-scale fisheries, in general, refer to the locus of smallest viable fishing units in a country or a province, with compatible fishing gear operations. Depending on the context, the definition could be based on:
• •
•
Whether or not the fishers are recognized as originating from a fishing caste, community or tribe. Whether or not the fishers are known to operate a specific regime of fishing craft or gear, or combination, and if they are at the bottom end of the hierarchy in a particular fishery in a country or province. Whether or not the fishers were fishing traditionally, but not necessarily confined to nearshore or inshore waters.
The definition of what constitutes traditional, artisanal or small-scale could be any one or a combination of these characteristics.
New technical changes in artisanal and small-scale fisheries The act of embracing motorization and the use of new technical accessories in the smallscale fishing sector in developing countries seems to have had a dual origin. On the one hand, it appears to be a reaction to unregulated mechanized fishing operations, such as
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in several Asian and African countries where bottom trawling and purse-seining intercepted and removed fish stocks before they could migrate into the estuaries or inshore waters where traditional fishing vessels, dependent on passive fishing techniques, used to fish. The small-scale subsector was thus forced to adopt fishing methods that would help them to fish in competition with the large-scale subsector. On the other hand, it also seems to be a response to burgeoning marketing opportunities as a result of growing demand for fish, coupled with easy availability of ice and credit. The latter phenomenon – response – in particular has contributed to a tremendous expansion of fishing capacity in small-scale fisheries. Motorization particularly suited those artisanal fishers who wanted to migrate to distant fishing grounds and those fisheries that were heavily dependent on labour power for the propulsion of larger pelagic fishing units. The technical flexibility offered by outboard motors and fish storage facilities was a significant factor in influencing artisanal fishers to motorize their fishing craft and exploit new fishing grounds. The fish merchants were enthusiastic to extend credit facilities to artisanal fishers and encouraged them to modernize their fishing operations. Fishermen’s cooperatives have also contributed to this process by extending credit and marketing facilities to their members in several developing countries. The small-scale subsector appears to have changed in several developing countries, away from one that was protected by legislation, most often through the enactment of an exclusive fishing zone, toward a situation where it is promoted by national and provincial governments at the expense of the large-scale industrial subsector. There are several examples of changing emphasis at government level in some countries, partly due to the growing realization that the smallscale sector makes better economic and social sense than the large-scale, industrial subsector. This perception recognizes the failure of an earlier model, which strongly emphasized investing in large-scale industrial fisheries (Tvedten and Hersoug, 1992; K. Dahou and M. Deme 2001, unpublished data).
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Such changes in official perception can be witnessed in important fishing nations, such as Senegal, India, the Philippines and even China, as well as among multilateral and bilateral aid agencies (e.g. World Bank et al., 1992).
Impact of new technical changes on small-scale fisheries During the pre-motorization phase under quasi-open access conditions, the fishing pressure that could be exerted by the small-scale fishing fleet was most often limited by ‘inefficient’ manual or wind-powered means of propulsion. These propulsion techniques also limited the size of the vessel and gear, and contributed to avoiding overfishing pressures. There was a division of labour practised by small-scale fishers, which often in effect acted like a limited-access regime. A beach-seine operator, for example, would not operate another gear type. Hook-and-line fishers would do only that type of fishing. Gillnet fishers often were categorized according to the particular species they specialized in fishing. From an exclusive dependence on manual or wind power, almost all these gear groups have now moved into dependence on sophisticated fishing technologies in many countries. From a technical point of view, outboard motors used by small-scale fishers are far more sophisticated than the diesel inboards currently used in most developing countries. Small-scale fishers have expanded the range of their fishing operations in several countries, to deeper as well as to distant waters, and their fisheries continue to be multi-species in nature. In some countries, such as Senegal, they have emerged as the most significant fishing power (Gaspart and Platteau, 2000). The technical developments, however, do not seem to have led to labour displacement as yet; on the contrary, they seem to have led to more fishing days and greater employment opportunities (Overa, 1998), possibly because the most labour-intensive (those requiring more labour per unit of output than other factors of production) of the range
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of craft–gear combinations are being used increasingly across seasons. The small-scale subsector thus seems to have maintained its labour-intensive character, with significant implications for poverty alleviation in rural areas. The sharing system prevailing on board small-scale fishing vessels in some instances, such as in Kerala, India, does not seem to indicate any major shift away from labour in favour of capital. Thus, from the employment and income points of view, the impact of technical changes in the small-scale subsector, at least in the short run, seems to have been positive in some labour-surplus fishing economies. Studies in Ghana, however, show that after the introduction of motorization in canoe purse-seine fishing, the owners of canoes and their families benefit more from new allocation regimes than do workers and their families. This is because the share of catches accruing to capital, and hence to owners, has increased since the introduction of motorization. The shares are in physical quantities of fish; therefore, the larger the share for owners, the greater the amount of fish available for their wives, who are fish processors or traders. This contrasts with the much lower volumes of fish for the wives of workers (Overa, 1998). There is growing inequality between those fishing for the domestic market and those fishing for the export market, with implications for gender relations in artisanal fishing communities. With growing dependence on the export market and export agents, there are negative impacts on access of women to artisanal fish production, for example in Senegal. Fish traditionally sold through fishers’ wives are now sold directly to export agents, often men from outside the fishing community (ICSF, 1997). The implications of such growing social inequalities have yet to be analysed in any systematic fashion. An expansion in the scope of small-scale fishing activities in the relatively limited time span of 10–20 years in developing countries has not been without its negative consequences. With the technical capacity to go after the quarry, as opposed to waiting for it in the nearshore fishing ground, the gear base of the most economically active small-scale
fishers, especially of those in the forefront of technical change, has been losing its diversity, becoming narrower and more standardized. The artisanal fisheries, as a result, have become far more differentiated; they now include both powered and unpowered vessels, and both active and passive gear groups. There has been a tremendous expansion of fishing capacity and increasing fishing pressure in the artisanal sector in many developing countries, especially in West Africa, South and Southeast Asia. In the Senegalese artisanal fisheries, for example, there has been a 42% increase in the number of pirogues between 1994 and 1997 (Gaspart and Platteau, 2000). In the traditional fisheries of Kerala, India, the number of plywood boats has increased by 300%, from less than 2000 in 1991 to close to 6000 in 1998, and all these craft are motorized fishing vessels using outboard motors (SIFFS, 1992, 1999). The rapid expansion of artisanal fishing capacity has begun to exert overfishing pressure on coastal fisheries resources, especially in Asia and Africa, which, until the beginning of the motorization phase, were caused mainly by the unregulated operations of non-selective, large-scale, industrial fishing units. As Gaspart and Platteau (2000) have pointed out in the case of Senegal, They [the fishermen] do not seriously consider the possibility of their being partly responsible for overfishing and, therefore, the idea that they could combat environmental degradation by restricting their own fishing effort seems alien to most of them. Revealingly, there is a clear tendency among Senegalese fishermen to externalize the problem by blaming industrial fishing vessels for the destruction of fish resources.
This is now true of small-scale fisheries in several other developing countries as well. It may not, however, be easy for the small-scale subsector to acknowledge biological and economic overfishing problems that they themselves are contributing to, unless largescale, non-selective industrial fisheries are regulated effectively. Although the size of the gear still remains small in comparison with gears used in large-scale fishing, those used in the
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small-scale subsector nevertheless are growing in size. These are also becoming less and less selective. In the Statement of the Workshop on Problems and Prospects for Developing Artisanal Fish Trade in West Africa, Dakar, Senegal, 30 May to 1 June 2001, organized by ICSF, Collectif National des Pêcheurs Artisanaux du Senegal (CNPS) and Centre de Recherches pour le Developpement des Technologies Intermediaries de Pêche (CREDETIP), and supported by the FAO-DFID Sustainable Fisheries Livelihood Programme (SFLP), participants from 13 West African countries, including small-scale fishworker representatives, were critical of the destructive impact on inshore fisheries resources of monofilament nets and ring seines in the small-scale subsector (ICSF, 2001a). There are also increasing conflicts within the small-scale subsector amongst different gear groups as a result of increased mobility of fishing vessels, capacity expansion and overfishing pressure. With motorization, the division of labour also seems to have broken down by making it easier for unskilled people to migrate into fishing activities. Built-in conditions of limited access regimes have broken down under pressures of motorization.
Small-scale Fishing Industry Perspective on an Ecosystem-based Approach to Fisheries Management In spite of the problems of biological and economic overfishing posed by small-scale fisheries in some contexts, an ecosystembased approach to fisheries management could help valorize small-scale compared with non-selective large-scale fishing, and could also help to bring about the required changes that might minimize threats to its existence. What could be a small-scale fishing industry perspective on an ecosystem-based approach to fisheries management? Since the fishing grounds of the smallscale subsector often are the richest and the most diverse, and since the impacts of pollution and destructive fishing practices
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are experienced most immediately by this subsector, an ecosystem-based approach to fisheries management could be of immense interest. Also, with greater mobility of fishers across borders following migratory fish stocks, for example in West African countries, an ecosystem-based approach to fisheries management might help to design coherent regional fisheries management strategies that take into account all factors that influence fish stocks in their entire range of movement, and all fisheries that interact with such stocks. An ecosystem approach, through better recognition of the relationship between the fishery, fishing resources and fish habitat, could help develop better regulatory measures to manage large-scale fishing operations that employ non-selective fishing gear and methods, and thus help minimize the cascade effect on fish stocks and on the livelihood of small-scale fishing communities. It could also help better regulate destructive fishing operations such as dynamiting and cyanide fishing, and help to regulate the use of fine-meshed nets by small-scale fishers themselves. An ecosystem-based approach can be applied to understand, and to prevent, land-based sources of pollution that have an adverse impact on plankton, which are food for smaller pelagic fish, the mainstay of small-scale fisheries in many Asian and African countries. Thus, concurrent with proposing and implementing measures that basically address the impact of fishing on fish stocks and the marine habitat, there is need to take steps to minimize the effect of pollutionrelated habitat degradation on fish stocks arising from coastal settlements, industries and agriculture, and oil spillage from the production and transport of crude oil. In addition, an ecosystem-based approach could be used when addressing reduction of nursery grounds as a result of destructive activities such as construction and reclamation in coastal areas, mangrove deforestation and brackish-water aquaculture, as well as the loss of marine biological diversity as a result of destruction of coral reefs due to global warming, dynamiting or cyanide fishing. Land-based sources of pollution could also have a direct negative impact on mollusc
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beds, the mainstay of artisanal fisheries in several Latin American countries. CONAPACH and FIUPAP – the artisanal fishers’ organizations of Peru and Chile, respectively – often have been complaining of such impacts arising from mine-tailings and fishmeal effluent. Agriculture and water-use regimes on land could also have an indirect impact on coastal fishing grounds through soil erosion and decreased fresh water discharges into the sea. Given that in modern times the fisheries sector, especially the small-scale subsector, is at the receiving end of land-based and seabased sources of habitat degradation (ICSF, 1994, 1996), an ecosystem-based approach can broaden the scope of fisheries management, especially in many developing countries, to address these forms of degradation effectively. An ecosystem-based approach could contribute to understand better the intricacies of natural factors and their impact on fish stocks. This is significant because many pelagic stocks, which are important for the small-scale subsector, are influenced by changing weather and climate conditions, and are highly vulnerable to oceanographic factors. Such understandings need to be articulated to fishers to enhance their understanding of the ‘prey in context’, mainly to draw the distinction between the impact of natural factors and of fishery-dependent factors on relative abundance or scarcity of fisheries resources. In Pulicat Lake, India, for example, the artisanal fishers argue that the mullet resources of the lagoon will simply perish if the salinity level exceeds that of the sea due to evaporation, zero exchange of water (as a result of spit formation at the mouth of the lagoon) and zero discharge into the lagoon from rivers (due to upstream dams). Fishers do not, therefore, believe that conservation of mullets under such conditions is possible just by refraining from fishing (Mathew, 1991). An ecosystem-based approach can facilitate a better understanding of prey–predator relationships at sea, and also the impact of fishing gear selectivity on marine living resources. Programmes designed to conserve charismatic species such as sea lions, dolphins and sea turtles sometimes become counterproductive when these resources multiply in
large number and compete with fishers for the quarry, in the process conflicting with the interests of small-scale fishers, and often adversely affecting their livelihood without significantly contributing to the health of the marine ecosystem. In Talara, northern Peru, for example, squid jiggers in the artisanal fisheries complain about predation of squid resources by sea lions and dolphins. FIUPAP has estimated that the annual damage caused by the southern sea lion to their fisheries is about US$64 million (Manuel, 1997), and similar complaints are also reported from the Maritimes, Canada. In spite of using selective fishing methods, the small-scale fishers of Orissa – the poorest province of India – are prohibited from fishing in their traditional grounds because of arbitrary declaration of sea turtle conservation zones for the protection of olive ridleys. Most importantly, an ecosystem approach can valorize and build upon the ecosystem principles inherent in traditional knowledge systems of artisanal and smallscale fishing communities around the world. Kurien, based on an analysis of Asian coastal proverbs, provides an insight into various facets of traditional knowledge about the complex ecological systems with which they interact (Kurien, 1998). Traditional ecological knowledge, based on locale-specific understanding of the components of the ecosystem, however, is often confined to an understanding of a limited number of environmental and oceanographic parameters, and restricted to the biology and behaviour of the target species and species that immediately predate on, or are prey for, these target stocks. Also, as J. Cordell (2000, unpublished data) tersely pointed out, customary marine tenure is designed to preserve ‘the social order, not the balance of nature’.
Ecosystem-based Approach to Managing Small-scale Fisheries in Developing Countries An ecosystem-based approach provides a framework to fisheries conservation, management and development and makes it
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possible to look at aspects of fisheries, including land- and sea-based parameters, as well as known and unknown factors. It can enable the subsector to address issues of both immediate and long-term concern, especially to prevent the impact of land-based sources of pollution and coastal degradation, to rebuild depleted fish stocks and to restore marine habitats. It can facilitate the building up and strengthening of traditional knowledge systems in artisanal and small-scale fishing communities.
Problems in managing small-scale fisheries The main challenge in applying an ecosystem approach to small-scale fisheries management is in negotiating the adverse impacts on the ecosystem arising from factors outside the control of the small-scale subsector. So, if we are talking about applying such an approach to small-scale fisheries, then we are confined to discussing input and output control measures, and institutional arrangements to regulate access to fishing grounds, especially when they are overcrowded or are in a state of ecological stress. In multi-species, multi-gear and multi-cultural fisheries, especially in the small-scale subsector, what indeed would be the best locus of measures to manage fisheries would be a moot point. Quota management regimes are ruled out because, by using such measures, it would be impossible to manage with any reasonable degree of success the ‘mosquito’ fleet operating from a multitude of landing centres in many developing countries. Moreover, the associated problems of such regimes, particularly high grading and concentration of ownership in the hands of a few, would only exacerbate social problems in labour-surplus, small-scale fisheries. While discussing the need for fisheries management in small-scale fisheries, especially effort control and limited-entry measures, the role of conventional management measures is limited by poor institutional arrangements. The problem is complicated further by numerous landing centres, too
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many fishing vessels as well as people in the fisheries. It would, therefore, be difficult, if not impossible, for governments to regulate marine fishing activities successfully, especially to introduce limited-entry regimes in small-scale fisheries, without the active participation of fishing gear groups or fishworker organizations. There is, however, a lack of such organizations in many developing countries. Although fisheries management programmes in industrialized countries can focus directly on fishing capacity, fisheries resources and fish habitat-related issues, such an approach may be difficult in developing countries, where the State, as a priority, may have to focus on the human dimension in the fisheries sector, especially the need for poverty alleviation and food security in coastal areas, before addressing fisheries management issues per se. The short-term goals of small-scale fisheries management under the aegis of the State cannot be exclusionary in nature, given the widespread poverty and unemployment in rural societies in many developing countries. A State that cannot provide an alternative employment to fishers may also not find it easy to ask people to leave the fishery to alleviate overcrowding in fishing grounds. However, such exclusionary regimes can be designed and implemented by the small-scale fishing industry itself and legitimized by the State machinery. We have yet to see effective fisheries management programmes in any labour-surplus, small-scale fisheries in developing countries that have been implemented successfully by the State. Even in large-scale fisheries, for that matter, there is hardly any success story of fisheries management, especially from developing countries. The large-scale fisheries of important fish-producing countries such as China, Thailand, India and Indonesia still do not figure as countries with effective management programmes. Given the collapse of fisheries even in countries such as Canada – which was believed to have an effective fisheries management system until the collapse of the Canadian Atlantic cod fisheries in the 1990s – the lack of political will, or confidence in the feasibility of fisheries management
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programmes, is understandable in many developing countries. There is no straightforward, universal solution to many of the vexing problems of overfishing and overcapacity in small-scale fisheries, however, and this calls for a better understanding of the structure of fisheries, the motives of and constraints on fishers, and the interaction between various components of fisheries, especially between the large- and small-scale, and between different gear groups within the small-scale subsector. Given all the failures – and indifference – of the past, new fisheries management initiatives should be based on a process of dialogue with the small-scale fishing industry, to arrive at long- and short-term goals for management, taking into account social, economic, ecological and other relevant aspects of labour-surplus fisheries in developing countries. Such initiatives can be taken by the State. One way to create room for such a dialogue would be to redistribute fishing space progressively to the small-scale fisheries subsector by phasing out large-scale, non-selective fishing units. Such a measure would also consolidate the recognition granted to small-scale fisheries by several governments since the 1990s and by the 1995 FAO Code of Conduct for Responsible Fisheries. Simultaneously, there should be a serious effort initiated by the State in the long term for greater institution building, such as building up fishworker organizations, that will help devolve principal fisheries management functions to the representative small-scale fishing industry organizations. As Jentoft and McCay (1995) point out, a devolutionary process should aim at delegating authority – not just decentralization – based on the subsidiarity principle, i.e. implementing management functions at the most effective level, starting from the bottom. In large countries such as China, Brazil, India and Indonesia, where it is almost impossible to have a centralized or even provincial-level effective fisheries management progamme, such an approach seems better sense. These institutions, however, should be designed in such a manner that they become true representative bodies, that they do not become hegemonic or
inequitable, or end up just as mere conduits for State patronage.
Building upon community-based fisheries management In developing countries, institutional structures that are ideal for undertaking effective fisheries management functions should be identified. Some lessons may be drawn from traditional community-based fisheries management initiatives involving fishing communities, especially to regulate access to fisheries and to limit fishing capacity. These tend to be more localized initiatives among homogenous gear groups, and often have a conflicting relationship with other gear types. They are forms of rights-based fisheries, often based on rotational access to fisheries resources, but their effectiveness is more confined to stationary or beach-based gear or to sedentary species, than to mobile gear or species. There are already several examples of such traditional arrangements in developing countries. The most salient aspect of these arrangements is that they have clearly defined rules of exclusion based on allegiance to a caste, community or a group. These arrangements, however, most often emphasize aspects of allocation, and are designed mainly to mitigate conflicts within their membership over access to marine fishing space. The fishing capacity of the members, however, could exceed the regenerative capacity of the resource and thus contribute to overfishing pressures, especially in the context of new technical changes in fisheries. In Pulicat Lake, India, for example, there is the padu system, a system of rotational access to shrimp fishing grounds, but it does not mitigate pressure on shrimp resources because different groups of members, in a rotational fashion, are harvesting the resources incessantly. Similarly, in several estuarine fisheries in Asia, although several stake-net groups practice rotational access, the mesh size is below the legal limit and it often contributes to overfishing of juveniles of diadromous species. It is also noticeable that
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traditional arrangements to regulate access are challenged under conditions of greater market demand, when non-member gear groups in coastal fishing villages refuse to recognize the legitimacy of these arrangements, and often do so with the support of the government (Mathew, 1991). As J. Cordell (2000 unpublished data) points out (emphasis added): Whereas available information may be sufficient to document general features of tenure practices, it is usually not sufficient to generate specific recommendations concerning how local tenure could be integrated with contemporary systems of marine resource use. Communities today must deal with environmental issues on a scale the ancestors were never confronted with.
The issue of legitimacy is exacerbated further by the conflicts between exclusionary traditional arrangements and the nonexclusionary formal arrangements under the auspices of the State. This can be tackled effectively if the governments throw their weight behind traditional systems. In exchange for lending formal recognition, the governments can insist that these arrangements should adopt and implement effective conservation measures.
Adopting a ‘crossword’ approach to small-scale fisheries management Conservation of fisheries resources, protection of fish habitats and allocation of fisheries resources to fishers are the three most important considerations in fisheries management. The vantage point to start from is the gear group or a bundle of compatible gear groups, because without their cooperation it would be impossible to adopt effective conservation measures and to protect fish habitats from fishery-related stress. It is thus the principal link in fisheries management, especially in small-scale fisheries in developing countries. Initiating fisheries management measures in small-scale fisheries in developing countries could be through a ‘crossword’ approach, i.e. filling up management niches
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that are relatively easy at first, and then moving on to more difficult ones with the aid of early breakthroughs or solutions. Stationary and beach-based gear groups, gear groups fishing around artificial reefs, and gear groups targeting sedentary stocks arguably are better candidates to collaborate in a fisheries management programme. The most difficult ones could be the migrant gear groups, who may have a vested interest in maintaining an open-access regime, like the longline fishers of Senegal (Gaspart and Platteau, 2000). Formal and traditional fisheries arrangements need to combine, to generate effective fisheries management policies and programmes. Simultaneously, measures should be drawn up to regulate large-scale fishing operations, including a proscription of fishing gear and fishing operations that are destructive or socially inappropriate. There should also be programmes to build up user participation in fisheries management, especially in small-scale fisheries, which are highly scattered and difficult, if not impossible, to ‘manage’ in a centralized fashion. In this context, devolutionary mechanisms are vital, and strong fishworker unions that can undertake fisheries management programmes successfully are required to be built with full legitimacy (Jentoft and McCay, 1995).
International cooperation to manage small-scale fisheries As a global solution to the national, provincial or local problems, an important responsibility of industrialized countries is to not sell their excess fishing capacity to developing countries at low prices, nor to send it as an article of aid (although it might appear to be the easiest solution to their problems of overcapacity), nor to transfer the excess capacity through joint ventures. What is in fact required is weeding out of the excess capacity problem, and northern countries should not, in the first place, be building up excess capacity. Subsidies are still extended for fleet expansion, for example, in several EC
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countries, and this practice should be strongly discouraged. To ensure the cooperation of the most significant gear groups in the small-scale subsector, incentive schemes might be considered. In this context, we have to look for global solutions to local and national problems. There are at least three possibilities that could be considered. First, for the management of overexploited fisheries in developing countries, there is need to set up a well-designed, time-bound, international fisheries management assistance fund in exchange for a commitment to manage fisheries in a consultative and transparent manner, within the framework of an ecosystem approach. However, governments in developing countries should also consider investing in fisheries management from existing revenue resources. Although net earnings from fisheries exports for many developing countries are quite high, few significant investments are made in conservation and management by most developing countries. For example, India, with a gross value of fisheries output of US$5000 million in 1997–1998 (at ex-vessel prices) and export earnings of over US$1000 million, spends insignificant amounts on activities that can be treated as fisheries management. As Willmann et al. (2000) point out, in 1999, when Norway spent about 8% of total gross revenue from marine fish landing on fisheries management, Iceland 3% and Newfoundland 20%, Thailand spent only 1.64%, although its fisheries have been beset with overcapacity and overfishing problems for some time. The mindset has yet to change from considering fisheries as an extractive industry, to seeing it as an industry based on renewable natural resources that have to be stewarded. There are at least two ways, for example, to set up such an international fisheries management assistance fund: through a Tobin tax (a tax on currency trades across borders, as proposed by the USA economist James Tobin in the 1970s), and Belgium has put such a tax on foreign exchange transactions on the discussion agenda for its current 6-month EU Presidency; and through a consumption tax on fish and fish products in industrialized countries.
The financial assistance for cash-starved developing countries with a clear political will to manage their fisheries could include assistance to:
• • • • • •
Bring about better control over the input of fishing effort and the output of fish. Introduce participatory and devolutionary management regimes and equitable property rights. Set up effective monitoring, control and surveillance (MCS) systems. Protect fish habitats (ICSF, 2001b). Conduct research on the status of fish stocks. Build up fishworker organizations at the local, provincial and national levels.
Secondly, for fisheries that are well managed, eco-labelling might provide an incentive to fishers, but it might be relevant only to exportable species in developing countries, principally for the USA and EU markets, and produced mainly by homogenous gear groups. Also, even if a particular small-scale fishing fleet is using a selective fishing gear and has acceptable levels of by-catch or discard, it may not qualify for an eco-label if it is targeting a stock that is subject to overfishing pressure from other fishing activities. Thirdly, for small-scale fisheries that are overcrowded as a result of demographic pressure in developing countries, industrialized nations may contribute to alleviating such pressure by facilitating temporary migration of surplus labour into their domestic or distant-water fisheries, particularly into fisheries that are characterized by labour shortage. The substitution of labour with capital in many developed country fisheries, inter alia, is believed to be a function of growing labour shortage. The average age of a Japanese and Korean fisherman, for example, is over 60 (OECD, 2000), and that of a Canadian fisherman in the Maritimes is around 47. Instead of substituting labour with capital, fisheries at low levels of technical intensity can be maintained, even in the event of chronic labour shortage in the North, if well-trained migrant workers from developing countries are recruited. Threats to immigration can be addressed by carefully designing time slots for transient accommodation of
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labour. Already, several OECD countries are employing migrant fishworkers from developing countries in their fisheries because of labour shortage. This is especially noticeable in Spain, France and Italy. It is just a matter of legalizing such arrangements. There are several examples of employment arrangements between the North and the South, especially in relation to the employment of computer and medical professionals from countries such as India in the USA and Europe. Needless to say, this will not be a solution to the problems arising from demographic pressure, but it would certainly be seen as a positive gesture from the North to the South.
Conclusion Unlike the single-species model in fisheries management, which is by far the most prominent model in most parts of the world, an ecosystem-based approach to fishery management could be an effective tool for small-scale fisheries in developing countries since it may take into account the complexity of the marine and coastal ecosystems, an attribute already factored in a limited way into the decision-making processes of several traditional, small-scale fishing communities. A multitude of species further exacerbates the problem in countries in the tropical belt. According to the FAO FishBase, in India, for example, about 263 out of the 1000 marine and brackish-water fishes identified so far are commercially significant, as against just 25 out of 250 in Norway, and 21 out of 300 in Iceland. In Indonesia and the Philippines, countries with the greatest marine biodiversity in the world, the figures are 681 out of 2511 and 616 out of 2255, respectively. Each of these fish will have several stocks, and the total number of stocks could run into thousands. Very little is known about the impact of fishing on these stocks. It is therefore quite doubtful whether an ecosystems approach can be considered, in its entire range, in tropical multi-species fisheries, for example. An attempt to deal with the complexity of the ecosystem should take into account
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the scale of fishing operations as well as the heterogeneity of fishers. There are small- and large-scale, active and passive, and there are responsible and destructive fishers. An ecosystem approach has therefore to be used in a dialectical sense: it should, on the one hand, take into account the effects of fishing on fish stocks, especially the unequal impact of small-scale and large-scale fishing on target fish stocks and the marine and coastal ecosystems, undertaken under different economic, social and political contexts. On the other hand, it should also take into account the effects of marine ecosystems on fishworkers. However, adopting an ecosystem approach is easier said than done. Developing the building blocks of an ecosystem-based approach with social sensitivity, and documenting the impact of fishing on targeted stocks and their habitats, as well as on other species in the ecosystem, are complex, difficult and expensive tasks, and require a ‘global partnership for sustainable development’, as quoted in the epigraph of this chapter. It should be based on a ‘crossword’ approach, which implies a realistic time frame to implement various components in a sequential manner. To persuade the small-scale subsector to adopt an ecosystem approach, governments should phase out all destructive forms of large-scale fishing, such as bottom trawling, as an incentive to the small-scale subsector, but subject to the subsector agreeing to improve its own fishing practices. There is a need to broaden the artisanal and small-scale knowledge base to encompass ecological parameters hitherto ignored or not understood sufficiently, e.g. the greater impact of natural factors, the broader picture of prey–predator relationship, the larger role of fish habitats, and factors that contribute to unprecedented habitat degradation, such as pollution. There should, however, be a sense of ‘historical continuity’ (Kurien, 1998), in an ecosystem-based approach, an attempt to build up on what already exists, especially to transmute the past traditions with new scientific insights to address the needs of the present meaningfully, or ‘the contemporary systems of marine resource use’, as Cordell puts it (Cordell, 2000). Of course, at the level of
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practice, it is indeed a challenge, with huge financial implications, which few developing countries can afford without international assistance.
Acknowledgements The author would like to acknowledge the constructive comments and suggestions on an earlier draft of this paper from Aliou Sall, Antonio Carlos Diegues, Brian O’Riordan, Chandrika Sharma, John Kurien, K.G. Kumar, Michael Belliveau, Nalini Nayak, C. Rammanohar Reddy, Rolf Willmann, E. Vivekanandan, V. Vivekanandan and the Scientific Committee of the Reykjavik Conference on Responsible Fisheries in the Marine Ecosystem. I am also grateful to R. Ramya of the ICSF Documentation Centre for invaluable assistance in the preparation of this chapter.
References CFFA (Coalition for Fair Fisheries Arrangements) (1999) Fishy Business, ACP-EU Fisheries Relations: Who Benefits at What Cost? CFFA, Brussels. DAGA (1984) International Conference of Fishworkers and their Supporters. Report. DAGA, Hong Kong. EC (European Commission) (2001) Green Paper: the Future of the Common Fisheries Policy, Vol. I. European Commission, Brussels. FAO (1995) Code of Conduct for Responsible Fisheries. FAO, Rome. FAO (1998) Integrated Coastal Area Management and Agriculture, Forestry and Fisheries: FAO Guidelines. FAO, Rome. FAO (1999) Number of fishers doubled since 1970. http://www.fao.org/fi/highligh/fisher/ c929.asp FAO (2000a) FishSTAT Plus. Universal software for fishery statistical time series, Version 2.3. FAO, Rome. FAO (2000b) Globalization and Implications for International Fish Trade and Food Security. Item 6 of the Provisional Agenda. Committee on Fisheries. Sub-Committee on Fish Trade. Seventh Session. FAO, Bremen.
FAO (2001) The State of World Fisheries and Aquaculture 2000. FAO, Rome. Gaspart, F. and Platteau, J.-P. (2000) Promotion of Coastal Fisheries Management: 1. Local-level Effort Regulation in Senegalese Fisheries. FAO Fisheries Circular No. 957/1. ICSF (International Collective in Support of Fishworkers). (1994) Proceedings of the Cebu Conference. ICSF, Madras. ICSF (1996) Proceedings of the South Asia Workshop and Symposium on Fisheries and Coastal Area Management. ICSF, Chennai. ICSF (1997) Globalization, Gender and Fisheries. Report of the Senegal Workshop on Gender Perspectives in Fisheries. ICSF, Chennai. ICSF (2001a) Report of the Workshop on Problems and Prospects for Developing Artisanal Fish Trade in West Africa. Dakar, Senegal, 30 May–1 June 2001. ICSF, Chennai. ICSF (2001b) Editorial. SAMUDRA Report, No. 28 (April). ICSF, Chennai. Jentoft, S. and McCay, B. (1995) User participation in fisheries management: lessons drawn from international experiences. Marine Policy 19(3), 227–246. Kurien, J. (1998) Traditional ecological knowledge and ecosystem sustainability: new meaning to Asian coastal proverbs. Ecological Applications 8, S2–S5. Le Sann, A. (1999) No more bounty. SAMUDRA Report, No. 24 (December). ICSF, Chennai. Manuel, M. (1997) The roar of the sea lion. SAMUDRA Report, No. 18 (July). ICSF, Chennai. Mathew, S. (1991) Study of territorial use rights in small-scale fisheries: traditional systems of fisheries management in Pulicat Lake, Tamil Nadu, India. FAO Fisheries Circular, No. 839. National Research Council (1999) Sustaining Marine Fisheries. National Academy Press, Washington DC. NMFS (USA National Marine Fisheries Service) (1999) Ecosystem-based fishery management: a report to the Congress by the Ecosystem Principles Advisory Panel. http:// www.nmfs.noaa.gov/sfa/reports.html OECD (Organisation for Economic Co-operation and Developement) (2000) Transition to Responsible Fisheries: Economic and Policy Implications. OECD, Paris. OECD (2001) Review of Fisheries in OECD Countries: Policies and Summary Statistics. OECD, Paris. Overa, R. (1998) Partners and competitors: gendered entrepreneurship in Ghanaian canoe
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fisheries. Dissertation. University of Bergen, Norway. SESRTCIC (1998) http://www.sesrtcic.org/ defaulteng.html Senate Committee on Fisheries (1993) http://www.parl.gc.ca./36/1/parlbus/com mbus/senate/com-e/fish-e/past_rep-e/ 93repen1.htm SIFFS (South Indian Federation of Fisherman Societies) (1992) A Census of the Artisanal Marine Fishing Fleet of Kerala 1991. SIFFS, Trivandrum. SIFFS (1999) A Census of the Artisanal Marine Fishing Fleet of Kerala 1998. SIFFS, Trivandrum. Tvedten, I. and Hersoug, B. (eds) (1992) Fishing for Development: Small-scale Fisheries in Africa. Scandinavian Institute of African Studies, Motala. UN (United Nations) (1955) Report of the International Technical Conference on the Conservation
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of the Living Resources of the Sea. UN, New York. UN (1983) The Law of the Sea: United Nations Convention on the Law of the Sea. UN, New York. UN (1992) Agenda 21: Programme of Action for Sustainable Development. UN, New York. UNDP (1999) Human Development Report. UNDP, New York. Van Dyne, G.M. (ed.) (1969) The Ecosystem Concept in Natural Resources Management. Academic Press, New York. Willmann, R., Boonchuwong, P. and Piumsombun, S. (2000) Fisheries management costs in Thai marine fisheries. In: Proceedings of the IIFET 2000 Conference. International Institute of Fisheries Economics and Trade, Oregon. World Bank et al. (1992) A Study of International Fisheries Research. World Bank, Washington, DC.
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An Environmentalist’s Perspective on Responsible Fisheries: the Need for Holistic Approaches Tundi Agardy Sound Seas, Bethesda, Maryland, USA
Abstract The new millennium marks a time when scientific opinion and environmentalist sentiment are at last converging on the perception that the world’s natural marine heritage is facing grave threats. What environmentalists have come to call the ‘marine biodiversity crisis’ is a pervasive and by now well-documented phenomenon, until recently occurring largely unnoticed beneath the deceptively unchanging blanket of the ocean’s surface. The fact that this problem is essentially an invisible one makes it all the more insidious, and our terrestrial bias makes combating the problem a huge and difficult task. Human impacts on our seas take many forms and result not only from activities that affect species directly – such as overfishing, in-filling of wetlands and coastal deforestation – but also from activities that affect oceans indirectly, such as through land-based sources of pollution, freshwater diversion from estuaries, invasive species and climate change. Due to the expanding scope of both global coastal degradation and fisheries conflicts, environmental groups recently have become more and more involved in fisheries management and conflict resolution. In tackling fisheries issues, most organizations attempt to base their projects and advocacy on the best available scientific information. These groups sometimes undertake in-house scientific research, predictive modelling and meta-analysis. However, in most cases, the non-governmental organizations (NGOs) are recipients of scientific information and liaise between the scientific community, decision makers and the public. The key scientific information underpinning campaigns and field projects addresses three facets of sustainability: (i) the levels of resource removal that can be realized without adverse impact on the ecosystem, given the particular environmental condition of the ecosystem at time of harvest; (ii) the least invasive means by which that harvest can be undertaken at desired levels of harvest, such that habitat impacts and by-catch are minimized; and (iii) the most appropriate stocks for large-scale harvest, namely protecting stocks that are sole representatives of genetically unique organisms and stocks whose ecological role is critically important and so not redundant. Environmental groups, however, are as diverse in their character, approach and constituencies as the environmental problems they address. They function variously as purveyors of information, as translators of scientific and management language to the vernacular, as honest brokers (although their own value systems cause some to question their honesty), as advocates and lobbyists for certain types of reform or regulatory measures, and as adversaries to management agencies and industry when invoking environmental litigation. In many of these roles, environmental groups have been seen as the antithesis to development, to business interests, and to the needs of many user groups. Yet, today, environmental groups play an increasingly important non-adversarial role in demonstrating how conservation and sustainable use can be accomplished, through practical, applied conservation projects that benefit users, community groups, business and national interests. If a common environmentalist response to fisheries-induced loss of marine © 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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biodiversity can be said to exist (and this is a dangerous assumption, given the diversity of groups and their approaches), it is to synthesize existing information, communicate it, and advocate change in policy and regulations where felt necessary. In addition, some groups go beyond fisheries-by-fisheries management reform to advocate: (i) shifting the burden of proof when evaluating fishing impacts on ecosystems; and (ii) establishing strictly protected marine reserves to further our understanding of and protect species, habitats and ecological processes. Such reserves are implemented in a variety of fashions: as components within larger, multiple-use protected areas that seek to accommodate a wide array of users; as single elements in scientifically designed reserve networks; and as one tool of many used in corridor approaches, coastal management and regional planning. From the environmentalist’s or conservationist’s perspective, solutions lie not in shutting down fisheries but rather in modifying the way we undertake management, and in using public awareness to help raise political will to conserve marine systems. Coupling current consumer awareness and purchasing power with strong and effective management could indeed alleviate pressure on many marine species and allow their subsequent recovery. Additionally, environmental groups will need to recognize and support real willingness among governmental agencies and decision makers to protect areas needed for fish spawning, feeding and migration through marine reserves, and help such forward-thinking agencies to enter into enforceable international agreements to protect shared or commons resources. By highlighting such potential successes, and by working to demonstrate how success can be achieved, environmental groups can begin to shed their image of extremist adversaries, and help decision-making bodies implement effective and beneficial management regimes. A common thread is now emerging from analysis of cases where fisheries management and marine conservation has succeeded – and we can well learn from this common thread. The central element in these initiatives is a holistic approach – one that considers renewable living resources as part of a wider, interconnected ecosystem, one that evaluates all aspects of production or development and one that treats humans as bone fide elements of living systems. These integrated approaches take into account ecosystem interconnections and the true ecological costs of fisheries, the entire production chain and its environmental costs, and human interconnections, and thus the social costs (and benefits) of fisheries development. Holistic solutions are those that recognize these connections and try to minimize ecological, environmental and social costs, while maximizing the benefits (and benefit sharing) that can accrue from engagement in well-managed marine resource use. Given the magnitude and complexity of global fisheries issues, only such holistic prescriptions will make it possible for nations to achieve responsible fisheries in the future.
Introduction Environmental groups increasingly influence the direction of resource management and habitat protection in many parts of the world, addressing marine as well as the more traditional terrestrial issues of conservation. Commonly held roles for such groups include synthesizing understanding about marine issues, highlighting findings in a way that can be communicated to the general public, and advocating for policy reform and incentives that will change human behaviour to make resource use more sustainable. In addition, and perhaps most importantly, environmental groups play a crucial role in leading by example – demonstrating how environmentally sound and socially beneficial resource use can be achieved through field projects and interventions.
In general, the environmental community espouses a widely held view that management of ocean and coastal resources is poorly handled by most governments, leading time and time again to conflict. Decision makers fumble in the dark with classically ineffective fisheries management, create marine parks and other management areas with little or no control over activities within them, and are so overwhelmed by the complexity of river basin management that they are largely unable to deal with problems of land-based sources of pollution and degradation. Effective measures to address declines in ocean health and productivity remain few and far between, and are often too little, too late. The basis of this crisis may rest in large part on our inability to communicate what is happening and why (Agardy et al., 1999). This lack of effective communication about ocean issues
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can and should be addressed by the environmental community – both non-governmental and intergovernmental – but in this we, too, have fallen short of the mark, as much as the governments we criticize. As a result of all these factors, the world’s natural marine heritage is facing grave threats in many regions. What environmentalists have now come to call the marine biodiversity crisis is thus a pervasive phenomenon, occurring largely unnoticed beneath the deceptively unchanging blanket of the ocean’s surface. The fact that this problem is essentially an invisible one makes it all the more insidious, and our terrestrial bias makes combating the problem a huge and difficult task (Wilder et al., 1999). The global marine environment seems to be undergoing dramatic change at hitherto unprecedented rates, revealing as folly our previously held notion of vast and limitless oceans. Human impacts on our seas take many forms and result from activities that affect species directly, such as overfishing, in-filling of wetlands and coastal deforestation, to activities that affect oceans indirectly, such as through land-based sources of pollution, freshwater diversion from estuaries, and climate change. Environmental groups are as diverse in their character, approach and constituencies as the environmental problems they address. They function variously as purveyors of information, as translators of scientific and management language to the vernacular, as honest brokers (although their own value systems cause some to question their honesty), as advocates and lobbyists for certain types of reform or regulatory measures, and as adversaries to management agencies and industry through the use of environmental litigation. In many of these roles, environmental groups are seen as the antithesis to development, to business interests and to the needs of many user groups. However, environmental groups play an increasingly important non-adversarial role in demonstrating how conservation and sustainable use can be accomplished, through on-the-ground conservation projects that benefit users, community groups, business and national interests. Due to the ballooning scope of global fisheries problems and conflicts, environmen-
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tal groups are becoming more and more involved in fisheries management and conflict resolution. In tackling fisheries issues, most organizations attempt to base their projects and advocacy on the best available scientific information. These groups sometimes undertake in-house scientific research, predictive modelling and meta-analysis. However, in most cases, the non-governmental organization (NGOs) are recipients of scientific information and act as a liaison between the scientific community, decision makers and the public. Key scientific information that forms the underpinnings of campaigns and field projects addresses three facets of sustainability: (i) the levels of resource removal that can be realized without adverse impact on the ecosystem given the particular environmental condition of the ecosystem at the time of harvest; (ii) the least invasive means by which that harvest can be undertaken at desired levels of harvest, such that habitat impacts and by-catch are minimized; and (iii) the most appropriate stocks for large-scale harvest – those being stocks that are not the sole representatives of a deme or particular genetic structure and those for which the ecological role of the species either is not critically important or redundant (Agardy, 2000b). If a common environmentalist response to fisheries-induced loss of marine biodiversity can be said to exist (and this is a dangerous assumption, given the diversity of groups and their approaches), it is to synthesize existing information, communicate it and advocate for change in policy and regulations. In addition, some groups go beyond fisheries-by-fisheries management reform to advocate: (i) shifting the burden of proof when evaluating fishing impacts on ecosystems; and (ii) establishing strictly protected marine reserves to further our understanding and to protect species, habitats and ecological processes. Shifting the burden of proof has received recent attention in the fisheries management community (Dayton, 1998), but there remain misconceptions (Agardy, 2000b). Much of the conservation community advocates shifting this burden of proof in evaluating the prospective impacts of new fisheries, expanded fisheries or new
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technologies, and for regulators to permit such fisheries development only when proof of no likely impact exists. From the environmentalist’s or conservationist’s perspective, solutions lie not in shutting down fisheries but rather in modifying the way we undertake management, and in using public awareness to help raise political will to conserve marine systems. Coupling current consumer awareness and purchasing power with strong and effective management could indeed alleviate pressure on many marine species and allow their subsequent recovery. This is addressed further below. Additionally, environmental groups will need to recognize where there exists real willingness among governmental agencies and decision makers to protect areas needed for fish spawning, feeding and migration through marine reserves, as well as entering into enforceable international agreements to protect shared or common resources. By highlighting such potential successes, environmental groups can begin to shed their image of extremist adversaries, and help decisionmaking bodies implement effective and beneficial management regimes. Already a common thread is emerging from analysis of cases where fisheries management and marine conservation have succeeded. The central element in these initiatives seems to be a holistic approach – one that considers renewable living resources as part of a wider, interconnected ecosystem, one that evaluates all aspects of production or development, and one that treats humans as bone fide elements of living systems. The following sections will discuss these aspects of interconnectedness, and conclude with holistic prescriptions to achieve responsible fisheries in the future.
Ecosystem Interconnections and Ecological Costs of Fisheries There are ample data to suggest that fisheries exploitation affects not only target stocks but also communities of organisms, ecological processes and even entire ecosystems (Jennings and Kaiser, 1998; NRC, 2000; Sumaila et al., 2000). Marine ecosystems and
the substantial biodiversity they support continue to be threatened by human activity the world over (NRC, 1995). As downstream recipients of degrading impacts caused by poor land use practices and simultaneously under increasing pressure to supply natural resources and space to accommodate human needs, the world’s coastal zones and shallow seas are affected both directly and indirectly. Multiple and cumulative threats have already caused the loss of both species and genetically unique stocks of organisms and have undermined the functioning of many marine systems (Dayton et al., 2000). The drive to exploit living marine resources stems from an increasing reliance on fisheries-derived protein to feed burgeoning human populations, livestock and cultivated aquatic organisms (NRC, 1995). This growing demand is exacerbated by poor agricultural practices that reduce the potential of terrestrial sources to meet these protein needs. Overexploitation stems not merely from need but from the tragedy of the global commons, i.e. the inability of governments adequately to regulate use of common property resources (Hardin, 1966), or to recognize valid and sustainable methods of communal management (McCay and Jentoft, 1996; Bavinck, 2001). Commercial fishing operations, whether large-scale industrial fisheries or small-scale operations (Jamieson, 1993; Holt, 1998), commonly overexploit stocks, in some cases collectively causing trophic mining (Pauly et al., 1998). Though the list of marine endangered and threatened species pales in comparison with that of terrestrial and freshwater systems, marine biodiversity is being lost at an alarming rate as genetically unique populations of marine organisms are extirpated (Vermeij, 1993; Dayton et al., 1995; Tegner et al., 1996). Even for cosmopolitan species, this reduction in genetic diversity is damaging (Jennings and Kaiser, 1998). So-called serial mismanagement of fisheries extends beyond overexploitation effects to include exploitation methods that compromise marine biodiversity. Fishing methods commonly used to catch highly valued species selectively affect many other species – not only fish, but also sea turtles, sea birds, porpoises and unutilized finfish (Dayton et al.,
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1995). These incidentally caught and wasted species, some of which are already endangered, may constitute a higher percentage of catch than the target species and, in some cases, nearly 30 times more by weight (Alverson et al., 1994). Surface longlining contributes to the mortality of thousands of seabirds annually, while mid-water longline has been implicated in the dramatic population decline of the leatherback turtle, and shrimp trawling has dramatically reduced populations of other sea turtle species (NRC, 1995). Habitat alteration, as exemplified by bottom trawling, which rakes the benthos, kills epibenthic plants and animals and interrupts key ecological processes, is also problematic (Auster, 1998; Engel and Kuitek, 1998; Watling and Norse, 1998). To understand the extent of the marine biodiversity crisis and the role of fisheries in it, a greater understanding of how selective removal of target species affects ecosystem health and productivity must be attained. By concentrating harvest on top predators (e.g. mako shark, billfish, bluefin tuna), our fishing practices dramatically affect biological communities by causing cascading effects down food webs that decrease diversity or productivity (Steele, 1998). Removal of such apex predators need not be at industrial scales to result in these effects, since many of these species are naturally rare or distributed patchily. At the same time, decreases in the abundance of valuable apex predators and other species high in the food chain have caused fishers to target less valuable resources at lower trophic levels (Pauly et al., 1998). Because of their lower inherent value, such fishing is undertaken with increased intensity, such that entire trophic levels can be affected in a phenomenon known as trophic mining. Decline in abundance of primary consumers removes important forage species for organisms higher in the food web, again with cascading effects (Dayton et al., 1995; Jennings and Kaiser, 1998). Such altered ecosystems may have impaired function and be unable to replenish lost resources. The conservation community is also concerned that in determining what constitutes sustainable levels, methods and targets of fisheries exploitation, we consider changing
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environmental conditions and contexts. For instance, resource removal that may have minimal impacts on ecosystem function and overall biodiversity in a relatively pristine ecosystem potentially could have devastating effects in ecosystems already stressed by pollution, eutrophication and alterations to primary or ecologically linked habitats (Zaitsev and Mamaev, 1997). When these ecological impacts caused by fisheries operations are coupled to general environmental degradation, such as the eutrophication of coastal waters, toxic pollution or global climate change, the capacity of marine systems to support sustainable fisheries is reduced (Costanza et al., 1993). Even more importantly, when essential habitat is lost, as in the conversion of wetlands or nursery areas for coastal development, the critical threshold levels inevitably move down (Dayton et al., 2000). The paradox is that marine ecosystems increasingly are less able to support demand, even as demand continues to increase. For this reason, a reasonably holistic approach to understanding the resource in its ecological context will require understanding changes in dynamics and thresholds as environmental conditions change. Marine protected areas can play a vital role in increasing this understanding in that they can provide control sites for experimental manipulation of key parameters. None the less, we must acknowledge that the work of evaluation into what constitutes sustainability is work that by its very nature can never be finished. Environmental groups have for some years been demanding better information on the true, ecosystem-wide impacts of fisheries activity, particularly in cases where new fisheries are being launched, where major gear modifications are taking place or where major expansion of fishing effort is occurring. Most groups also advocate greater use of marine protected areas and fisheries reserves as a tool to strengthen management and to provide control sites to further scientific understanding and promote adaptive management. Finally, environmental groups have played a key role in developing case studies where government bodies work in concert with user groups and communities to demonstrate how
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co-management can be achieved. In pushing for these approaches to marine conservation, environmentalists underscore the need for a holistic perspective on conservation problems, and a holistic approach to their solution.
Production Interconnections and Environmental Costs of Fisheries One of the great failings of both resource managers and environmentalists addressing fisheries issues has been to focus on only a single aspect of marine fisheries production, namely resource removal. However, fisheries affect the environment – both terrestrial and marine – at many points along the production chain. Resource removal can have dire consequences for marine ecosystems when harvest levels exceed maximum sustainable yields, when certain trophic levels are heavily targeted and when the method of capture damages habitat or causes excessive by-catch, and these effects are increasingly well documented. However, fish product processing can also have severe environmental effects, occurring worldwide, from tuna processing plants in Pacific atolls to improperly regulated fish meal plants in the high temperate latitudes. As but one example, unregulated effluent discharge of fish waste into Madang Lagoon on the north coast of Papua New Guinea has dramatically affected a region of exceptionally high biodiversity, and threatens to affect the health of local communities as well. Recent lawsuits against processors suggest the environmental community is waking up to this problem, but it has certainly not yet resonated with the public or with decision makers. Packaging of fisheries products can also incur environmental costs, especially when packaging plants occur in countries with minimum industry oversight and pollution regulations. These production impacts on the environment are substantial, and hold equally for aquaculture products and for wild-caught fisheries products. Similarly, fisheries products travel the world in greater frequency than even international conservationists – and the pollution and energy costs of this aspect of
international fisheries trade must be part of the equation. A holistic view of fisheries management should thus extend not only to ecosystem impacts caused by the fishing itself, but beyond to impacts along the entire production chain. The environmental community is by no means united on this issue. Many groups have championed the efforts of the Marine Stewardship Council (MSC) and other certification programmes, which work to harness consumer interest in environmental issues to provide incentives for fisheries to practise their trade in a more sustainable manner. Green-labelled or eco-labelled products are those that can be certified independently as sustainably harvested. Note that the number of fisheries that can achieve even this simplistic level of certification are few and far between; at the time of writing, only seven fisheries had received MSC certification. However, an even greater problem may be that certification ignores the rest of the production chain – and in so doing may totally mislead consumers into thinking the products they are purchasing have exacted no environmental costs whatsoever. The can of skipjack tuna that comes from a fishery in which by-catch is minimized, dolphin encirclement is banned and habitat impacts are nonexistent none the less may have contributed to massive environmental degradation and loss of biodiversity due to processing operations far from the harvest site. Ensuring the sustainability of fisheries operations requires more than merely effective control of resource extraction. Even if management measures for marine fisheries were perfect, these fisheries would not replenish themselves without additional conservation measures that protect key habitats. Many of these key habitats are coastal, and are degraded both directly and indirectly by human activity. Such impacts do not affect solely coastal species, since even many pelagic species have life stages at least partially spent in nearshore areas. While anadromous and catadromous species clearly require intact coastal and even freshwater habitats, other species, including sharks, many pelagic and neritic fishfish, and many commercially important kinds of shellfish spend early life
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history stages in estuaries and other nearshore areas. Even for species that remain pelagic throughout their life, food sources often originate in estuaries or other coastal areas. Paradoxically, in some parts of the world, fisheries-related industries are destroying the critical habitats upon which the target species (and other equally important, though not as economically valuable, species) depend. This is the case with aquaculture (often touted as the best way to relieve pressure on wild stocks, a claim rarely substantiated (Golding and Triplett, 1997)), and fish processing and packaging industries. What can be done about this aspect of interconnectedness? First, understanding that these connections occur is key to a holistic and effective approach. Instead of asking the question, ‘How can fisheries management for target species be improved?’, it makes sense to ask, ‘What is needed to maintain these fisheries over time?’ Admittedly, this question is difficult enough to answer in the narrow realm of setting limits to exploitation. The question becomes even more complex – though no less important – when one attempts to apply it to the fisheries industry along its entire production chain. Secondly, raising awareness about these connections, and highlighting the importance of critical areas that sustain fisheries, must be a challenge to which environmentalists rise with enthusiasm. Groups that advocate certification and green-labelling will have to address the broader complexities and realities of measuring fisheries sustainability. Finally, environmental groups can catalyse the protection of critical habitats through regulatory reform and marine protected area designations, in order to complement conventional fisheries management measures and move us towards effective conservation.
Human Interconnections and Social Costs of Fisheries Environmentalists often have been labelled ‘nature-centric’ – a disparaging term used to contrast the value systems of those who would put biodiversity conservation above
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considerations of human need. While it is true that many environmental groups see their niche in extremist positions that provide counterbalance to the wise-use movement and other equally extreme antienvironmental philosophies, most groups practise a human-centric conservation that recognizes human needs, especially those of marginalized coastal communities and traditional fishers. These groups work to understand the social costs of massive economic development, the tension between artisanal and commercial fisheries, and the social impacts of large-scale fisheries and aquaculture operations that act to disempower local peoples and steer profits away to large multi-national corporations (Kurien, 1978; Kurien and Achari, 1994). In fact, when developing small-scale models of integrated coastal management, environmental groups only seem to succeed when these more local concerns have been appraised and addressed adequately. When humans are not considered bona fide elements of ecosystems, and human needs are ignored in either the rush to develop or the defensive move to protect the environment, social conflicts worsen. The resulting social effects include social disruption (Acheson, 1987; Berkes, 1987); migration and resulting interference with traditional sustainable patterns of resource use in areas of in-migration; environmental refugee movements that put pressures on scarce resources or vulnerable ecosystems; undermined national security; and resource use conflicts that sometimes escalate into resource wars (Poggie and Polnac, 1988; McGoodwin, 1990). Again, we must recognize these interconnections before we can help turn unsustainable fisheries into responsible ones. In places where artisanal and commercial fisheries clash with intensity, such as much of coastal West Africa, documenting and empirically assessing the problem is crucial to being able to deal with it. Further, resource management authorities must look to ensure equitable sharing of benefits whenever possible (De Fontaubert et al., 1996), and explore ways to address more local needs in addition to national economic interests. Co-management has much potential in this regard, and
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may prove to be the way forward in many settings (Tamanaha, 1993; Stroud, 1994). Co-management that recognizes the legitimacy of traditional management regimes, such as those demonstrated by marine tenure or community-imposed limited access (Ruddle and Johannes, 1985; Ruddle, 1988; Ruddle et al., 1992), have especially good chances of succeeding (Dyer and McGoodwin, 1994). However, this does not automatically invoke a push for pluralistic approaches to marine resource use and management, as described by Hooker (1975) and others, since pluralism can lead to chaos and erode responsibility to institute effective management. Successful models of fisheries management that have addressed social issues well seem to be those that clearly define roles and responsibilities of local communities and government authorities, and recognize the benefits of participatory planning and management (Pinkerton, 1987; Jentoft and McCay, 1995).
Holistic Solutions that Recognize Interconnections Nothing described in these pages is new, nor are holistic approaches that move us away from previously faulty sectoral management an invention of enlightened environmentalists. However, few scientists or managers seem willing to talk about these complexities – leaving a vacuum that the environmentalist community can and should work to fill. The complicated nature of marine conservation has heretofore shrouded our collective understanding of issues and made it impossible for us consistently to apply or even advocate effective solutions. We seem, if nothing else, paralysed by the complicated position that humans occupy in the natural world, and keep falling back on simplistic approaches that fail us time and again. Given the current situation, the roles that environmentalists can play in brokering information and communicating it in ways that the public can grasp are increasingly critical. In adopting a holistic approach to describing human impacts on marine environments, we
need to be objective, and scientifically rigorous, but we also must present the big picture (Caddy, 1993). Recognizing linkages is a prerequisite and, hard as it may be, we need to weed through all human impacts that affect marine systems and fisheries potential simultaneously. It will be critically important not only to recognize and prioritize threats to ecosystem health and function, but also to identify the underlying drivers behind these threatening human activities (Table 5.1). Perhaps the biggest challenge here will be to describe these complicated situations in ways that will be fathomable, yet at the same time will not act to frighten people away by the enormity of it all. The other role of environmental groups in catalysing more responsible fisheries and overall coastal and marine management is in field projects and interventions that demonstrate how these principles can be applied effectively. Marine protected areas are key here, because they are most often the venues for such demonstration projects. In addition, marine reserves and other protected areas are crucial in serving as benchmarks and baselines in furthering our understanding of how ecosystems function, and how humans affect such functioning.
Marine Protected Areas as a Supplement to Conventional Management Marine protected areas (MPAs) are being selected increasingly from the portfolio of options available to marine resource managers, largely because conventional measures to manage fisheries and conserve marine ecosystems have failed repeatedly (Agardy, 1994). This failure has started to enter the realm of public consciousness as signs that mismanagement affects consumers as well as fishermen have become apparent. Limiting fisheries management to controls on quantity of effort or catch ignores the potentially significant impact that fisheries activities have on ecosystems and their function. The use of spatial and temporal regulations, as made possible by area closures, ensures that
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Table 5.1.
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Types of threat to marine ecosystems, drivers and possible policy responses.
Type of threat Habitat loss or conversion Coastal development (ports, urbanization, tourism-related development, industrial sites)
Driver
Policy responses
Requirement for environmental impact assessment (EIA), technology transfer and training in planning, tourism policies, identification of ecologically critical areas, marine protected areas (MPAs) Shift to market economies, demand MPAs, training for alternative Destructive fisheries methods, green-labelling and for aquaria fish and live food fish, (dynamite, cyanide, other certification policies, cost–benefit poison-fishing, bottom trawling) increasing competition in light of analyses and awareness raising diminishing resources Access to and training in use of Lack of alternative materials, Coastal deforestation alternative materials, cost–benefit increased competition, poor (especially mangrove analyses and awareness raising national policies deforestation) Mining (coral, sand, minerals, maintenance dredging) Civil engineering works
Environmental change brought about by war and conflict
Aquaculture-related habitat conversion
Habitat degradation Eutrophication from land-based sources (LBS) (agricultural waste, sewage, fertilizers)
Pollution: toxics and pathogens from LBS
Pollution: dumping and dredge spoils
Population growth, poor siting due to undervaluing and lack of knowledge, poorly developed industrial policy, tourism demand, environmental refugees and internal migration
Lack of alternative materials, global Cost–benefit analyses, technology transfer for employing commons perceptions alternatives Mainstreaming marine Transport and energy demands, conservation into national policy, poor public policy, lack of cost–benefit analyses, awareness knowledge about impacts and raising their costs Social policy reform, policies to Increased competition for scarce deal with immigrants, limited resources, political instability, access (in special cases), inequality in wealth distribution mainstreaming marine conservation into foreign policy International security of trade and International demand for luxury items and regional demand for fish its impacts, international agreements, training in food, decline in wild stocks and environmentally sensitive decreased access to fisheries (or aquaculture, certification inability to compete) Urbanization, lack of sewage treatment or use of combined storm and sewer systems (CSS), unregulated agricultural development, loss of wetlands and other natural controls
Lack of awareness, increasing pesticide and fertilizer use (especially as soil quality diminishes), unregulated industry Lack of alternative disposal methods, increased enforcement and stiffer penalties for land disposal, belief in unlimited assimilative capacities, waste as a commodity
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Regional agreements involving both sources and sinks, common policies and actions on waste treatment, adequate financing for waste treatment and run-off mitigation (river basin and watershed management), artificial wetlands construction and stricter protection of existing wetlands International standards and compliance on toxics, awareness raising Technology transfer and training in appropriate methods for dumping, enforcement of treaties (LDC and MARPOL), financing and construction of wasteaccepting stations in ports of call continued
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Table 5.1.
Continued.
Type of threat
Driver
Policy responses
Pollution: shipping-related
Substandard shipping regulations, no investment in safety, policies promoting flags of convenience, increases in ship-based trade Demand for electricity and water, territorial disputes
International regulations on ship hulls and safety measures, establishment of Sensitive Sea Areas (SSAs) Cost–benefit analyses, technology transfer for water re-use Ballast water discharge regulations, active controls against spreading of invading species
Salinization of estuaries due to decreased freshwater inflow Alien species invasions
Global warming and sea level rise
Overexploitation Directed take (low value, high volume) exceeding sustainable levels
Directed take for luxury markets (high value, low volume) exceeding sustainable levels Incidental take or by-catch
Directed take at commercial scales decreasing availability of resources for subsistence and artisanal use
Lack of regulations on ballast discharge, increased aquaculture-related escapes, lack of international agreements on deliberate introductions Adherence to emission Controls on emissions lacking, standards, sea level rise poorly planned development (vulnerable development), stressed defences ecosystems less able to cope Demand for subsistence and market (food and medicinal), industrialization of fisheries, improved fish-finding technology, poor regional agreements, breakdown of traditional regulation systems, lack of enforcement, subsidies Demand for specialty foods and medicines, aquarium fish and curios, lack of awareness or concern about impacts, technological advances, commodification Subsidies, by-catch has no cost
Unempowered local peoples, breakdown of traditional structures
the benefits of management are extended beyond just the target stock to wider segments of ecosystems themselves (Davis, 1989). Thus fully protected closed areas, when used in conjunction with other forms of regulation, can move fisheries management away from largely ineffective speciesby-species fisheries management to more ecosystem-based conservation (Jamieson and Lessard, 2001). MPAs are fundamentally different from terrestrial protected areas, though whether these differences are in kind or degree is
International agreements (global codes of conduct, regional management) and enforcement, consumer awareness, quotas, limited access
International pressure on key demand hot spots, trade embargoes, quotas, limited access, MPAs Regulations on gear, MPAs for critically important habitats, seasonal closures MPAs, campaign to highlight the problem, artisanal fishing cooperatives
debatable. An important factor underlying these differences is the nebulous nature of boundaries in the fluid environment of the sea (Steele, 1998), making it difficult to attach boundary conditions to marine ecological processes and threats to those processes. While this is also true for inland freshwater systems, these ecosystems commonly have discernible outer bounds and distinct thermoclines that delimit biotic communities. To a far greater extent than on land, it is impossible to ‘fence in’ living marine resources or the critical ecological processes
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that support them, just as it is impossible to ‘fence out’ the degradation of ocean environments caused by land-based sources of pollution, changes in hydrology, or ecological disruptions occurring in areas adjacent or linked to a protected area. Long distance dispersal and the vastness of linkages between critical habitats in a coastal and marine ecosystem require comprehensive management of all its parts (Caddy and Sharp, 1986; Costanza et al., 1993; Mooney, 1998). The open nature of coastal and ocean areas exists as a spectrum ranging from relatively fixed and ‘land-like’ systems to highly dynamic and complex systems. Coral reef ecosystems, for instance, harbour organisms that are largely confined in their movements to the specific habitats of reef, surrounding soft or hard benthos, and coastal wetlands (Hatcher et al., 1989; Roberts, 1995a). The structural framework for reef systems is fixed in place and can be mapped, much as a tropical forest provides a relatively fixed framework for the interactions of the forest community. The functional links between the water column in reef areas and the benthos are strong, so one can treat the marine organisms together with reef structures themselves (Bohnsack, 1998). In contrast, temperate open ocean systems such as estuarine–gulf–banks complexes are highly dynamic and in no way ‘fixed’. Here, living marine resources move in space and time according to physically dominated, largely non-deterministic patterns (de Groot, 1992). The ecology of the water column is not strongly linked to that of the benthos, and physical reference points for the system cannot be mapped easily. However, even highly dynamic open ocean systems can benefit from MPAs, as long as the dynamics are considered in planning design (Hyrenbach et al., 2000). Thus, a wide array of system types presents a challenge to conservationists and resource managers, requiring that protected area measures be appropriate to the system in question (Agardy, 1997). Since identification of critical areas, public education and enforcement are achieved more easily in coral reef and other relatively ‘fixed’ ecosystems, MPA work has proliferated in these systems (Quinn et al., 1993; Agardy, 1995; Jennings and Polunin, 1996; van
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Ginkel, 1998). That is not to say, however, that closed areas in temperate and boreal systems are unfeasible, nor should it suggest that potential benefits of such protected areas are fewer in non-tropical systems (Auster and Malatesta, 1995). It merely suggests that random application of terrestrial models to the marine environment may not result in a viable means of protecting resources and the underlying ecology that gives rise to them. New paradigms are needed, and the newest MPAs reflect both acknowledgement of fundamental differences between marine and terrestrial systems, and existence of new information and planning technologies that can optimize MPA design. Clear evidence exists that MPAs can be designed to help make fisheries and coastal management more effective (Guenette and Pitcher, 1999). In the last 5 years, new, rigorous and defensible evidence has emerged to show that MPAs do indeed improve fish yields while conserving biological diversity more generally (Jennings and Polunin, 1996; Jennings and Kaiser, 1998). These benefits have included increased fish stock size inside the reserve, as well as spillover effects in which fish populations have also increased outside the reserve in the Caribbean (Roberts and Polunin, 1991; Reynard, 1994; Roberts, 1995b; Rakitin and Kramer, 1996), Philippines (Russ and Alcala, 1996, 1997) and in numerous other areas (Castilla and Duran, 1985; McCormick and Choat, 1987; McClanahan and Shafir, 1990; Ballantine, 1991; Dugan and Davis, 1993; Bohnsack, 1996a,b; McClanahan and Kaunda-Arara, 1996). The ideal situation seems to be the establishment of closed areas within the context of a larger multipleuse protected area such as a coastal biosphere reserve, marine sanctuary or other large-scale MPA. Area closures that are designated specifically to protect ‘seed banks’ or sources of recruits are becoming more common (Roberts, 1995a; Russ and Alcala, 1996). The link between certain coastal areas and maintenance of marine fisheries resources has been clearly established (Odum, 1984; LozanoAlvarez et al., 1993). Although recruitment dynamics are often complex and seemingly unpredictable (Holt, 1990; Fogarty et al., 1991),
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dispersal pathways for recruits can be identified readily in some ecosystems (e.g. Gaines and Bertness, 1992). The important biological processes that support fisheries productivity include spawning, migratory pathways, feeding, settlement and concentrated feeding (de Groot, 1992). Such ecologically critical processes in nearshore ecosystems often are concentrated in areas that can be identified easily by physical parameters such as reef formations, extensive shallow water areas, certain types of coastal wetlands, continental shelf breaks and frontal systems. An additional role for MPAs is to serve as control sites for scientific research and experimentation, especially to foment true adaptive management in which the controls on use serve as experiments to test management effectiveness. Without control areas and rigorous hypothesis testing, management cannot be adaptive in the true sense of the term. Unfortunately, many managers and the public at large tend to think that any management that is revised over time or is flexible is adaptive management. Without the science behind it, such flexibility is, in essence, only bet-hedging. Although the usefulness of closed areas and harvest refugia increasingly is being documented as resource managers turn to this management option, there are undeniable constraints to the broad applicability of this measure (Allison et al., 1998; Russ and Alcala, 1998). Limited scientific knowledge on population replacement rates, dynamics, recruitment patterns and impacts of fishing pressure on ecosystem function have all been used as excuses hindering establishment of no-take reserves. The stochastic nature of many marine systems also undermines the usefulness of this approach, particularly if closed areas are treated as static and immutable entities instead of flexible management measures. There may also be social constraints limiting the applicability of closed areas. The fishing industry is notoriously hard to regulate, precluding the acceptance of any new, potentially effective management tool. Closures having a scientific basis can be viewed by the fishing community as exclusionary practices that are somehow rooted in social discrimination. This predisposes user groups to reject the idea of
area closures even before they have the chance to discover exactly why and how these would be beneficial to them. Despite these constraints, closed area designations can be an effective tool to complement other fisheries regulation if they are carefully planned and grounded in good scientific understanding of ecosystem dynamics. The prospect of increased management and enforcement that the implementation of closed areas entails will not be embraced readily by most fishing communities, but only until the effectiveness of such areas in maintaining and even increasing catch is demonstrated. Managers using this technique will have to be responsive to changes in scientific information, in the status of the resources and in management needs in order to make MPAs optimally effective. This will require adopting management techniques that require refinement based on periodic reassessment of zone boundaries, regulations and overall extent of the protected area. MPA design and implementation have clearly entered a new phase of sophistication as more rigorous approaches to protected area planning have emerged, and as experiential learning over several decades has increased. Certain scientifically rigorous criteria now guide the selection of MPA sites as well as the subsequent size, shape and management regime of individual protected areas. These criteria (Table 5.2) relate directly to the specific objectives that the protected area or protected area system are established to achieve. Such objectives include, inter alia, habitat protection for overall biodiversity conservation, fisheries management and stock enhancement, naturebased tourism development, protection of traditional use and tenure, and scientific research. MPAs can be classified according to these objectives with objective-specific subsets of criteria for selection and design. Specific examples are given for each subset, spanning the spectrum from small-scale, community-based marine protected areas, to large-scale protected areas and networks of protected areas administered by centralized government authorities. Finally, we must recognize that though serious advances have been made in MPA planning, the ‘science’ of MPA site selection and design is still
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Table 5.2.
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Relationship between marine protected area (MPA) objectives, size and design complexity.
Specific MPA objective
Relative size
Complexity
Protecting an endangered species Protecting a migratory species Protecting habitat from single threat Protecting habitat from multiple threats Preventing overfishing Enhancing stocks Protecting an area of historic or cultural interest Providing a coastal zone management (CZM) model or empowering local people Promoting marine ecotourism Providing site(s) for scientific research Conserving biodiversity
Small to medium Large (or network) Medium Medium to large Small Small to medium Small Small to medium
Simple Simple to complex Simple Complex Simple Simple Simple Simple to complex
Small Small Large (or network)
Simple Simple Simple to complex
something of an art, and neither hard and fast rules for optimizing design nor a model MPA can be said to exist.
Multiple-use zoning in large-scale protected areas MPAs are site-based conservation initiatives that run the gamut from small-scale, strictly protected fisheries reserves (no-take areas), to larger-scale marine parks and sanctuaries. Although usually thought of as purely in-water designations, MPAs can also have terrestrial and aquatic components. Larger-scale protected areas, especially those encompassing land and estuarine areas, generally use zoning to allow different levels and kinds of use in different areas. Such designations are known as multiple-use MPAs. When a functional approach is adopted, i.e. where the object of conservation is not a single stock of resources or a single species but the reef ecosystem and its processes, MPAs will have to be large and encompass many types of linked habitats (Agardy, 1994), or will have to be designed as part of a linked or integrated network. The large-scale, multiple-use protected area can demonstrate the practical application of ecosystem-based management (Hatcher et al., 1989; Costanza et al., 1993). The underlying ecology and movement of species define the outer boundaries for the area of protection, or management unit (Dayton et al.,
2000). In recognizing these linkages, MPA planners can work towards conserving ecosystem integrity, not just individual resources or ecosystem structure. Virtually all multiple-use MPAs that serve to maintain ecosystem integrity (as opposed to merely providing sites in which the segregation of uses can be accomplished in the effort to resolve user conflicts) rely on no-take or strictly protected core areas (Roberts, 1995a; Allison et al., 1998). These critical areas usually protect ecological functions such as breeding, dispersal and growth in populations of organisms that are crucial – by virtue of their economic value, ecological value or role as indicators (Guenette and Pitcher, 1999). This ecological information, however, is but a part of the types of information required for developing multiple-use areas. Equally important is socio-economic information on uses of the area, both current and prospective. Multiple-use MPAs that are of appropriate ecosystem scales, that contain management units grounded in ecology and that allow multiple uses by establishing zoning to protect areas deemed most critical, most sensitive or most amenable to monitoring and evaluation, are a crucial tool for conservation. In many cases, such protected areas provide managers the opportunities to test various management regimes and practise adaptive management (Agardy, 1997). In what may be more appealing to decision makers and government bodies, these MPAs can also serve as
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small-scale demonstration models for what can and should be done to integrate terrestrial, aquatic and marine management most optimally. While it is true that no model MPA can fit all circumstances of environment, society and economics, the process by which these multiple-use protected areas are designed and implemented can be thought of as a generic one. Whatever the specific objectives of a protected area, ecological information must be harnessed to determine, with as much certainty as possible, where the ecologically most critical areas are that need the strictest possible protection (Agardy, 2000a). Utilizing zoning to protect these key areas is analogous to how we employ medicine to keep us healthy: we target the vital organs first and foremost, and do everything we can to maintain them in their life support roles. So it is with estuaries, sea-grass beds, mangrove forests, sea mounts, migration pathways and other key elements of marine systems. However, because our understanding of marine ecology is still weak (Hillborn and Walters, 1992), we must out of necessity turn to unconventional sources of information. As planners and manager are becoming more comfortable using traditional knowledge to supplement what we understand from scientific study, this is leading to a more precautionary approach to marine conservation (Johannes, 1998). Thus large-scale multipleuse protected areas target ecological processes and conserve linkages, demonstrating how a functional approach is both possible and effective.
Marine reserve networks Even large-scale multiple-use MPAs cannot protect what is ecologically most critical on a regional scale. However, networks of marine reserves can be used to accomplish this. Such networks present little opportunity cost resulting from restrictions of use in any one local area, since these restrictions are spread out equitably over a much wider geographical area. In addition, the benefits accruing from use of strategic and well-
designed networks of reserves can be enormous – and are also spread across wide areas. For this reason, networks of reserves are now being viewed as an exciting new tool to conserve whole ecosystems or even ocean regions. Fundamental principles of ecology and biological oceanography underlie the way networks are designed and constructed. The open nature of marine ecosystems and the large-scale patterns of dispersal that characterize many species mean that terrestrial models of protected area planning become largely irrelevant when applied to ocean systems. Larval dispersal and subsequent recruitment is not only geographically widespread in many marine ecosystems, it is also a highly variable process in which non-linear dynamics and stochasticity plays a large role (Holt, 1990; Fogarty et al., 1991; Gaines and Bertness, 1992). Despite this natural variability, biologists have been able to study many systems to determine sources and sinks for larval recruits, in order to arrive at maps that show the most critical habitats to protect in order to conserve dispersal patterns and recruitment (Murray et al., 1999; Hyrenbach et al., 2000), although not all critical areas determined in this way are necessarily key habitats continually – some areas that are fundamental to constructing networks to preserve ecological integrity may be only seasonally important, such as sites for spawning aggregations of fishes (Johannes, 1998). New studies have shown that even very small marine reserves can have significant positive impact on marine biodiversity and productivity (Ratikin and Kramer, 1996; Russ and Alcala, 1996, 1997, 1998). Small MPAs that are linked in a systematic network that protects a large proportion of critical habitats or particularly important sources of recruits in a region provide even more benefits. Source sink modelling has been employed to determine the relative value of various sites within a single MPA such as the Great Barrier Reef Marine Park (James et al., 1995) and the Florida Keys National Marine Sanctuary (Bohnsack, 1996a). Such studies of connectivity are also important in the development of networks of reserves, as proposed by Roberts (1998) for the northwest Caribbean.
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Such systems of protected areas aiming to conserve overall ecosystem function will have to focus not only on intact and biologically valuable areas, but also on the most highly threatened areas or degraded areas needing restoration. This approach requires evaluating threats to ecosystems and the degree to which areas are degraded, in order to establish a system of MPAs allowing restoration of sites (and replenishment of resources) as quickly as possible. Though few systematic attempts to identify coastal and marine areas in need of restoration exist, the ongoing restoration programme for South Florida (including the Everglades area, Florida Bay and the Florida Keys) is a good example of an analytical approach to establishing a network of protected areas for restoration purposes (Murray et al., 1999). In some MPA examples, the restoration effort targets a single species or stock, as in the restoration of a historically overexploited fishery. Such protected areas are known as no-take zones, and they often become a starting point for subsequent more comprehensive and effective protected area management. The innovative new networks, however, attempt to amalgamate these strategies by identifying the ecologically most important and most threatened areas, together with those most likely to contribute to ecosystem functioning once restored. When threats to biodiversity are direct, such as destructive fisheries or habitat conversion, then it is easy to see how a protected area will have the potential to abate the threat (Roberts and Polunin, 1991; Dugan and Davis, 1993). However, MPAs can also help address indirect threats, such as land-based sources of pollution. The protected area becomes a tool for addressing indirect threats either when the scope of a single protected area or network of protected areas is expanded to encompass the area where such threats originate, or when the perceived value of the area ultimately affected by the threat is increased. In other words, MPA designations can help focus attention on areas being degraded from afar, and help generate the political will to address the source of the problem. In order to protect the most ecologically critical components of a large marine
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ecosystem or a coastal region, we must harness information to identify the key elements. For reserve networks that aim to maintain productive fisheries, the network of reserves may target key spawning areas, larval retention areas, nursery grounds, feeding areas and migration pathways (Bohnsack, 1996b, 1998). For networks that aim to maintain the health of key habitats within a region, the picture will be a bit more complex, requiring us to incorporate terrestrial reserves and river basin and watershed areas into the network. We must also look at what is threatening these critical areas, and tailor our network to stave off these threats. If the major threat is from overexploitation of fisheries, then strict controls must be placed in some areas to create no-take marine reserves (Bohnsack, 1998; Roberts, 1995a). If pollution is undermining the health of the system, then pollution input areas where pollution is controlled should become part of the reserve network. There is a general sense that these marine reserve networks can only utilize MPAs that are small and simply managed as no-take areas (Roberts, 2000). However, there is no reason why multiple-use MPAs cannot be used in networks (Agardy, 1995). Virtually all the world’s coasts and nearshore areas are characterized by conflict within and amongst user groups or jurisdictional agencies, or, at a minimum, a serious lack of communication between these factions (de Fontaubert et al., 1996). For instance, shipping and mineral extraction are uses that often conflict with recreational use of coral reef areas. Fishing, both commercial and subsistence, conflicts with skin and scuba diving and nature-based tourism. Different fisheries, at both commercial and artisanal scales, often conflict. In such cases of conflict, zoning can be used to accommodate a wide variety of user groups in relative harmony, and can be a tool for dispute resolution where conflicting uses clash. Networks must act to secure the rich and valuable coastal ecosystems of entire regions by focusing marine conservation action at the most ecologically critical places. This will entail providing involved parties with the means to put a systematically linked network of terrestrial and MPAs into place, through either establishment of new areas or
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strengthening of existing ones. It will also require that all prospective development impacts in a region are understood and evaluated systemically, and that integrated management of terrestrial, aquatic and marine systems is accomplished. The keys to these essential features are in turn harnessing science, raising awareness, building stakeholder constituencies and involvement, finding alternatives to harmful development practices, developing integrated coastal and tourism planning policies, and providing the means for long-lasting MPAs.
Corridor approaches Can even a well planned and extensive system of MPAs, linked in functional networks, really act to conserve large ocean regions? The answer is yes, but only if the effort does not stop with the establishment of individual protected areas within the network. Truly effective conservation requires a holistic approach – one that deals with the entire suite of impacts affecting marine and coastal systems. This means protecting the context in which sit the islands of protection created by MPAs. The corridor concept is the most sophisticated, comprehensive and integrative of all the three MPA approaches described in this review. ‘Corridor’, however, may be a misleading term, since the application in the marine realm is not strictly analogous to terrestrial corridor concepts. On land, corridors for conservation are created by identifying crucial areas for conservation, undertaking a gap analysis to highlight areas that fall outside of protection and implementing new protection regimes physically to link up existing protected areas and allow transmigration and genetic exchange. At sea, corridors describe a new way to package integrated coastal planning with a strong MPA component, and it is the direction in which all coastal planners should move in years to come. In creating marine corridors, planners seek to conserve the most critically important terrestrial, aquatic and marine habitats, which together function to maintain ecosystem
integrity and productivity. To do this, the best available scientific and traditional knowledge information is harnessed to identify the ecologically most critical areas (terrestrial, aquatic and marine) of a particular coastal region. The latter type of information, i.e. traditional knowledge or user-based knowledge, is crucial in most parts of the world, since scientifically derived ecological information is in short supply (Johannes, 1998). Once these most vital areas are identified, the corridor approach applies gap analysis to highlight conservation needs. Here the process is very much like that for constructing a strategic network of reserves, although the geographic scope of the enterprise is usually larger since terrestrial portions of the region are fully included. The corridor approach attains new heights of sophistication since it is built on establishment of reserve networks, but goes well beyond such networks. While land use planning and protected areas, both terrestrial and marine, are essential elements of a corridor approach, the method provides a way to protect the context in which these ‘islands of protection’ are found. The corridor approach can highlight areas where additional MPAs need to be established, or where existing MPAs needs to be strengthened or expanded. The collective protected areas constitute the sum total of vital organs. These critical habitats can then be linked by a ‘virtual corridor’, i.e. targeted policy reform that ensures that the connectivity is preserved and that these most vital parts are not degraded by direct and indirect impacts of human activity. The ecological linkages that connect life in the water with life on land necessitate that we do a better job of integrating management of watersheds, coastal lands and the oceans. Corridor approaches provide a framework to allow us to achieve that integration, in a strategic and efficient way. This tool allows place-based management to be implemented with the big-picture view in mind, bringing management interventions up to the scale of regional conservation. The natural extension of a holistic approach is fully to embody regional, ecosystem-based management (Gislason et al., 2000). In practice, such broad-based management is perhaps
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best exemplified in the Antarctic regional management framework of CCAMLR (the Convention for the Conservation of Antarctic Marine Living Resources) (Constable et al., 2000), which serves as a model for international cooperation around a conservation agenda. As regional approaches begin to be replicated in other large marine ecosystems and semi-enclosed seas, the environmental community will probably be called upon to watch for and facilitate the connections between site-level conservation and broader, multilateral policies and agreements. We must start to recognize these interconnections, too – between feet-on-the-ground interventions that effectively conserve biodiversity and at the same time confer benefits to stakeholders on the small scale, and head-in-the-cloud negotiations that represent the national interests of governments and the global interests of the international community on the very largest scale. When, and only when, local, regional and global level interests can be made to work in parallel or in harmony with one another will true marine conservation be achieved.
Conclusions Environmental groups play an ever more important role in public understanding of fisheries issues, in communicating technical information in lay terms, and in advocating for reform through a wide array of measures. In addition, these groups increasingly highlight and demonstrate ways forward, by assessing where and how fisheries management has worked to conserve both stocks and ecosystems, and where benefits are shared equitably while costs are minimized. As groups take on more and more of the latter function, their roles have shifted from adversaries to welcome partners catalysing change. However, environmental auditing of fisheries, and indeed all marine development, will only increase in the future and, unless fisheries reform is truly achieved, the fishing industry and government decision makers can continue to expect pressure
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from these increasingly powerful nongovernmental bodies. A review of some key environmental groups working in the fisheries field shows just how varied approaches to the fisheries problem can be. Many environmental groups lobby and advocate for policy reform and for formation of international instruments and agreements (such as FAO’s noteworthy Code of Conduct for fisheries – a non-binding agreement that describes holistically oriented principles for fisheries management). Some of these groups continue to play adversarial roles in their dealings with the fisheries industry, entering into environmental litigation when management bodies cannot be convinced to reform or are found to be ineffective in upholding regulations. Examples of these activities include efforts by the Natural Resources Defense Council, Environmental Defense, Greenpeace, the Ocean Conservancy (formerly the Center for Marine Conservation, based in Washington, DC) and the newly emergent Oceana (also based in Washington, DC). In contrast, the Worldwide Fund for Nature (World Wildlife Fund in the USA) has concentrated its recent efforts on trying to harness consumer awareness and purchasing power through the establishment of the Marine Stewardship Council. This independent body certifies capture fisheries that have been deemed ‘sustainable’ by a set of scientifically developed guidelines. In a slightly different vein, the Nature Conservancy (TNC) has taken on a particularly troubling set of fisheries: those that support the live fish trade centred in Southeast Asia. TNC has underwritten scientific study of the trade and its impacts, and has elevated the issue in the minds of the public in both western and eastern countries. In trying to stop the destructive use of cyanide in the trade, TNC partners work closely with other environmental NGOs – most notably the International Marine Life Alliance (IMA) and its broad network of partners operating in the countries where it undertakes training of fishers in less destructive techniques. Alternative methods also figure prominently in the work of Conservation International in its Gulf of California Program, where environmentalists have worked
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with the shrimp industry to install by-catch reduction devices and where it has initiated grass-roots level projects in sustainable mariculture. Finally, the World Resources Institute and the International Union for the Conservation of Nature (now the World Conservation Union) both engage in systematic, global-scale analyses to track trends in fisheries exploitation and prescribe remedies for overcoming environmental destruction and social inequity. Common to nearly all these groups (most of which are based in the USA – not because all the activity is occurring there by any means, but rather because of the author’s geographic bias) is the recent drive to lead by example. Many groups engage in assembling the scientific information on ecosystems and their use in order to help craft solutions to destructive or overexploitative fisheries, usually involving MPAs. Some of these MPAs are small scale and highly focused, such as the community-based marine reserves established in the Philippines and Indonesia. Others are very much larger scale, involving either zoned multiple-use MPAs that can accommodate a wide range of uses, or MPA networks. In perhaps the most useful demonstration of all, non-governmental environmental organizations are now working closely with governments within large regions to help foster corridor approaches that conserve ecosystems and promote sustainable uses of coastal and open oceans across wide geographies, using the collaborative and cooperative approaches that such regional conservation requires. This scaling up of effort, working with user groups instead of against them, seems the only logical way forward if we are to achieve truly responsible and beneficial fisheries in the future.
Acknowledgements Many thanks are due to Mike Sinclair for allowing a bona fide environmentalist into this fisheries forum, and for continuing to encourage open-mindedness in the fisheries management community.
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St. Lucia. Caribbean Parks and Protected Areas Bulletin 5(2), 5–7. Roberts, C. (1995a) Effects of fishing on the ecosystem structure of coral reefs. Conservation Biology 9(5), 988–994. Roberts, C. (1995b) Rapid build-up of fish biomass in a Caribbean marine reserve. Conservation Biology 9(4), 815–826. Roberts, C. (1998) Sources, sinks, and the design of marine networks. Fisheries 23(7), 16–19. Roberts, C. (2000) Selecting marine reserve locations: optimality versus opportunism. Bulletin of Marine Science 66(3), 581–592. Roberts, C. and Polunin, N. (1991) Are marine reserves effective in management of reef fisheries? Reviews in Fish Biology and Fisheries 1, 65–91. Ruddle, K. (1988) Social principles underlying traditional inshore fishery management systems in the Pacific Basin. Marine Resource Economics 5, 351–363. Ruddle, K. and Johannes, R.E. (1985) The Traditional Knowledge and Management of Coastal Ecosystems in Asia and the Pacific. UNESCO, Jakarta. Ruddle, K., Hviding, E. and Johannes, R.E. (1992) Marine resource management in the context of customary tenure. Marine Resource Economics 7, 249–273. Russ, G. and Alcala, A. (1996) Marine reserves: rates and patterns of recovery and decline of large predatory fish. Ecological Applications 6(3), 947–961. Russ, G. and Alcala, A. (1997) Do marine reserves export adult fish biomass? Evidence from Apo Island, Central Philippines. Marine Ecology Progress Series 132, 1–9. Russ, G. and Alcala, A. (1998) Natural fishing experiments in marine reserves, 1983–1993:
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community and trophic responses. Coral Reefs 17, 383–397. Steele, J. (1998) Regime shifts in marine ecosystems. Ecological Applications 8(1), S33-S36. Stroud, R.H. (ed.) (1994) Conserving America’s Fisheries. Marine Recreational Fisheries, No. 15. National Coalition for Marine Conservation, Savannah, Georgia. Sumaila, U.R., Guenette, S., Adler, J. and Chuenpagdee, R. (2000) Addressing ecosystem effects of fishing using marine protected areas. ICES Journal of Marine Sciences 57, 752–760. Tamanaha, B.Z. (1993) The folly of the ‘social scientific’ concept of legal pluralism. Journal of Law and Society 20(2), 192–217. Tegner, M., Basch, L. and Dayton, P. (1996) Near extinction of an exploited marine invertebrate. Trends in Ecology and Evolution 11, 278–280. van Ginkel, R. (1998) Zostera marina in dispute: management regimes in the Dutch eelgrass industry. In: Symes, D. (ed.) Property Rights and Regulatory Systems in Fisheries. Fishing News Books, Oxford, pp. 230–243. Vermeij, G. (1993) Biogeography of recently extinct marine species: implications for conservation. Conservation Biology 7, 391–397. Watling, L. and Norse, E. (1998) Disturbance of the seabed by mobile fishing gear: a comparison to forest clearcutting. Conservation Biology 12(6), 1180–1197. Wilder, R.J., Tegner, M.J. and Dayton, P.K. (1999) Saving marine biodiversity. Issues in Science and Technology Spring 1999, 57–64. Zaitsev, Y. and Mamaev, V. (1997) Marine Biological Diversity in the Black Sea: a Study of Change and Decline. UN Publications, New York.
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Ecological Geography as a Framework for a Transition Toward Responsible Fishing Daniel Pauly, Reg Watson and Villy Christensen Fisheries Centre, University of British Columbia, Vancouver, Canada
Abstract Meeting the widely expressed requirement that fisheries should somehow be managed on an ‘ecosystem basis’ implies that fisheries-relevant ecological processes, and the fisheries themselves, need to be documented in the form of maps. This allows recovery, in intuitive fashion, of at least some of the many dimensions of the complex ecosystems in which the fisheries are embedded. The implied transition, in fisheries science, from bivariate time series, to maps as major heuristic devices has a number of implications – some obvious, some less so – of which a number are here discussed and illustrated. Among the issues covered are: (i) the requirement for a consensus taxonomy of large marine ecosystems; (ii) the need to construct fisheries catch maps in the absence of positive records of what was caught where; (iii) the proper identification of one’s audience; and (iv) the mapping of marine protected areas and reserves. The seriousness of the fisheries crisis is emphasized, and the case is made that fisheries, if ever they are going to achieve some measure of sustainability – however defined – ultimately will have to be limited not only through the amount of effort they can effectively deploy, but also limited in space, leading to a change to the defaults under which fisheries operate, currently set such that all aquatic wildlife can be exploited, if under some restrictions.
Introduction Fisheries worldwide are in serious trouble. There is perhaps no need to document this, but we shall still present a single graph, a time series of global marine catches, with and without the catches from China and of Peruvian anchoveta, which jointly mask the clear declining trend evident in the rest of the world’s fisheries and species (Fig. 6.1). This crisis has many aspects and proposed solutions, and one of the latter is the widely expressed requirement that fisheries
management should somehow be put on an ecosystem basis – even though what this means is not yet very clear (NRC, 1999). The most common exhibits, in fisheries science, so far have usually been time series of key variables, e.g. catch, fishing mortality or spawning biomass. Such time series are usually hard to assemble, and their value increases with time (i.e. with the number of generations they encompass); hence the enormous value (in terms of both the costs they embody and the insights they led to) of the recruitment time series assembled by R.A.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Fig. 6.1. Trend in world marine fisheries catches as reported by FAO, with and without: (i) the catches from China (including Hong Kong and Macau, but excluding Taiwan) from Statistical Area 61 (Northwestern Pacific), which are massively over-reported (Watson et al., 2001); and (ii) the catches of Peruvian anchoveta, whose fluctuations largely reflect El Niño events. Removal of these two series unmasks a strong declining trend for the rest of the world’s fisheries, confirming the perception of widespread fisheries failures.
Myers and collaborators (e.g. Myers et al., 1999). Indeed, the bivariate plots representing time series serve among fisheries scientists as a key heuristic device: we work hard to assemble them, show them to colleagues (see Fig. 6.1) and jointly ponder on their features, such as the amount of contrast they do or do not incorporate (Hilborn and Walters, 1992). Also, a huge number of methods have been developed to analyse such plots, ranging from time series analysis (Chatfield, 1984) and other statistical methods (regression, etc.), to simulation models, e.g. ECOSIM (see Walters et al., 1997), designed to produce (i.e. predict) time series. However, fisheries, embedded as they are in natural ecosystems, and relying as they do on natural fluxes of these systems, depend on the features of places. Thus, while we emphasize the variability of fisheries in time, we tend to lose track of their variability in space (Samb and Pauly, 2000). Indeed, we hardly use maps to discuss fisheries (except for tunas, see below). Maps, clearly, will be an
important part of ecosystem-based management – though obviously they will not be all. Maps were crucial to the emergence of the modern world, as they catalogued the countries newly discovered by the European powers, and the best routes to their riches. The emergence of physical oceanography as a discipline of its own was also mediated by maps; indeed they are the currency that Commodore Matthew Maury (1806–1873), one of the founders of physical oceanography, used to ‘pay’ for the current, wind and depth observations that mariners sent him. Even our humble science of fisheries used maps to represent some of its newly acquired knowledge (Garstang, 1909; Fig. 6.2). Why we later neglected the device so successfully used by this and other pioneers of fisheries science to summarize their knowledge on the biology of North Sea or other fish need not be pursued here. What we can do, however, is to point out that the availability of powerful, PC-based geographical information system (GIS) technology makes it possible for maps to return to the central role that they had formerly in
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Fig. 6.2. Schematic representation of the distribution of plaice (Pleuronectes platessa) in the North Sea, illustrating the key aspects of its life history (modified from Garstang, 1909). Contrast this information-rich map with the text that would be required to convey the same amount of information.
summarizing knowledge in intuitive fashion, and hence this contribution. A number of issues will have to be sorted out, however, before fisheries maps become a routine tool in fisheries sciences:
• • • •
Ecosystem taxonomy and classification. Fisheries catch maps and related issues of scale. Using maps to reach new audiences. Maps and space-based fisheries management.
Ecosystem Taxonomy and Classification Using fisheries maps for putting fisheries in an ecosystem context assumes that some
agreement exists as to the definition and location of marine ecosystems. Indeed, without a prior definition of ecosystem boundaries, there is a real danger that the fisheries to be studied will themselves be used to define the boundaries of ecosystems pragmatically, as happened earlier with traditional biogeography, in which the distribution of the diverse groups mapped by specialists led to the definition of taxa-specific geographies, all mutually incompatible (see Ekman, 1967). Moreover, these taxa-specific geographies ended up being useless, even to those who had proposed them, due to the circularity of their definition. Thus, if the distribution of species within a given taxon defines a system of, say ‘provinces’, then features of these provinces diverging from what would be suggested (given the underlying physical
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structure of the ocean) will not be identified, nor any resulting improvement to the system of provinces. The way out of this circularity is, of course, to use predefined ecosystem definitions (and boundaries), independent of (but hopefully with deep affinities to) the taxa or processes (here: fisheries) that are being mapped. There are at present three broad taxonomies-cum-classification systems representing the world ocean at scales below that defined by entire oceans, or their major basins, namely:
• •
The system of 18 Statistical Areas used by FAO to report global fisheries catches. The Large Marine Ecosystems defined by K. Sherman and collaborators.
Box 6.1.
•
The system of four Biomes and 57 Biogeochemical Provinces described by Longhurst (1998) and presented in Box 6.1.
The FAO’s system of statistical areas at present is the only device routinely used for breaking global catches into geographic space (Fig. 6.3). These 18 FAO areas are rather large, and have boundaries based largely on political considerations. Therefore, they cannot be used directly to put fisheries into an ecosystem context; however, they do provide some constraints for the rule-based construction of fisheries maps described in Box 6.2. Initially, large marine ecosystems (LMEs) were only what the three words in their name
Biogeographical provinces.
Until recently, a ‘geography of the sea’ did not exist that was suitable for describing, in standardized fashion, the distribution of all marine organisms, despite a history of oceanographic research starting with the Challenger Expedition (1872–1876). Numerous maps did exist in which this or that oceanographic parameter, or the distribution of a few organisms had been used to draw a map of some sort (see, for example, Ekman, 1967). However, no tests were conducted of the ability of these maps to predict distributions other than those from which they were derived: circularity reigned supreme. Reasons for this are easy to imagine, from the excessive preoccupation of various specialists with their favorite taxonomic groups, to the absence, before the computer revolution, of analytic tools that were up to the task. However, the real reason is probably that developing a truly synoptic vision of the ocean was impossible before the advent of satellite-based oceanography. Satellites cannot ‘see’ very deep into the sea. However, what satellites do see is the very stuff that generates fundamental differences between ocean provinces: sea surface temperatures and their seasonal fluctuations, and pigments such as chlorophyll, and their fluctuations. Marine systems differ from terrestrial systems in that their productivity is essentially a function of nutrient inputs to the illuminated layers. This gives a structuring role to the physical processes that enrich surface waters with nutrients from deeper layers, such as wind-induced mixing, fronts, upwellings, etc. (Longhurst, 1995). Thus, the location, duration and amplitude of deep nutrient inputs into different oceanic regions – as reflected in their chlorophyll standing stocks – largely define the upper trophic level biomasses and fluxes that can be maintained in these regions. This is the reason why satellite images reflect fundamental features of the ocean, while maps based on the distribution of various organisms – even ‘indicator’ organisms – can only reflect second-order phenomena. T. Platt, S. Sathyadranath and A.R. Longhurst are among the first to have realized this, and thus their stratification of the ocean, and the estimates of global primary production based thereon, are far superior to earlier attempts. The system of biomes and biogeochemical provinces defined in the process was refined further in a book by Longhurst (1998), the review of which (Pauly, 1999) provides the basis for this box. One interesting aspect of this stratification (or classification) is that the biogeochemical provinces (BGCPs) in the ‘coastal biome’ thus defined largely overlap with the large marine ecosystems (LMEs) of Sherman and collaborators (Pauly et al., 2000; see Figs 6.4 and 6.5a and b). This correspondence should make it possible to integrate in a common framework the vast amount of geo-referenced information on marine biological processes that is now available, and finally to make widely available to practitioners the data that so many of them still claim we do not have.
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Fig. 6.3. System of 18 statistical areas used by FAO to report on global marine fisheries statistics. Note the large size of these areas, encompassing very different ecosystems and faunas.
Box 6.2.
Construction of fisheries maps (by R. Watson).
The records used for constructing detailed catch maps (here: by half a degree lat./long. cells) for an ocean basin (or worldwide for global maps) for a given year are based either; (i) exclusively on FAO statistics on the countries fishing in that basin, or on the FAO global statistics; (ii) on FAO statistics complemented with time series of discards, estimates of illegal catches, etc.; or (iii) statistics that substitute for those of FAO, e.g. International Council for the Exploration of the Sea statistics in the Northeastern Atlantic, complemented as in (ii) or not (A in Fig. 6.6a). These are processed as a set of database records by first disaggregating the statistics for the generalized group into records at lower taxonomic levels (B in Fig. 6.6a), as necessary for many countries where the reported catch composition is very aggregated, such as China, and using a catch composition interpolated from that of immediate neighbours with detailed statistics (here Taiwan and South Korea). Then, each taxon represented in a landing record is looked up in a database of species-specific spatial distributions that identifies the subset of spatial cells of the world’s oceans from which the catch record in question could originate. The country reporting (fishing) is then looked up in; (i) a database of fishing access agreements (updated from Anon. 1998); and (ii) a database identifying the exclusive economic zones (EEZs) of the world’s countries (see Table 6.1), which jointly identify the spatial cells that are available for that country to fish in (including the EEZ of other countries for which arrangements exist). The FAO area that the statistic was reported from is also used to identify a set of spatial cells from which the catch may originate. These sets of spatial cells are then compared and if there are no overlapping cells the landing is not allocated and an ‘error report’ is logged (see ‘no’ in Fig. 6.6b). Otherwise, the reported landing is assigned among overlapping cells in proportion to their areas. Thus, landing rates (t km−2 year−1) are accumulated in each cell as each record is processed (currently, we are able to allocate > 95% of the world catch to cells; the remainder reflects error reports whose resolution we expect to contribute to cleaning up the underlying databases. In this way, a grid map of landing rates is built up as each landing record is processed (D in Fig. 6.6a and b). Though each record is processed for the taxonomic level it is reported at (after disaggregation), the results can be reassembled into larger groups as required, e.g. for statistical models (E in Fig. 6.6a and b). Alternatively, the taxon- and cell-specific catch records can be multiplied by its corresponding market price, yielding maps of catch value, a new product for which the Sea Around Us Project envisages a large range of uses.
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imply, namely marine ecosystems defined so as to cover a large area (200,000 km2 or more). Gradually, however, and mainly due to the work of K. Sherman and collaborators, LMEs became restricted to an explicit list of 50 coastal entities (Fig. 6.4), with a dozen recently added (see www.edc.uri.edu/lme/default. htm), broadly defined by physical features (presence of shelves, coastal currents, fronts, etc.), and documented in a growing number of books (listed in www.edc.uri.edu/lme/ publications.htm). One major conservation-oriented nongovernmental organization (NGO), the World Conservation Union (IUCN), has endorsed the concept, with the intention of using it for reporting on marine biodiversity. Similarly, FishBase, the global database of fish, has linked all species of marine fishes in the world (~ 15,000 species) with the LME in which they occur (see Table 6.1). One important features of LMEs is that, while covering only 18% of the world’s oceans, they accounted for 75% of the world’s fisheries catches in 1999. These figures, based on the 50 LMEs listed in Sherman and Duda (1999), will increase when recalculated based on the 62 LMEs in Fig. 6.4. The most rigorous division of the world ocean, at least in terms of biological oceanography, is, however, the system described by Longhurst (1998), based on Platt and Sathyendranath (1988), Sathyendranath et al. (1995) and Longhurst (1995) (see Box 6.1 and Fig. 6.5a and b). At the highest level, this hierarchical classification is based on a division of the world ocean into four biomes. In the Polar biome, covering only 6% of the world ocean, vertical density structure is determined very largely by low-salinity water derived from ice-melt each spring. In the Westerlies biome, between the Polar fronts and the Subtropical Convergence, large seasonal differences in mixed-layer depth are forced by seasonality in surface irradiance and wind stress, inducing strong seasonality of biological processes, characteristically including a spring bloom of phytoplankton. The Trade-wind biome lies across the equatorial regions, between the boreal and austral Subtropical convergences, where low values for the Coriolis parameter, a strong density gradient across the permanent
pycnocline and weak seasonality in both wind stress and surface irradiance, result in relatively uniform levels of primary production throughout the year. Finally, the Coastal Boundary biome is composed of the continental shelves and the adjacent slopes, i.e. from the coastlines to the oceanographic front usually found at the shelf edge (Pauly et al., 2000). Next, the biomes are subdivided into 57 biogeochemical provinces (BGCPs), defined by satellite imagery and physical oceanography. Each BGCP is characterized by a distinct regime of physically driven water mixing, leading to a distinct pattern of (seasonal) supply of nutrients to the euphotic zone, and hence primary production (Longhurst et al., 1995; Longhurst, 1998; see also Table 6.1). In this scheme, the BGCPs comprising coastal biomes largely overlap with the area covered by the LME mentioned above, and hence the suggestion of a consensus system in Pauly et al. (2000), currently being implemented through a collaboration between members of various teams represented by the authors of the consensus statement. A further advantage of the consensus approach implied by the structure provided by LME/BGCP is that it leads to emphasizing benthic–pelagic coupling, as a single set of ecosystems is proposed for the neritic (shelf) areas of the world. This is appropriate, as it counters the misguided tendency to separate the pelagic and benthic realms, which leads to ecosystem representations that are exceedingly ‘open,’ and in which benthic–pelagic coupling must be represented explicitly (and thus quantified). Rather, benthic–pelagic coupling should be allowed to appear as an emergent property of neritic food webs, as will occur when one’s ecosystem representation includes predators feeding both on benthic and pelagic organisms, and detritus (e.g. marine snow) that is consumed both while sinking, and after it has sedimented. Conversely, there is no need for benthic–pelagic coupling in representations of open ocean systems, where the pelagic (sub)system is largely independent of benthic processes, and can be modelled as such; see, for example, Kitchell et al. (1999). Another reason to be wary of uncoupling the benthic and pelagic components of neritic
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Fig. 6.4. System of large marine ecosystems (LMEs) identified by K. Sherman and collaborators. This maps includes 12 recently defined LMEs (notably around Australia; see Table 6.1).
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Fig. 6.5. (a) The four ‘biomes’ in the global ocean stratification of A.R. Longhurst and colleagues (Polar, Westerlies, Trade-Winds and Coastal Boundary). Note their overall match with a global climate map (insert, from Anon., 1991). (b) Biogeochemical provinces (BGCPs) in the system of A.R. Longhurst and collaborators. Note that each BGCP fits into one of the four biomes in (a), thus allowing for a nested hierarchy of comparable ecosystems.
systems is that fishing itself tends to turn ecosystems dominated by benthic organisms (in terms of biomass or species numbers) into systems dominated by (small) pelagics and planktonic organisms. This feature, initially documented as a response to the stress generated by the combined effects of pollution and overfishing in the North American Great Lakes, has now been shown capable of being induced by fishing alone – at least in principle
(see Parsons, 1996). Broadly speaking, this would be due to trophic cascades, wherein fewer piscivores → more small pelagics → fewer zooplankton → more phytoplankton. Such indirect effects are very hard to identify in practice, given the contribution of terrigenous fertilizers in regions plagued by algal blooms, such as the northern Gulf of Mexico (Turner and Rabalais, 1994) or the inner Gulf of Thailand (Piyakarnchana,
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1999). However, the possibility that such effects can occur provides a good additional reason for constructing models of neritic systems that integrate the entire water column, and not only their benthic or pelagic components.
Construction of Catch Maps and Issues of Scale The scope of fisheries science, and of the related components of marine biology, traditionally has been defined by the scale of the fisheries studied (Pauly and Pitcher, 2000), which may range from a few square kilometres or even less (e.g. in the case of fisheries for sessile invertebrates) to thousands of square kilometres in the case of high sea fisheries. However, basin-level analyses are rare, except for tuna fisheries (for which, incidentally, mapping frequently is used; see below). Over 75% of fisheries landings (in value) are consumed in countries other than those owning the exclusive economic zone (EEZ) in which these landings were realized (based on FAO, 2000). In contrast, only 4–5% of the rice grown in the world is traded internationally
Table 6.1.
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(Maclean, 1997). This, by itself, provides a rather good reason why global fisheries maps are appropriate to our times – not to mention the need to quantify the global impacts of fisheries on marine ecosystems. One objection frequently heard relating to the feasibility of large-scale fisheries maps is the absence of suitable data, widely understood to consist of positive records of where some fishing unit may have caught, at a certain time, a certain quantity of fish (as plotted in tuna atlases) (Fonteneau, 1997; see Table 6.1 for FAO atlas). Such records, usually supplied by the industry, or costly observer programmes, are indeed rather scarce and, when available, are either presented at very coarse scales (e.g. in 5° scale for the FAO tuna atlas, to mask small-scale patterns with the high concentrations so dear to the industry) or pertain to small areas, and the catch of a limited set of gear. Constructing global fisheries maps from such data, i.e. from the ‘bottom-up’, does indeed seem unfeasible. However, such ‘positive’ records are not required to construct fisheries catch maps. These can also be constructed from the geographic range of the exploited taxa, and constraints on which parts of that range led to the reported catches, i.e. from ‘negative’ records as it were. The
Databases (on-line or CD-ROM) used for the construction of fisheries catch maps.
Data type
Organization
URL
Fisheries landings Tuna and billfish landings by 5° cells Taxonomy for all species, and ranges for many Distribution of commercial fish and invertebrates Physical ocean data (depths, temp. etc.) Primary productivity
FAO FAO
www.fao.org/fi/statist/FISOFT/FISHPLUS.asp www.fao.org/fi/atlas/tunabill/english/home.htm
FishBase
www.fishbase.org
FAO
www.fao.org/fi/sidp/default.htm
NOAA
www.ngdc.noaa.gov/mgg/global/global.htm
Coral reefs Seamounts Sea ice extent Exclusive economic zones Fishing agreements
www.me.sai.jrc.it/me-website/contents/shared_utilities/ frames/index_windows.htm www.reefbase.org ReefBase www.ngdc.noaa.gov/mgg/global/global.htm NOAA nsidc.org/index.html Univ. Colorado Veridan Information www.maritimeboundaries.com/main.htm Solutions (see FAO, 1998) Contact FAO FAO JRC of the EU
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procedure used for this is presented in Box 6.2 and Fig. 6.6a and b, while Table 6.1 indicates sources of information on the key databases used in the process. This approach works straightforwardly at larger scales (FAO areas, biomes, ocean basins or global), but not at smaller scales, wholly comprised within the geographic range of a number of resources species, where positive knowledge on fleet operations is required. Issues of scale also show up when dealing with the definition of an ecosystem. Indeed, such issues are implicit in its definition as an ‘area where a set of species interact in characteristic fashion, and generate among them biomass flows that are stronger than the flows linking that area to adjacent ones’ (Pauly and Froese, 2001). This definition applies to the large scale inherent in the LMEs and BGCPs presented above, but also the small scale (a few hundred metres), where organisms interact to form the food webs that characterize coral reefs, or small lagoon or estuarine systems.
Using maps to reach new audiences It is not the fishing industry that is asking for fisheries management to be put on an ecosystem basis, but politicians, pressed by conservation-orientated NGOs, themselves expressing public unease about the way marine ecosystems, and especially their more charismatic components (marine mammals, turtles and birds), are being affected by fishing. Thus, progress in putting fisheries on an ecosystem basis will have to be reported to that audience – quite a change for fisheries scientists accustomed to generate total allowable catches (TACs), communicated to highlevel bureaucrats by their superiors, fought by industry representatives, then applied or not to contain a fishery on the ground. It may be argued that the public at large will not understand the message conveyed by maps of fisheries and their ecosystem impacts. Yet, every day, a large part of the population, in most countries of the world,
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watch television weather programmes, and understand the sophisticated weather maps presented therein, although they are based on millions of data points analysed in quasi-real time by supercomputers, and combine dynamic displays of temperature, atmospheric pressure, cloud cover, risk of precipitation, etc. The public has been able to learn the ‘language’ of weather maps because: (i) it matters and (ii) visual displays presented in intuitive fashion can convey far more information than a text that is read or heard (Tufte, 1983). Thus, engaging the public and our political representatives in debates about the state of fisheries resources, and about alternative approaches to their utilization and long-term sustainability, should be possible, if we use the proper format for conveying that
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information. We believe that maps provide that format, and we have documented this with maps illustrating the decline of piscivorous fishes in the North Atlantic (see Box 6.3). It is our belief that without such engagement with the public, the fisheries sector, including the fisheries science that studies it, and the largely captive regulatory agencies that ‘manage’ the resources, will not be able to halt the decline illustrated by the biomass maps described in Box 6.3.
Maps and space-based fisheries management What is striking when examining the catch or biomass maps (Fig. 6.7) is that all show
Fig. 6.6. (a) (Opposite) Schematic representation of algorithm for construction of catch maps in the absence of positive, georeferenced catch records: the algorithm is initiated (in A) with global catch statistics from FAO, or from regional or national sources; its output is cells to which catches have been assigned (see also Fig. 6.5b and Box 6.2). (b) Schematic representation of algorithm for construction of catch maps in the absence of positive records. A catch record (from FAO database or other source) is evaluated via three criteria (what taxon, by which country, in which FAO area), and can be assigned only when one or more cells meet these criteria. Over 95% of the world catches from FAO can be assigned straightforwardly in this fashion; the remainder providing pointers to corrections of the assignment rules (see also (a) and Box 6.2).
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Box 6.3.
Examples of fisheries ‘weather’ maps for the North Atlantic (by V. Christensen).
The North Atlantic is one of the best studied marine areas of the world, which is not surprising since the marine sciences emerged largely along its shores, from 150–100 years ago. As a consequence, and contrary to the stubborn beliefs, by various colleagues, in a widespread ‘lack of data,’ abundant data sets exist which can be used to evaluate the impact of fisheries on the North Atlantic ecosystem. This is one of the goals of the of the Sea Around Us Project (SAUP; see www.fisheries.ubc.ca/projects/SAUP and Pauly and Pitcher, 2000). However, these data sets do not have the form required for analysis. (Note though, that this is always true: it is the analytical process itself which determines the format data should have.) After opting to present the SAUP analyses in form of ‘weather maps,’ a two-step approach was used for their construction: 1. Construction of mass-balance food web models for all major shelf areas, and a representative subset of oceanic areas to quantify biomasses at different trophic levels, and for different periods. 2. Extension, using a statistical model, of the biomass estimates in (1) to the entire North Atlantic, and the period from 1950 to 2000. Item (1) relied on 17 models constructed by a vast number of authors, most associated with national research institutions in the countries concerned (for details, see contributions in Guénette et al., 2001). Importantly, these models included data-rich representations of the North Sea in 1880, and the Newfoundland shelf in 1900, both implying higher biomass of predatory fishes than in the corresponding 1980s models, and several other model pairs, contrasting present biomass with those in the 1970s or 1960s. The fish biomasses in these models, all constructed using the ECOPATH WITH ECOSIM software (EwE), were put on a spatial basis using the Ecospace routine of EwE (Walters et al., 1999), using the same half-degree spatial cells also used to construct catch maps (see Box 6.2). Item (2) then consisted of identifying a general linear model of the form Biomasstyc = f (catchyc; year and physical attribute of half-cell) wherein the subscripts are t = trophic level, y = year and c = catch in each half-degree cell (mapped as presented in Box 6.2), and where the (bio-)physical attribute (assumed invariable in time) of each cell include mean depth and temperature, primary production, ice cover and other properties (see Christensen et al., 2001, for details). Examples of the maps thus generated, which indicate a strong decline of predatory fish biomass from 1950 to the present, indicative of ‘fishing down marine food webs’ (Pauly et al., 1998, 2001) are available on-line (www.fisheries.ubc.ca/projects/SAUP). These maps highlight processes that are rather worrisome, and none of the persons (both specialists and laypersons) to whom they have been shown has failed to perceive their analogy to weather maps, and to very bad weather developing over the North Atlantic.
the same trend, at least for the North Atlantic, the only ocean basin we have examined so far in some detail. This is not surprising, given that we did not include local components in the algorithms used to generate these maps. The point, though, is that at the scale we were working (with ~ 21,000 pixels of half a degree latitude and longitude), there were no marine reserve, or other refugia (with biomass trends different from the general downward trend for the North Atlantic as a whole) to consider. Put differently: there are – at the scale of our pixels, appropriate to represent the
distribution ranges of all but small intertidal species – no areas of the North Atlantic where fishes are not exposed to nets and other gear designed for, and extremely effective at, catching them. The deleterious effects of the sort of default setting implied here (i.e. that fish can be exploited anywhere, unless regulations state otherwise, rather than the converse) are discussed by Walters (1998). Put as a map, contrasting areas with fishing (say red) versus areas without any fishing, this would imply a single colour for the entire North Atlantic, without any green, or other shades (Fig. 6.8). We are
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Fig. 6.7. Catch map for the North Atlantic (here: 1990s), constructed as explained in Box 6.2 and Fig. 6.6a and b, with darker shades indicating higher catch rates, in t km−2 year−1. Colour versions of this and similar maps for other periods and areas are available on-line; see Box 6.3.
Fig. 6.8. Map of the North Atlantic, with dark identifying those areas where fishing is allowed, and fish killed, and light the areas where no fishing is permitted, thus allowing the resource to recover. Unfortunately, there are no light areas to be seen at the scale of half-degree cells. (The authors welcome corrections that would identify recently created refugia.)
confident that, as for the weather maps presented above, the meaning of this monochromatic map will be widely understood by lay audiences. Figure 6.8 would be less dire had it been designed to illustrate the extent of protection for coral reefs and other sensitive coastal
systems (e.g. kelp beds). For these, the idea of area-based protection is well accepted, and a number of (mainly small) marine reserves have been created. Here, high-resolution maps are understood as playing the key role in defining the terms of the debate between different
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stakeholder groups, namely local residents, small-scale fishers, dive-resort operators, environmental NGOs, etc. In contrast, it may take a while for the management of commercial fisheries to rely on area-based protection as its key tool. When this happens, maps will be there to help us see where we are, and where we should be going.
Acknowledgements We thank Mr A. Gelchu for shape files of fish distribution, Dr U.R. Sumaila for estimating the fraction of the world’s fish catch value that is exported, and our colleagues at FAO for fruitful exchanges. We also thank the Pew Charitable Trusts for their support of the Sea Around Us Project.
References Anon. (1991) Bartolomew Illustrated World Atlas. Harper Collin, Edinburgh. Chatfield, C. (1984) The Analysis of Time Series: an Introduction. Chapman and Hall, London. Christensen, V., Guénette, S., Heymans, S., Walters, C.J., Watson, R., Zeller, D. and Pauly, D. (2001) Estimating fish abundance of the North Atlantic, 1950 to 2000. In: Guénette, S. Christensen, V. and Pauly, D. (eds) Fisheries Impacts on North Atlantic Ecosystems: Models and Analyses. Fisheries Centre Research Reports 9(4), 1–25. www.saup.fisheries.ubc.ca/report/ report.htm Ekman, S. (1967) Zoogeography of the Sea. Sidgwick and Jackson, London. FAO (1998) FAO’s fisheries agreements register (FARISIS). Committee on Fisheries, 23rd Session, Rome, Italy, 15–19 February 1999 (COFI/99/Inf.9 E). FAO (2000) Fisheries trade flow (1995–1997). FAO Fisheries Circular, No. 961. Fonteneau, A. (1997) Atlas of Tropical Tuna Fisheries. Edition ORSTOM, Paris. Froese, R. and Pauly, D. (eds) FishBase 2000: Concepts, Design and Data Sources. Los Baños, Philippines. Garstang, W. (1909) The distribution of the plaice in the North Sea, Skagerrak and Kattegat, according to size, age and frequency. Report of the trawling investigations of the research
streamers from October 1902 to July 1907. Rapports et Procès-Verbaux des Réunions du Counseil Permanent International pour l’Exploration de la Mer 11, 136–138. Hilborn, R. and Walters, C.J. (1992) Quantitative Fisheries Stock Assessment: Choice, Dynamics and Uncertainty. Chapman and Hall, New York. Kitchell, J.F., Boggs, C.H., He Xi and Walters, C.J. (1999) Keystone predators in the Central Pacific. In: Ecosystem Approaches for Fisheries Management. Alaska Sea Grant College Program, AK-SG-9901, pp. 665–684. Longhurst, A.R. (1995) Seasonal cycles of pelagic production and consumption. Progress in Oceanography 36, 77–167. Longhurst, A.R. (1998) Ecological Geography of the Sea. Academic Press, San Diego. Longhurst, A.R., Sathyendranath, S.A., Platt, T. and Caverhill, C.M. (1995) An estimate of global primary production in the ocean from satellite radiometer data. Journal of Plankton Research 17, 1245–1271. Maclean, J. (ed.) Rice Almanac. IRRI, Los Baños, Philippines. Myers, R.A., Bowen, K.G. and Barrowman, N.J. (1999) The maximum reproductive rate of fish at low population sizes. Canadian Journal of Fisheries and Aquatic Sciences 56, 2404–2419. NRC (National Research Council) (1999) Sustaining Marine Fisheries. National Academy Press, Washington, DC. Parsons, T.R. (1996) The impact of industrial fisheries on the trophic structure of marine ecosystems In: Polis, G.A. and Winnemiller, K.D. (eds) Food Webs: Integration of Patterns and Dynamics. Chapman and Hall, New York, pp. 352–357. Pauly, D. (1999) Review of A. Longhurst’s ‘Ecological Geography of the Sea’. TREE 14(2), 118. Pauly, D. and Froese, R. (2001) Fish stocks. In: Levin, S. (ed.) Encyclopedia of Biodiversity, Vol. 2. Academic Press, San Diego, pp. 801–814. Pauly, D. and Pitcher, T.J. (2000) Assessment and mitigation of fisheries impacts on marine ecosystems: a multidisciplinary approach for basin-scale inferences, applied to the North Atlantic. In: Pauly, D. and Pitcher, T.J. (eds) Methods for Evaluating the Impacts of Fisheries on North Atlantic Ecosystems. Fisheries Centre Research Reports 8(2), pp. 1–12. www.saup.fisheries.ubc.ca/report/report. htm Pauly, D., Christensen, V., Dalsgaard, J., Froese, R. and Torres, F.C., Jr. (1998) Fishing down marine food webs. Science 279, 860–863.
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Pauly, D., Christensen, V., Froese, R., Longhurst, A., Platt, T., Sathyendranath, S., Sherman, K. and Watson, R. (2000) Mapping fisheries onto marine ecosystems: a proposal for a consensus approach for regional, oceanic and global integration. In: Pauly, D. and Pitcher, T.J. (eds) Methods for Evaluating the Impacts of Fisheries on North Atlantic Ecosystems. Fisheries Centre Research Reports 8(2), pp. 13–22. www.saup.fisheries.ubc.ca/report/report. htm Pauly, D., Palomares, M.L., Froese, R., Sa-a, P., Vakily, M., Preikshot, D. and Wallace, S. (2001) Fishing down Canadian aquatic food webs. Canadian Journal of Fisheries and Aquatic Sciences 58, 51–62. Piyakarnchana, T. (1999) Changing state and health of the Gulf of Thailand large marine ecosystem. In: Sherman, K. and Qisheng Tang (eds) Large Marine Ecosystems: Assessment Sustainability and Management. Blackwell Science, Malden, UK, pp. 240–250. Platt, T. and Sathyendranath, S.A. (1988) Oceanic primary production: estimation by remote sensing at local and regional scales. Science 241, 1613–1620. Samb, B. and Pauly, D. (2000) On ‘variability’ as a sampling artefact: the case of Sardinella in north-western Africa. Fish and Fisheries 1, 206–210. Sathyendranath, S.A., Longhurst, A.R., Caverhill, C.M. and Platt, T. (1995) Regionally and seasonally differentiated primary production
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in the North Atlantic. Deep-Sea Research 42, 1773–1802. Sherman, K. and Duda, A.M. (1999) An ecosystem approach to global assessment and management of coastal waters. Marine Ecology Progress Series 190, 271–287. Tufte, E.R. (1983) The Visual Display of Quantitative Information. Graphic Press, Cheshire, Connecticut. Turner, R.E. and Rabalais, N.N. (1994) Coastal eutrophication near the Mississippi river delta. Nature 368, 619–621. Walters, C.J. (1998) Designing fisheries management systems that do not depend upon accurate stock assessments. In: Pitcher, T.J., Pauly, D. and Hart, P. (eds) Reinventing Fisheries Management. Fish and Fisheries 23. Kluwer Academic, Dordecht, The Netherlands, pp. 279–288. Walters, C.J., Christensen, V. and Pauly, D. (1997) Structuring dynamic models of exploited ecosystems from trophic mass-balance assessments. Reviews in Fish Biology and Fisheries 7(2), 139–172. Walters, C.J., Pauly, D. and Christensen, V. (1999) ECOSPACE: prediction of mesoscale spatial patterns in trophic relationships of exploited ecosystems, with emphasis on the impacts of marine protected areas. Ecosystems 2, 539–554. Watson, R., Pang, L. and Pauly, D. (2001) The marine fisheries of China: development and reported catches. Fisheries Centre Research Reports 8(2). www.saup.fisheries.ubc.ca/ report/report.htm
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The Functioning of Marine Ecosystems: a Fisheries Perspective Philippe Cury1, Lynne Shannon2 and Yunne-Jai Shin3 1IRD
Oceanography Department, University of Cape Town, Rondebosch, South Africa; 2Marine and Coastal Management, Cape Town, South Africa; 3IRD Centre de Recherche Halieutique Méditerranéenne et Tropicale, Sète Cedex, France
Not only is the science incomplete, but the [eco]system itself is a moving target, evolving because of the impact of management and the progressive expansion of the scale of human influences on the planet (C.S. Holling, 1995)
Abstract There is considerable evidence that environmental variability plays a major role in controlling abundance and distribution of marine populations and that fisheries alter ecosystem functioning and state. This overview documents emergent, i.e. visible to us as observers, ecosystem-level ecological patterns and addresses important questions regarding the exploitation of marine resources. Do marine ecosystems function differently from terrestrial systems? Do multiple stable states exist in marine ecosystems? Does removal of top predators in marine ecosystems result in fundamental changes in the plankton communities (top-down ‘trophic cascades’), as observed in lakes? Alternatively, are marine ecosystems characterized by bottom-up control such that fishing predatory fish does not disturb community structure and function? Does heavy exploitation of forage species, such as anchovies and sardines, cause changes in the functioning of upwelling ecosystems? The key to answering these questions and exploring whether general principles apply lies in understanding the energy flow within the ecosystems. The chapter reviews different types of energy flow in marine ecosystems, i.e. bottom-up control (control by primary producers), top-down control (control by predators) and wasp-waist control (control by numerically dominant species). No general theory can yet be ascribed to the functioning of marine ecosystems. Ecological understanding and models of ecosystem functioning are provisional and subject to change, and common sense is not sufficient when studying complex dynamic systems. However, tentative and partial generalizations are proposed, namely that bottom-up control predominates; top-down control plays a role in dampening ecosystem-level fluctuations; trophic cascades seldom occur; and wasp-waist control is most probable in upwelling systems. Moreover, alternation and large-scale synchronized fluctuations in fish stocks, stability of fish communities and emergent features such as size spectra are potentially important patterns when assessing states and changes in marine ecosystems. New and meaningful indicators, derived from our current understanding of marine ecosystem functioning, can be used to assess the impact of fisheries and to promote responsible fisheries in marine ecosystems.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Introduction Although the term ‘ecosystem’ is quite recent (Tansley, 1935), it is now part of the mainstream of ecological science. An ecosystem is defined as ‘a spatially explicit unit of earth that includes all of the organisms, along with all the components of the abiotic environment within its boundaries’ (Likens, 1992). This definition remains vague as ecosystems have no apparent boundaries and lack the sort of clear objective or purpose that can be ascribed to other, more tractable, biological or ecological entities (e.g. cell, individual or population). A marine ecosystem contains water, detritus and hundreds of kinds of organisms, including bacteria, phytoplankton, zooplankton, fish, mammals and birds. All these components are connected in a complex food web by evolving interactions (Fig. 7.1). Until recently, fisheries management has been based largely on single-species approaches (Beverton, 1984). However, ecosystem-based management represents a paradigm shift, as well as a new attitude towards the exploitation of renewable marine resources (Christensen et al., 1996). The ecosystem is now viewed as an integrative level for ecological studies, and its overall complexity is perceived as critical to its sustainability (Costanza et al., 1997). Ecosystems comprise a diverse array of processes that provide both goods and services to humans. It also becomes important to understand what impacts an ecosystem can tolerate before major structural changes occur, and how reversible these changes are. In this respect, improved understanding of ecosystem dynamics is critical to predict and manage the consequences of environmental variability and human impacts, such as those induced by marine fisheries, an activity targeting specific species and size classes. This overview aims to address questions regarding the exploitation of marine resources, such as:
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Do marine ecosystems function differently from terrestrial systems? Are there multiple stable states in marine ecosystems?
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Which species are most critical, and which ecological processes are most sensitive, to exploitation? Does the removal of top predators have a strong impact on lower trophic levels? Does the removal of large portions of forage fish species, such as anchovies and sardines, result in changes in the functioning of upwelling systems?
The key to answering these questions and finding out whether general principles apply lies in understanding the mechanisms responsible for the observed emergent ecological patterns. Marine ecosystem functioning depends on its structure, diversity and integrity. Alteration or disturbance of one or several components of marine ecosystems can have strong effects on higher or lower trophic levels, depending on whether food webs are controlled by resources or by predators. This chapter reviews different types of energy flow in ecosystems and how they possibly influence the dynamics of marine communities. They are presented in a didactic way in order to structure in an organized manner the different theories developed for linking trophic levels. For this purpose, the chapter presents recent theoretical ecological knowledge, illustrates it with case studies, and explores whether simple questions, such as those posed above, can have simple answers.
Bottom-up Control: the Very Small Drive the Very Large Victor Hensen is considered to have been the father of quantitative marine ecology. In 1887, he thought of planktonic populations as rapidly revolving links in a food chain leading from the very small to the very large (Smetacek, 1999). Using an analogy with agronomy, where crop yields can be predicted from the control of the input, Hensen made the assumption that food supply regulates adult fish stocks, and quantitative studies of phytoplankton and zooplankton production would permit predictions of fish yields (Verity, 1998). From this deduction was born the notion that ecosystems were ‘bottom-up’
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Fig. 7.1. Food webs in the northern and southern Benguela showing the complexity of the interactions between the different components of the ecosystems. Important differences in the functioning of the ecosystem have been observed between two close and fairly similar upwelling systems (adapted from Shannon and Jarre-Teichmann, 1999).
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controlled (Fig. 7.2a). In other words, the regulation of food-web components derives from either primary producers, or the input of limited nutrients (Pace et al., 1999). Recently, Micheli (1999) analysed 20 natural marine systems, and found that nutrients generally enhance phytoplankton biomass. Plants dominate terrestrial ecosystems but the ocean contains less than 1% of plant biomass (Smetacek, 1999). The realization that the ocean’s animals are fed by a thin soup of minute algae, the phytoplankton, and that this resource can limit the global productivity of marine ecosystems, was puzzling. Environmental forcing controls the carrying capacity and fish biomasses in marine ecosystems, but apparently not in a simple way. It took time to realize that the marine environment is a dispersive and heterogeneous one. Species
are not evenly distributed spatially, and their marine populations, particularly of fish, fluctuate widely from year to year. Since the pioneering work of Hjort in 1914, which is still influential today, it was recognized that renewable processes in fish population dynamics are highly irregular, depending on recruitment strength, and that marine fish species comprise many selfsustaining populations (Sinclair, 1997). Currently, there is considerable evidence that natural variability in ocean circulation and mixing plays a major role in generating fluctuations in marine productivity, as well as in the distribution of populations. Food availability and physical constraints such as retention, concentration or enrichment processes that are associated with currents and turbulence are now considered to be
Fig. 7.2. (a) Bottom-up control within a simplified four-level food web in a marine ecosystem. (b) The physical environment, being less favourable, controls the decrease in abundance of the phytoplankton, which in turn has a negative impact on the abundance of the zooplankton. The decline in zooplankton abundance controls the decrease in abundance of the prey fish, which itself leads to a decrease in the abundance of the predators (the control factor is a dashed line and the responses are solid lines).
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important factors that affect larval survival, fish recruitment and, ultimately, stock abundance (Cury and Roy, 1989; Bakun, 1996; Chambers and Trippel, 1997). Fish populations have geographical closure of their life cycles, and climatic factors can affect the spatial context of marine populations in many ways by modifying the dynamics of the spawning or feeding areas, consequently changing recruitment success of migrational patterns (Sinclair, 1988). Several recurrent patterns can illustrate how the physical environment plays a structuring role in shaping abundance and distribution of marine populations in space and time.
Ecosystem responses to drastic environmental changes The structure and function of marine ecosystems respond drastically to inter-annual changes and inter-decadal climatic variations. This has been documented for the California Current, the Gulf of Alaska (McGowan et al., 1998), the North Atlantic (Aebischer et al., 1990) and off Chile (Hayward, 1997). Parallel long-term trends across four marine trophic levels, ranging from phytoplankton, through zooplankton and herring, to marine birds, have been related to environmental changes in the North Sea (Aebischer et al., 1990). Even though the mechanisms behind the parallelism of trends remain unclear, the effect of the environment was identified as the driving force for structuring several components of the ecosystem. Using trophic mass-balance models, the multiple and complex changes that occurred in the Bering Sea ecosystem between the 1950s and 1980s were shown to be largely driven by environmental changes (Trites et al., 1999). Inter-annual environmental fluctuations, such as El Niño events, affected the structure of the plankton community, the spatial distribution of fish and invertebrates, the recruitment success of pelagic fish and the mortality of birds and mammals in the northern Pacific (McGowan et al., 1998). Large-scale perturbations have taken place during the past 20 years in the Pacific, where a dramatic shift of the
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atmospheric forcing occurred in the mid-1970s (Hayward, 1997). Inter-decadal regime shifts, such as the one experienced in the entire North Pacific Basin and the California Current from the late 1970s to the early 1980s, appear to have altered the productivity of marine ecosystems, at various trophic levels (Polovina et al., 1994; Francis et al., 1998). There has been a generally increased frequency of southern species moving north, a substantial lowering of secondary productivity and fish landings, a major decline in seabirds, and changes in species composition in most sectors of these ecosystems (McGowan et al., 1998). However, the biological response to the inter-decadal regime shift in the Gulf of Alaska is thought to have been in the opposite direction to that of the California Current. It seems that in the ocean there are large-scale biological responses to low-frequency climatic variations. However, the mechanisms by which climate exerts its influence vary as components of the ecosystem are constrained by different limiting environmental factors. Thus similar species at the same trophic level may respond quite differently to climate change (Hayward, 1997). Findings from one system cannot necessarily be extrapolated to others, and predicting the effects of global-scale environmental change on ecosystems does not appear to be a straightforward exercise.
Regime shifts and synchronized large-scale fluctuations Changes in the abundance of pelagic fish species have been recorded in many marine ecosystems, based on catch statistics (e.g. Schwartzlose et al., 1999), biomass surveys (e.g. Hampton, 1992) and records of seabird guano harvests (e.g. Crawford and Jahncke, 1999). It was hotly debated whether collapses of pelagic fish stocks were caused by overfishing, which allowed competing species to dominate (Francis and Hare, 1994). However, records of scale deposition from anaerobic sediments show that large-amplitude fluctuations of pelagic fish stocks (e.g. sardine or anchovy) occurred even in the absence of any
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fishery over a period of 2000 years (Soutar and Isaacs, 1974). Sediment records of δ15N and biological indicators were used to reconstruct the abundance of Pacific salmon over the past 300 years (Finney et al., 2000). Marked shifts in sockeye salmon populations occurred over decades during this period, and some pronounced changes appear to be related to climatic change. These regime shifts can alter the nutrient cycles and may have significant impacts on the productivity of the ecosystems. Since the beginning of modern fisheries, the emerging patterns of decadal-scale variation in pelagic fish populations have also exhibited a substantial degree of global synchrony, sometimes between remote areas (Schwartzlose et al., 1999) (Fig. 7.3a). This
synchrony most probably is driven by global climatic teleconnections (Bakun, 1996; Klyashtorin, 1997). Drastic change of state in one abundant prey resource is expected to have major consequences for the functioning of the ecosystem. Small pelagic fish are forage fish in marine systems; they represent an important source of food for numerous top predators, such as large pelagic fish, demersal fish, marine birds and mammals (Anon., 1997). The collapse of a prey species, induced by climate or/and fisheries, most often is associated with massive mortality of mammals, birds and predatory fish (Cury et al., 2000). However, the collapse of an abundant forage fish stock can also have an impact on other species at the same trophic level.
Fig. 7.3. Global patterns of decadal abundance, as illustrated by pelagic fish catch (in million t) in several ecosystems. (a) Decadal-scale regime shifts suggest the existence of multiple stable states in pelagic fish assemblages. The synchrony of fish populations between remote ecosystems suggests strong climatic connection. (b) The alternation between different pelagic fish (full line = sardine; dotted line = anchovy) suggests replacement between redundant species. (Adapted from Schwartzlose et al., 1999; and Bakun, 1998.)
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Alternating steady states and pelagic fish assemblages Strong environmental effects on fish populations result in large fluctuations in species composition. It also appears that alternating steady states are observed at the level of fish assemblages on decadal scales. For example, upwelling systems tend to be dominated by one species of sardine (or sardinella) and one species of anchovy, but most often only one of the two is dominant at any particular time. Alternating patterns between small pelagic fish species have been observed in most upwelling ecosystems over past decades (Fig. 7.3b). A regime shift between two species is considered to operate when, after following the decrease in the biomass of one species, for example, due to removal by environmental effects or fishing, the total biomass is restored by density compensation of the other species. This process may occur between two redundant species, i.e. belonging to the same guild or functional group (Lawton and Brown, 1993). The mechanisms that generally are invoked in direct competition are not completely satisfactory to explain species alternation (Hall, 1999) because sardine and anchovy usually do not have the same spatial distribution (sardine are usually found farther offshore) and do not eat the same type of food (e.g. off South Africa, anchovy preferentially feed on large zooplankton whereas sardine prefer phytoplankton and small zooplankton (Van der Lingen, 1994)). These arguments led several authors to consider that competition is magnified by schooling behaviour within mixed-species schools (Bakun and Cury, 1999). Analysing changes in abundance of pelagic species in response to environmental changes, Skud (1982) concluded that the dominant species responds to environmental factors, while the subordinate species responds to the abundance of the dominant one. From an ecosystem perspective, climatic factors are thought to affect fluctuations in abundance of a species, whereas its absolute density is controlled by intraspecific competition (Skud, 1982; Serra et al., 1998). Under bottom-up control, the physical environment drastically
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affects the overall productivity (i.e. the carrying capacity) of ecosystems, but, more importantly, also the dynamics of fish assemblages in a more or less predictable way. For pelagic ecosystems, recurrent patterns in fish abundance can be expected. These decadal-scale patterns of alternation are important for longterm management, as exploitation reduces the biomass of the dominant species, which is usually the target species, and sometimes precipitates its collapse (Beverton, 1990). Decadal-scale regime shifts suggest the existence of multiple stable states in pelagic communities, resulting in sustained or unsustained pelagic fisheries. Within a pelagic community, alternation also suggests that harvesting a prey species will favour a competing species, provided that the latter is only lightly exploited. However, in the northern Benguela ecosystem off Namibia, commercially valuable sardine began to show signs of collapse in the 1970s. The fisheries targeted anchovy heavily, with the view that reducing anchovy would benefit its competitor, i.e. sardine (Butterworth, 1983). The attempt to enhance sardine abundance failed, and both anchovy and sardine underwent major declines in the late 1970s. In comparison, off South Africa, anchovy were managed conservatively when sardine collapsed in the late 1960s, allowing anchovy to reach large biomasses and support a large fishery during the 1980s. Alternation of pelagic fish species is a rule that has exceptions. What has relatively few exceptions is the observation that, when not replaced by another species, the collapse of a dominant prey species due to exploitation or other natural causes alters the abundance and distribution of predator communities (see also the section below on wasp-waist control). The conceptual model based on food limitation and responses to increased resource availability by elevated standing stocks is regarded as the paradigm (Hunter and Price, 1992). Bottom-up control offers a comprehensive framework for understanding how different components would react to environmental changes or to changes at the bottom of the food chain (Fig. 7.2b). However, it appears that certain taxa are better than others at regulating the flux of materials through the food web, and that predation is sometimes as
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important as resource limitation (Verity, 1998).
Top-Down Control: the Very Large Drive the Very Small As species mostly interact through predation, the existence of top-down control, which means the regulation of lower food-web components by one or several upper-level predators, should be critical in the functioning of
marine ecosystems (Fig. 7.4a). Predation mortality is estimated to be the major source of mortality for marine exploited species. An analysis of six marine ecosystems (Benguela Current, Georges Bank, Balsfjord, East Bering Sea, North Sea and Barents Sea) suggests that predation represents between two and 35 times fishing mortality (Bax, 1991). This does not mean that fishing has negligible effects on species dynamics, but rather implies that it can affect the whole ecosystem, as species are tightly connected through the process of predation.
Fig. 7.4. (a) Top-down control within a simplified four-level food web in a marine ecosystem. (b) The decreasing size of the top predator populations leads to reduced predation on the prey, which in turn leads to an increase in abundance of the prey fish. Increased predation of fish prey on zooplankton leads to a decrease in the zooplankton population size. The smaller zooplankton abundance reduces the grazing pressure on the phytoplankton, which consequently becomes more abundant. The control factor is represented by a dashed line, and the responses are solid lines.
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When bigger fish eat smaller fish In terrestrial and aquatic ecosystems, the range of potential prey for a given species depends largely on their morphometric characteristics. For different taxonomic groups, the mean size of prey increases with predator size. This is the case for aquatic species, for which body size is considered to be the main constraint in the predator’s ability to catch its prey (Lundvall et al., 1999). According to Sheldon et al. (1977), size-based predation in the aquatic environment is supported by the fact that fish live in a medium that is 800 times denser than air, where only a streamlined morphology facilitates active and efficient movements. In this context, the development of appendages, which would help to handle and capture large sized prey, is not common among fishes. Thus, a predatory fish must have a jaw large enough to swallow its prey whole. A lion can catch prey bigger than itself, but a fish generally cannot. As the size of the jaw is related to fish size, predation is believed to be determined largely by the size ratio between predator and prey (Fig. 7.5a). As suggested by Ursin (1973) ‘[fish] stomach contents are a simple function of local prey availability and suitability, this latter often simply being a function of size’. Feeding in marine food webs can be considered as opportunistic and less dependent on prey taxonomy than on prey size. Strong patterns emerge from this feeding behaviour (Fig. 7.5b). First, fish diets comprise large prey diversity. Generally, fish larvae feed at the base of the food web and, when they become adults, they occupy higher trophic levels and feed at one or several trophic levels below their own (Rice, 1995). Fish species often tend to have multiple predators and multiple prey. Secondly, cannibalism is also common in fish communities and can represent an important source of pre-recruit mortality (Claessen et al., 2000). For instance, for the Eastern Baltic stock of cod (Gadus morhua), between age 0 and age 2, a year class may lose about 31–44% of its initial individuals as a result of cannibalism (Neuenfeldt and Köster, 2000). Finally, eggs and larvae are all located at the base of
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piscivorous trophic levels (Jones, 1982). Particular to all species of teleosts is the rather homogeneous size of their eggs, about 1 mm in diameter (Cury and Pauly, 2000). The first consequence is that pre-recruits are subjected to what might be called ‘community predation’ (Sissenwine, 1984), with every fish species potentially competing with every other (Fig. 7.5b). As stated by Gulland (1982), ‘fish have no direct terrestrial counterparts – a fox or lion does not start competing with mice’. The second consequence is that two species can be simultaneously a predator or a prey of each other, according to the stage in their life cycle (i.e. their size). For instance, North Sea cod is known to be a predator of herring, but it is also its prey, since adult herrings feed on cod larvae (Stokes, 1992). This suggests two competing top-down control mechanisms, on a species basis, but one unidirectional top-down control type on the basis of size (Fig. 7.5b). Thus, considering the number of potential interactions between the different species, trophic levels or age groups, marine food webs appear to have complex and evolving dynamics. However, patterns of trophic interactions have been shown to exhibit strong emergent properties at the level of the ecosystem. As stated by May (1974), ‘if we concentrate on any one particular species our impression will be one of flux and hazard, but if we concentrate on total community properties [. . .] our impression will be of pattern and steadiness’. A recurrent observation is the relative stability of the total fish biomass compared with that of individual species in marine ecosystems (Fig. 7.5c). For example, in the North Sea during the 1970s, fisheries experienced important variations in the species composition of catch, but the total catch remained relatively stable: herring and mackerel catches collapsed while those of gadoids increased (May et al., 1979), even though a global decreasing trend for total fish biomass has been observed over the past five decades (Pauly, et al., Chapter 6). May et al. (1979) assumed that year-class strength is regulated by top-down control: fishing reduced the biomass of mackerel and herring, which in turn reduced predation pressure on the larvae of
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Fig. 7.5. (a) Bigger fish eat smaller fish: fish prefer prey smaller than about one-quarter to one-third their own size, as predators are constrained by the size of their jaw. (b) Who is eating whom? This simple opportunistic feeding behaviour generates complex trophic webs, wherein fish have multiple predators, multiple prey and multiple competitors. A fish can feed on different trophic levels (omnivory), on its own progeny (cannibalism) and on early life stages of its predators (e.g. eggs and larvae). Three species are represented on the vertical axis, and four size classes on the horizontal axis. Along the axes, the thin arrows correspond to the potential predation interactions between species and size classes. Cannibalism is represented by loops along the vertical axis. Within this framework, the arrows relating fish correspond to a theoretical example of a trophic web. (c) Complexity–stability: a recurrent pattern is the relative stability of the total fish biomass compared with that of individual species. Size-based predation implies multiple and weak trophic interactions between species, which have been proved theoretically to favour stability. (d) Size-based predation provides an explanation for observed size spectra in marine ecosystems. A remarkably linear relationship is obtained when the logarithm of the numbers of fish in a size class is plotted versus the logarithm of the median size of the size class.
gadoids, and consequently improved recruitment. In the context of size-based predation, fish can be considered as general predators that may represent stabilizing forces on
populations because they eat a variety of prey and target the most abundant species (Bax, 1998). Therefore, top-down control may operate through multiple and weak trophic
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interactions between species, a case which has been proved theoretically to favour stability (McCann, 2000; Shin and Cury, 2001). The relative stability of the size spectrum is also a recurrent feature at the level of the community (Fig. 7.5d). Contrasting with changes in species composition, the size spectra of marine ecosystems exhibit remarkably constant shapes. Fish number or biomass are decreasing functions of fish size, which are linear or dome-shaped depending on the metrics used (e.g. Bianchi et al., 2000). This observation suggests that, beyond strict species interactions, size-based interaction controls energy transfer in the marine environment. In this regard, primary production (bottom-up control) may act as a scaling factor that determines global productivity of the ecosystem, but the stabilization process may be under top-down control. In this context, fishing acts as an apex predator, targeting the largest size classes. It has been shown that this top-down effect can be assessed by the variations in the slopes and intercepts of ecosystem size spectra. In a comparison between North Sea and Faeroe Bank ecosystems, Pope and Knights (1982) showed that heavier exploitation in the North Sea led to a steeper slope of the observed size spectrum.
All species are not equal As everything is not strongly connected to everything else, there is no need to measure or understand everything, but rather to determine the significant interactions. Once this idea was accepted, it was rapidly recognized that certain key species play a more important role than others in structuring ecosystems.
Keystone species The most widely used definition for keystone species is one ‘whose impact on its community or ecosystem is large, and disproportionately large relative to its abundance’ (Power et al., 1996), i.e. keystone species affect processes at the community or ecosystem level to a greater extent than would be
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expected based upon their relative abundance alone (Bond, 1993). Keystone species are likely to occur near the top of the food chain, although they are not necessarily at the highest trophic levels (Power et al., 1996) and they are, by definition, not abundant. They have an impact on other species through consumption, competition, etc., and also by physically modifying habitat characteristics (ecosystem engineering). Ecologists have devoted a lot of attention to identifying keystone species in nature, as it has been suggested that the future of conservation management might lie in maintaining keystone species rather than attempting to protect and manage all species subjectively considered to be important or vulnerable (Power et al., 1996). Paine (1966) examined the top-down effects of predatory starfish in rocky intertidal communities, showing for the first time that some species play key roles in ecosystems by preventing single species from monopolizing a limited resource. In the presence of the starfish Pisaster ochraceus, the intertidal community at Mukkaw Bay, Washington, comprised a diverse assemblage of algae, mussels, barnacles, chitons, limpets, sponges and nudibranchs (Fig. 7.6a). When the keystone predator, P. ochraceus, was removed experimentally, its most important prey species, the mussel, Mytilus californianus, was able to proliferate, reducing species diversity and effectively reverting the ecosystem to a monoculture of mussels (Fig. 7.6b). Sanford (1999) found that the strength of the Pisaster–Mytilus interaction is reduced during periods of cool upwelling, concluding that a change in keystone interaction strength as a result of environmental changes could have large impacts on ecosystems. There are very few examples of marine keystones that are not from the intertidal region. One example is the top-down control by the jellyfish Aurelia aurita, shown to determine the structure of the zooplankton community in a shallow cove in Denmark (Oleson, 1995). Species newly introduced to an ecosystem (e.g. invasive aliens) may have strong effects that are disproportionately large relative to their biomass, i.e. they may be considered to be keystones even if they are
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Fig. 7.6. Schematic representation of the keystone role of predatory starfish Pisaster ochraceus in an intertidal ecosystem in Washington (based on Paine, 1966). (a) Pisaster sp. predation maintains a diverse community. (b) Removal of Pisaster ochraceus allows mussels to dominate, and reduces species diversity.
not formally components of the ecosystems. Nevertheless, these species subsequently may become dominant as they benefit from the absence of predators and diseases in their new environment (Power et al., 1996). Kitchell et al. (1999) examined the possible keystone effects of apex predators, such as sharks, tunas and billfishes, in the Central North Pacific, finding that no single species at high trophic levels could be considered as a true keystone. In their simulations using an ECOSIM model, Kitchell et al. (1999) found no evidence that strong predatory effects were propagated through the system, affecting species at lower trophic levels. Even removal of fisheries, shown to act in a similar way to a keystone predator, did not have effects on the first two trophic levels of the ecosystem. Although keystone species are not identified frequently in marine ecosystems, in some cases they can cause, in addition to other
changes in dominant species, major changes to ecosystem structure and functioning through trophic cascades down the marine food web.
Trophic cascades Trophic cascades are defined as reciprocal predator–prey effects that alter the abundance, biomass or productivity of a population or trophic level across more than one link in a food web (Pace et al., 1999) (Fig. 7.7, left). True trophic cascades imply keystone species (Paine, 1980), taxa with such top-down dominance that their removal causes precipitous change in the system. They result in inverse patterns in abundance or biomass across trophic links in a food web. These trophic interactions were first described in lakes (for a review see Carpenter and Kitchell, 1993) and intertidal zones
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Fig. 7.7. Trophic cascade illustrated in the Aleutian archipelago, western Alaska. On the left it is shown how the ecosystem was organized when the sea otters are abundant: they prey heavily on the sea urchin biomass, which remains low; the resulting weak grazing intensity allows a high density of kelps. On the right, the addition of killer whales as an apex predator limits the sea otter abundance, allowing sea urchin biomass to develop and the resulting grazing intensity constrains the kelp density to low levels (redrawn from Estes et al., 1998). Heavy arrows represent strong trophic interactions; light arrows represent weak interactions.
(Paine, 1980; Estes and Duggins, 1995). They were thought to be a relatively unusual sort of food web mechanics, and a form of biological instability (Strong, 1992) restricted to particular types of marine ecosystems (Hall, 1999). However, new examples are
emerging from studies in several contrasting ecosystems, suggesting that cascade effects can be revealed in diverse marine ecosystems, even in unexpected places such as the open ocean (Pace et al., 1999). Trophic cascades can have strong impacts on
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ecosystems and can stabilize them in alternate states. An example is the otter–urchin–kelp interaction in Alaska (Estes and Duggins, 1995). When sea otters, considered to be keystone species, are abundant, they stabilize a system of abundant kelp forest by reducing grazing by sea urchins (Fig. 7.7, left). When present in low abundance, sea otters shift the system to urchin dominance, with substantial reductions in kelp coverage and productivity (Fig. 7.7, right). As stated by Pace et al. (1999), this illustrates how trophic cascades can induce dramatic shifts in both the appearance and properties of ecosystems. Another example is the possibility that feeding by the pink salmon (Oncorhynchus gorbuscha) controls summer macrozoo- and phytoplankton biomass in the sub-Arctic North Pacific. By exploring relationships between species at different trophic levels, the biomass of the planktivorous pink salmon was found to be inversely related to zooplankton biomass, which in turn is inversely related to phytoplankton biomass (Shiomoto et al., 1997). In pelagic marine ecosystems, alterations of consumer abundance can cascade down food webs to affect phytoplankton biomass, but this effect is rarely detected (Micheli, 1999). Zooplanktivores tend to decrease mesozooplankton abundance, but the mesozooplankton commonly has no effect on the phytoplankton, making the loose coupling between herbivores and plants pervasive (Micheli, 1999). Humans interact with top predators. In western Alaska, killer whales (Orcinus orca) recently may have begun to prey on sea otters, driving a population decline with drastic effects on urchins and kelps (Fig. 7.7, right). It is possible that the behaviour of killer whales toward sea otters has changed recently due to the collapse of their preferred food, the marine mammals. Further, the cause of the decrease in pinniped abundance might be related to overfishing and climate change (Estes et al., 1998). As stressed by Pace et al. (1999), this provides a good example of one of many cases in which it appears that fisheries and fish management are altering trophic cascades, with profound consequences for food webs
in coastal ecosystems. Humans also compete with top predators for valuable marine resources. South African fur seals feed on several commercially important fish species. The problem apparently is an easy one to solve. If top predators compete with fisheries, then fisheries should also compete with predators by culling the expanding seal population. In fact, and as illustrated by Yodzis (2001), the likely results of doing so are controversial as no obvious cascading effects or probable increase of fish resource are to be expected from such culling. This might be the case for most mammal populations for which direct competition with fisheries appears to be limited (Trites et al., 1997). Even though direct competition between fisheries and marine mammals for prey appears to be rather limited, indirect competition might occur for primary production, which sustains both fisheries and marine mammals. The rapid expansion of fisheries may thus lead to so-called ‘food-web competition’ (Trites et al., 1997). Trophic cascades are transitory dynamic interactions and hence exhibit variations in their strength and duration. Not all cascades propagate to lower trophic levels or have significant impacts on ecosystem processes, as numerous compensatory mechanisms dampen or eliminate them (Pace et al., 1999). Fishing usually greatly reduces the abundance of top predators, and it stands to reason that the abundance of prey populations and their effects on marine communities will increase after release from predator control (Steneck, 1998). Many trophic cascades that formerly arose might have disappeared after decades of intense fishing (Steneck, 1998). In this instance, defining proper baselines for both fisheries and conservation objectives will be laborious. Few species are keystone species, and sometimes a keystone species is only revealed in certain configurations of the ecosystem. It would be unreasonable to manage fisheries by arguing solely that a particular species is a keystone species and that cascades actually occur, unless strong evidence supports such mechanisms (Hall, 1999). Despite such difficulties, adopting a top-down approach can help us to understand several observed ecological patterns and
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to get a glimpse, at the ecosystem level, of the possible consequences of removing top predators. Although much attention has been devoted to determining which species are keystones in various marine ecosystems, it may be more useful for management purposes to focus on the strength of interactions between species, as proposed by Mills et al. (1993). Support for this approach lies in the fact that exploited species are rarely keystones, and thus changing their abundances may have small or inconsistent effects on their prey or competitors (Jennings and Kaiser, 1998). At the same time, removing large proportions of forage species may have impacts on their prey, competitors and predators similarly large to those species with typical ‘keystone’ attributes (Shannon and Cury, unpublished data). The effect of removing large portions of pelagic fish species, such as anchovies and sardines, on the functioning of upwelling systems is explored in the next section.
Wasp-waist Control: Small Pelagic Fish Drive Both the Very Large and the Very Small Dominant species in upwelling systems Many marine ecosystems share one striking aspect in the structure of their biological communities: they typically contain a large number of species at lower trophic levels. They also contain a substantial number of predatory fish, seabirds or marine mammals that feed at the upper and near-apex trophic levels. However, in many of the productive ecosystems of the world, and particularly in upwelling ecosystems (Canary, Benguela, California and Humboldt currents), there is an intermediate trophic level, occupied by a limited number of species of small, plankton-feeding pelagic fish, comprising massive populations that are exploited intensively and vary radically in abundance (Cury et al., 2000). Small pelagic fish seem to exert a major control on energy flows in upwelling ecosystems, and this has been termed wasp-waist
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control (Fig. 7.8a). In 20 open marine ecosystems, Micheli (1999) found that inter-annual fluctuations in mesozooplankton biomass were negatively correlated with those of zooplankton-feeding fish, indicating that fish predation controls mesozooplankton biomass. Similar top-down control of zooplankton by sardine, sardinellas, herring or anchovy was also detected off South Africa, Ghana, Japan, in the Black Sea (Cury et al., 2000), and in the northern Baltic (Arrhenius, 1997). Conversely, bottom-up control of fish predators by small pelagic fish has been observed as predatory fish suffer from the collapse of their prey in the Benguela, the Guinea and the Humboldt currents (Cury et al., 2000). Once food becomes abundant again, the recovery of the depleted predator biomass may be immediate, or delayed by shorter or longer periods, highlighting the complex response of the ecosystem to change. Many top predators are buffered against large fluctuations in their food supply by, for example in the case of seabirds, high annual survivorship, protracted longevity, delayed sexual maturity and a relatively low reproductive rate (Hunt et al., 1996). Successful cases of prey switching have been recorded among several seabirds off California and South Africa (Crawford, 1999). However, this plasticity in life history characteristics is sometimes insufficient to dampen the effects of longer term fluctuations in prey resources (Crawford, 1999). This was the case off Namibia in the 1970s, when sardine in the diets of birds was replaced mainly by horse mackerel and pelagic goby. Because these fish were either distributed too far to the north, or occurred too deep in the water column, they were unavailable to penguin and gannet colonies situated south of Lüderitz, causing massive decreases in seabird populations in this region (Crawford et al., 1985). These examples illustrate wasp-waist control, where small pelagic fish constitute mid-trophic-level populations that exert both top-down control on zooplankton and bottom-up control on top predators (Fig. 7.8b). The collapse of a dominant prey can generate drastic changes at higher, but,
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Fig. 7.8. (a) Wasp-waist control within a simplified four-level food web in a marine ecosystem. (b) The abundance of the prey fish (small pelagic fishes), which is dependent on the environment, controls both the abundance of predators and primary producers. A decrease in abundance of prey fish negatively affects abundance of the predators. The same decrease in abundance of prey fish reduces predation on zooplankton, which becomes more abundant. A larger zooplankton population increases grazing pressure and diminishes phytoplankton abundance (the control factor is a dashed line and the responses are a solid line). The environment is considered here to have a direct physical effect on pelagic fish recruitment (sensu Sinclair, 1988), but no effect on primary productivity.
most surprisingly, at lower, trophic levels. As fisheries have removed substantial amounts of small pelagic fish during recent decades, one must carefully consider the implications for the other component species. Again, it appears useful to state that numerous compensatory mechanisms tend to dampen or eliminate expected straightforward consequences, as predators can switch to another type of prey or can migrate to other feeding grounds. In spite of this, as noted by Cury et al. (2000), it is doubtful that the global pelagic fish catch will continue to increase at an annual rate of 4.3%, as has been the case worldwide since the 1950s, without any ecosystem disruptions at different trophic levels.
Discussion – From ‘Common Sense’ and ‘Pet Concept’ Toward an Ecological Framework for Dealing with Responsible Fisheries in Marine Ecosystems? The success of fisheries management in the future will depend on research directed at the mechanisms underlying ecosystem dynamics and fisheries interactions (Murawski, 1991). ‘Ecosystem management’ presumes a reasonable understanding of the interactions among and between species complexes, as well as with their environment (Larkin, 1996). Nevertheless, the ability of marine ecology to
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contribute multispecies and ecosystem information useful to fisheries management has remained very limited (Botsford et al., 1997; Hall, 1999). There are several reasons for this. One is the lack of detailed knowledge of most of the dynamic interactions that underlie multispecies modelling efforts (Rose et al., 1996). Another is the intrinsic complexity of ecological systems that are driven by interactions at multiple levels and scales. Three different theoretical ways of considering energy flows through ecosystems have been presented. Considering bottom-up, top-down or wasp-waist control produces different model ecosystem dynamics, and consequently different possible ecosystem responses to fisheries activity and management (Figs 7.2b, 7.4b and 7.8b). Obviously, the difficulty lies in the ability to determine the controlling factors within an ecosystem. Climate (bottom-up) as well as fishers (topdown) alter the functioning of marine ecosystems. However, an ecosystem is not driven entirely by only one type of control or another, but by a subtle and changing combination of control types that might depend on the ecosystem’s state, diversity and integrity. Several case studies have been presented to illustrate the functioning of marine ecosystems, but clearly only the few that are documented currently have been used. Notably, alteration of the carrying capacity, species replacement or the existence of keystone species and trophic cascades are notoriously difficult to demonstrate. No general theory can be ascribed to the functioning of marine ecosystems, except in the light of the evolutionary theory, which results in poor predictive power for fisheries management. Recently, tentative and partial generalizations have been proposed, for example, that trophic cascades are found mostly in lakes, or in hard substrata marine ecosystems and mainly for less complex food webs, whereas wasp-waist control is most probable in upwelling systems. This restrains the field of possibilities and introduces opportunities for stimulating comparisons and generalizations. As mentioned earlier, the definition of ‘ecosystem’ is fairly new, and the interest of the vast majority of marine ecologists is even more recent. Terrestrial ecology has a long
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tradition of studying ecosystems and has its own ‘pet’ concepts. Several ‘nomad’ concepts, such as keystone species or trophic cascades, can be applied in terrestrial as well as in aquatic studies. However, due to strong differences that exist between these ecosystem types (Chase, 2000; Cury and Pauly, 2000), more attention should be given to promoting the development of new concepts for the functioning of marine ecosystems that will integrate such specificities (Cury, 1994; Frank and Leggett, 1994; Bakun, 2001). Nevertheless, ecological understanding and models of ecosystem functioning are provisional and subject to change (Christensen et al., 1996), and common sense is not sufficient when studying complex dynamic systems. For the time being, we must admit ignorance of the true importance of the effects of fisheries acting through species interactions in marine systems (Hall, 1999). Several decades might be necessary for marine ecologists to refine concepts and to find the appropriate data to strengthen their theories on the functioning of marine ecosystems. These difficulties do not mean that an ecosystem approach to fisheries management should be abandoned or that we should just wait for additional results on the functioning of ecosystems. Major steps are urgently needed to define an ecological framework for dealing with responsible fisheries in marine ecosystems. Ecological questions have to be addressed on the right scale, which often means an uncomfortably large scale (May, 1999). Comparative and retrospective studies between marine ecosystems should be widely promoted to assess changes due to human activities (Jackson et al., 2001). Moreover, another objective should be to evaluate states and changes in marine ecosystems by defining new ecosystem indicators, to assess the usefulness of these indicators for management purposes and to apply them to various ecosystems. This is a complex issue that needs to integrate our simplistic and disparate views of nature (Cury and Cayré, 2001). A framework for defining sustainable reference systems and indicators is already being promoted (Garcia and Staples, 2000), and new, meaningful indicators can be used to assess the impact of fisheries on ecosystems
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(Christensen, 2000; Pauly et al., 2000). These contributions constitute major steps towards a new framework for fisheries management incorporating our recent and incomplete, but consequential, theoretical background on the functioning of marine ecosystems.
Acknowledgements This chapter constitutes an overview of numerous works and comprehensive syntheses. Thanks to M. Sinclair, A. Bakun, V. Christensen and P. Fréon for helpful comments. We thank Cathy Boucher, who drew the figures, and Penny Krohn for the references. This study was funded by FAO and IRD as part of the IDYLE Research Unit dedicated to the study and modelling of marine ecosystems. This work is part of SCOR-IOC Working Group 119 activities to develop quantitative ecosystem indicators for fisheries management.
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fishery management science. In: Pitcher, T.J., Hart, P.J.B. and Pauly, D. (eds) Reinventing Fisheries Management. Chapman and Hall, London, pp. 331–358. Bakun, A. (2001) ‘School-mix feedback’: a different way to think about low frequency variability in large mobile fish populations. In: McKinnell, S.M., Brodeur, R.D., Hanawa, K., Hollowed, A.B., Polovina, J.J. and Zhang, C.-I. (eds) Pacific Climate Variability and Marine Ecosystem Impacts from the Tropics to the Arctic. Progress in Oceanography 49, pp. 485–511. Bakun, A. and Cury, P. (1999) The ‘school trap’: a mechanism promoting large-amplitude out-of-phase population oscillations of small pelagic fish species. Ecology Letters 2(6), 349–351. Bax, N.J. (1991) A comparison of the fish biomass flow to fish, fisheries and mammals in six marine ecosystems. ICES Marine Science Symposium 193, 217–224. Bax, N.J. (1998) The significance and prediction of predation in marine fisheries. ICES Journal of Marine Science 55, 997–1030. Beverton, R.J.H. (1984) Dynamics of single species. In: May, R.M. (ed.) Exploitation of Marine Communities. Springer Verlag, Berlin, pp. 13–58. Beverton, R.J.H. (1990) Small marine pelagic fish and the threat of fishing; are they endangered? Journal of Fish Biology 37(Supplement A), 5–16. Bianchi, G., Gislason, H., Graham, K., Hill, L., Jin, X., Koranteng, K., Manickchand-Heileman, S., Paya, I., Sainsbury, K., Sanchez, F. and Zwanenburg, K. (2000) Impact of fishing on size composition and diversity of demersal fish communities. ICES Journal of Marine Science 57, 558–571. Botsford, L.W., Castilla, J.C. and Peterson, C.H. (1997) The management of fisheries and marine ecosystems. Science 277, 509–515. Bond, W.J. (1993) Keystone species. In: Schulze, E.-D. and Mooney, H.A. (eds) Biodiversity and Ecosystem Function. Springer Verlag, Berlin, pp. 237–253. Butterworth, D.S. (1983) Assessment and management of pelagic stocks in the southern Benguela region. In: Sharp, G.D. and Csirke, J. (eds) Proceedings of the Expert Consultation to Examine Changes in Abundance and Species Composition of Neritic Fish Resources. San Jose, Costa Rica, April 1983. FAO Fisheries Report No.291 (2). Carpenter, S.R. and Kitchell, J.F. (1993) The Trophic Cascade in Lake Ecosystems. Cambridge University Press, Cambridge.
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Hunt, G.L., Barrett, R.T., Joiris, C. and Montevecchi, W.A. (1996) Seabird/fish interactions: an introduction. In: Hunt, G.L. and Furness, R.W. (eds) Seabird/Fish Interactions, with Particular Reference to Seabirds in the North Sea. ICES Cooperative Research Report, No.216, pp. 2–5. Hunter, M.D. and Price, P.W. (1992) Playing chutes and ladders: heterogeneity and the relative roles of bottom-up and top-down forces in natural communities. Ecology 73, 724–732. Jackson, J.B.C. et al. (2001) Historical overfishing and the recent collapse of coastal ecosystems. Science 293, 629–638. Jennings, S. and Kaiser, M.J. (1998) The effects of fishing on marine ecosystems. Advances in Marine Biology 34, 201–352. Jones, R. (1982) Ecosystems, food chains and fish yields. In: Pauly, D. and Murphy, G.I. (eds) Theory and Management of Tropical Fisheries. ICLARM Conference Proceedings No.9, pp. 195–239. Kitchell, J.F., Boggs, C.H., He, X. and Walters, C.J. (1999) Keystone predators in the central Pacific. In: Ecosystem Approaches for Fisheries Management. Alaska Sea Grant College Program, AK-SG-99–01, Fairbanks, pp. 665–683. Klyashtorin, L.B. (1997) Global climatic cycles and Pacific forage fish stock fluctuations. In: Forage Fishes in Marine Ecosystems. Proceedings of the International Symposium on the Role of Forage Fishes in Marine Ecosystems. Alaska Sea Grant College Program Report No. 97–01. University of Alaska, Fairbanks, pp. 545–557. Larkin, P.A. (1996) Concepts and issues in marine ecosystem management. Reviews in Fish Biology and Fisheries 6, 139–164. Lawton, J.H. and Brown, V.K. (1993) Redundancy in ecosystems. In: Schultze, E.-D. and Mooney, H.A. (eds) Biodiversity and Ecosystem Function. Springer Verlag, Berlin, pp. 255–270. Likens, G. (1992) An Ecosystem Approach: Its Use and Abuse. Excellence in Ecology, Book 3. Ecology Institute, Oldendorf/Luhe, Germany. Lundvall, D., Svanbäck, R., Persson, L. and Byström, P. (1999) Size-dependent predation in piscivores: interactions between predator foraging and prey avoidance abilities. Canadian Journal of Fisheries and Aquatic Sciences 56, 1285–1292. May, R.M. (1974) Stability and Complexity in Model Ecosystems. Princeton University Press, Princeton, New Jersey. May, R.M. (1999) Crash tests for real. Nature 398, 371. May R.M., Beddington, J.R., Clark, C.W., Holt, S.J. and Laws, R.M. (1979) Management of multispecies fisheries. Science 205, 267–277.
McCann, K. (2000) The diversity–stability debate. Nature 405(11), 228–233. McGowan, J.A., Cayan, D.R. and Dorman, L.M. (1998) Climate–Ocean variability and ecosystem response in the Northeast Pacific. Science 281, 210–217. Micheli, F. (1999) Eutrophication, fisheries, and consumer-resource dynamics in marine pelagic ecosystems. Science 285, 1396–1398. Mills, L.S., Soule, M.E. and Doak, D.F. (1993) The keystone-species concept in ecology and conservation. BioScience 43(4), 219–224. Murawski, S.A. (1991) Can we manage our multispecies fisheries? Fisheries 16(5), 5–13. Neunfeldt, S. and Koster, F.W. (2000) Trophodynamic control on recruitment success in Baltic cod: the influence of cannibalism. ICES Journal of Marine Science 57(2), 300–309. Oleson, N.J. (1995) Clearance potential of jellyfish Aurelia aurita, and predation impact on zooplankton in a shallow cove. Marine Ecology Progress Series 124(1–3), 63–72. Pace, L.P., Cole, J.J., Carpenter, S.R. and Kitchell, J.F. (1999) Trophic cascades revealed in diverse ecosystems. Trends in Ecology and Evolution 14(12), 483–488. Paine, R.T. (1966) Food web complexity and species diversity. American Naturalist 1000, 65–75. Paine, R.T. (1980) Food webs: linkage, interaction strength and community infrastructure. Journal of Animal Ecology 49, 667–685. Pauly, D., Christensen, V., Froese, R. and Palomares, M.L. (2000) Fishing down aquatic food webs. American Scientist 88, 46–51. Polovina, J.F., Mitchum, G.T., Graham, N.E., Craig, M.P., Demartini, E.E. and Flint, E.N. (1994) Physical and biological consequences of a climatic event in the Central North Pacific. Fisheries Oceanography 3, 15–21. Pope, J.G. and Knights, B.J. (1982) Simple models of predation in multi-age multispecies fisheries for considering the estimation of fishing mortality and its effects. In: Mercer, M.C. (ed.) Multispecies approaches to fisheries management advice. Canadian Special Publication in Fisheries and Aquatic Science 59, 64–69. Power, M.E., Tilman, D., Estes, J.A., Menge, B.A., Bond, W.A., Mills, L.S., Daily, G., Castilla, J.C., Lubchenco, J. and Paine, R.T. (1996) Challenges in the quest for keystones. Bioscience 46(8), 609–620. Rice, J. (1995) Food web theory, marine food webs, and what climate change may do to northern marine fish populations. In: Beamish, R.J. (ed.) Climate change and northern fish populations. Canadian Special Publication in Fisheries and Aquatic Science 121, 561–568.
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Rose, K.A., Tyler, J.A., Singh-Dermot, D. and Rutherford, A.S. (1996) Multispecies modeling of fish populations. In: Megrey, B. and Moksness, E. (eds) Computers in Fisheries Research. Chapman and Hall, London, pp. 195–221. Sanford, E. (1999) Regulation of keystone predation by small changes in ocean temperature. Science 283, 2095–2097. Schwartzlose, R.A. et al. (1999) Worldwide largescale fluctuations of sardine and anchovy populations. South African Journal of Marine Science 21, 289–347. Serra, R., Cury, P. and Roy, C. (1998) The recruitment of the Chilean sardine Sardinops sagax and the ‘optimal environmental window.’ In: Durand, M.H., Cury, P., Mendelssohn, R., Roy, C., Bakun, A. and Pauly, D. (eds) From Local to Global Changes in Upwelling Systems. ORSTOM, Paris, pp. 267–274. Shannon, L.J. and Jarre-Teichmann, A. (1999) Comparing models of trophic flows in the northern and southern Benguela upwelling systems during the 1980s. In: Ecosystem Approaches for Fisheries Management. University of Alaska Sea Grant, AK-SG-99-01, Fairbanks, pp. 55–68. Sheldon, R.W., Sutcliffe, W.H., Jr and Paranjape, M.A. (1977) Structure of pelagic food chains and the relationship between plankton and fish production. Journal of the Fisheries Research Board of Canada 34, 2344–2353. Shin, Y.-J., and Cury, P. (2001) Exploring fish community dynamics through size-dependent trophic interactions using a spatialized individual-based model. Aquatic Living Resources 14(2), 65–80. Shiomoto, A., Tadokoro, K., Nagasawa, K. and Ishida, Y. (1997) Trophic relations in the subarctic North Pacific ecosystem: possible feeding effect from pink salmon. Marine Ecology Progress Series 150, 75–85. Sinclair, M. (1988) Marine Populations: an Essay on Population Regulation and Speciation. Washington Sea Grant Program. University of Washington Press. Sinclair, M. (1997) Prologue. Recruitment in fish populations: the paradigm shift generated by ICES Committee A. In: Chambers R.C. and Trippel, E.A. (eds) Early Life History and Recruitment in Fish Populations. Chapman and Hall, London, pp. 1–27. Sissenwine, M.P. (1984) Why do fish populations vary? In: May, R.M. (ed.) Exploitation of Marine Communities. Springer Verlag, Berlin, pp. 59–94.
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Skud, B.E. (1982) Dominance in fishes: the relation between environment and abundance. Science 216, 144–149. Smetacek, V. (1999) Revolution in the ocean. Nature 401, 647. Soutar, A. and Isaacs, J.D. (1974) Abundance of pelagic fish during the 19th and 20th centuries as recorded in anaerobic sediment off California. Fishery Bulletin 72, 257–274. Steneck, R.S. (1998) Human influences on coastal ecosystems: does overfishing create trophic cascades? Trends in Ecology and Evolution 13, 429–430. Stokes, T.K. (1992) An overview of the North Sea multispecies work in ICES. In: Payne, A.I.L., Brink, K.H., Mann, K.H. and Hilborn, R. (eds) Benguela Trophic Functioning. South African Journal of Marine Science 12, 1051–1060. Strong, D.R. (1992) Are trophic cascades all wet? Differentiation and donor-control in speciose ecosystems. Ecology 73(3), 747–754. Tansley, A.G. (1935) The use and abuse of vegetational concepts and terms. Ecology 16, 284–307. Trites, A., Christensen, V. and Pauly, D. (1997) Competition between fisheries and marine mammals for prey and primary production in the Pacific Ocean. Journal of North West Atlantic Fisheries Science 22, 173–187. Trites, A., Livingston, P., Vasconcellos, M.C., Mackinson, S., Springer, A.M. and Pauly, D. (1999) Ecosystem considerations and the limitations of ecosystem models in fisheries management: insights from the Bering Sea. In: Proceedings of the Symposium on Ecosystem Considerations in Fisheries Management. Alaska Sea Grant College Program Report No. 99–01. University of Alaska, Fairbanks, pp. 609–619. Ursin, E. (1973) On the prey size preferences of cod and dab. Meddr. Danm. Fisk.- og Havunders 7, 85–98. Van Der Lingen, C.D. (1994) Effect of particle size and concentration on the feeding behaviour of adult pilchard Sardinops sagax. Marine Ecology Progress Series 109, 1–13. Verity, P.G. (1998) Why is relating plankton community structure to pelagic production so problematic? In: Pillar, S.C., Moloney, C.L., Payne, A.I.L. and Shillington, F.A. (eds) Benguela Dynamics. South African Journal of Marine Science 19, 333–338. Yodzis, P. (2001) Must top predators be culled for the sake of fisheries? Trends in Ecology and Evolution 16, 78–88.
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Food Webs in the Ocean: Who Eats Whom and How Much? Andrew W. Trites
Marine Mammal Research Unit, Fisheries Centre, University of British Columbia, Vancouver, Canada
Abstract Over 100 food webs have been published for marine ecosystems to describe the transfer of food energy from its source in plants, through herbivores, to carnivores and higher order predators. The webs suggest that the lengths of the chains that form food webs are typically short (3–4 links), and that ecosystems with long food chains may be less stable than those with shorter food chains. Stomach contents have been the primary means for determining what marine organisms eat. More recently developed techniques include faecal analysis and fatty acid signatures from blood or fat samples. Consumption has been estimated from the volume of food found in stomachs, from the feeding rates of captive individuals and from bio-energetic modelling. Consumption of marine organisms, expressed as a percentage of an individual’s body weight per day, ranges from about 4–15% for zooplankton, to 1–4% for cephalopods, 1–2% for fish, 3–5% for marine mammals and 15–20% for sea birds. Immature age classes consume about twice as much (per unit of body weight) as do mature individuals. Furthermore, consumption is not constant throughout the year, but varies with seasonal periods of growth and reproduction. Most groups of species consume 3–10 times more than they produce, and export or pass up the food web about 70–95% of their production. Marine organisms tend to be larger at successive trophic levels and are limited in the sizes of food they can consume. Humans are one of the few species that can prey upon almost any level of the food chain and any size of prey. Food web analysis and estimates of consumption are essential for understanding which ecosystems can support additional species, and which may be less stable and susceptible to species loss through the synergistic effects of fishing or culling. They are also critical tools for understanding changes in ecosystem dynamics as highlighted by a case study from the eastern Bering Sea.
Introduction The understanding of predator–prey relationships within marine food webs has increased enormously since the first simple food web was drawn for herring in the North Sea (Hardy, 1924). What were once simple
qualitative depictions of the interrelationships between species are being replaced increasingly with quantitative descriptions of entire ecosystems. This in turn is allowing fisheries biologists to estimate better the amount of food consumed by various species, and to make predictions about the effects of
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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fisheries on predator–prey relationships, and of the effects of food web dynamics on fisheries. The following provides an overview of marine food webs and the amounts of food that marine organisms consume. Emphasis is placed on apparent patterns that a number of biologists have noted in comparing food webs among different ecosystems regarding ecosystem stability, food chain lengths and food web complexity. The efficiency of energy transfer between trophic levels is also discussed in the context of digestion and assimilation efficiencies of marine organisms; and the interplay between food webs and consumption is highlighted with a case study from the eastern Bering Sea.
Food Webs A food chain delineates one possible pathway for the transfer of energy from plants to herbivores, to carnivores and top predators. Myriads of food chains within an ecosystem form a food web, which biologists typically have drawn as a series of interconnected species joined by lines showing the presence or absence of interactions (e.g. Fig. 8.1). This is the simplest depiction of who eats whom. Species may be identified individually or may be grouped into functional categories (e.g. groundfish or benthic invertebrates) based on similar life history characteristics or other traits. A more refined depiction of food webs (energy flow food webs) highlights the trophic level of each species and the relative strength of the interactions (based on the amount of energy flowing from producers and consumers, e.g. Fig. 8.2). Here, species or groups of species are placed according to their trophic level (calculated as 1.0 plus the mean trophic level of the species that they consume), and the size of each box is relative to the biomass of the species. In general, trophic levels of functional groups tend to cluster about integer values (e.g. Fig. 8.2). Energy flow food webs convey considerably more information about the ecosystem than do topological food web drawings.
However, both depictions of food webs (energy flow and topological) fail to convey which interactions are critical to maintaining the ecosystem in its normal state. For example, regardless of the amount of energy consumed or produced in a near-shore marine ecosystem, removing limpets would have little effect compared with removing urchins, because urchins dramatically alter the physical structure of the ecosystem by consuming kelp. Functional food webs thus differ from the other two depictions of food webs by highlighting the linkages that are most important to community structure (Paine, 1980; Huxham et al., 1995). However, functional relationships are not yet well enough understood to make this a practical means of depicting food webs. Food webs are built by knowing which producers are eaten, and in what proportion consumers eat them. Stomach contents have been the principal source of dietary data and continue to be the main tool for identifying the numbers, sizes and types of species consumed by fish (e.g. Christensen, 1995). Seabird and marine mammal diets are determined in a similar manner (e.g. Pauly et al., 1998; Santos et al., 2001), although attempts are being made to determine diet from alternative means. Fatty acid signatures are one of the newer methods of identifying prey of whales and seals, using milk samples, blood samples and blubber biopsies (e.g. Grahl-Nielsen and Mjaavatten, 1991; Iverson et al., 1997; Kirsch et al., 1998). Seal diets are being determined more commonly now from the hard parts (bones, eye lenses and beaks) found in stomachs or faecal remains collected from sites where animals haul out and rest (e.g. Tollit et al., 1997; Kirkman et al., 2000; Cottrell and Trites, 2002). Stable isotope analysis is yet another recently developed technique for determining diet (e.g. Wada et al., 1991; Kaelher et al., 2000; Kelly, 2000).
Comparative analyses Over the past 80 years, food web research has sought to reduce complexity and identify recurring patterns that might infer underlying mechanisms or represent constraints on
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Fig. 8.1. A simplified depiction of the Bering Sea food web: (1) ice algae; (2) phytoplankton; (3) copepods; (4) mysids and euphausids; (5) medusae; (6) hyperid amphipods; (7) seabirds; (8, 9) pelagic fishes; (10) walrus; (11) seals; (12) basket stars; (13) ascidians; (14) shrimps; (15) filter-feeding bivalves; (16) sand dollars; (17) sea stars; (18) crabs; (19) bottom feeding fishes; (20) polychaetes; (21) predatory gastropods; (22) deposit feeding bivalves (from McConnaughey and McRoy, 1976).
ecosystem structure (e.g. Summerhayes and Elton, 1923; Elton, 1927; Cohen, 1977, 1978; Pimm, 1982, 1991; Lawton and Warren, 1988; Lawton, 1989; Cohen et al., 1990; Winemiller, 1990; Christensen and Pauly, 1993; Hall and Raffaelli, 1996; Raffaelli, 2000). Based on comparisons of food webs (mostly terrestrial ecosystems), a number of generalizations have
been postulated about the sizes of organisms, the lengths of food chains and the stability of ecosystems. Elton (1927) was among the first to remark that animals occupying successively higher trophic levels tended to be larger, and that there were upper and lower limits on the size of food that they can eat. Humans are one
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Fig. 8.2. Flowchart of trophic interactions in the eastern Bering Sea during the 1980s. All flows are in t km−2 year−1. Minor flows are omitted, as are all backflows to the detritus. The size of each box is roughly proportional to the biomass therein (from Trites et al., 1999).
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of the few species that can consume prey of any size and can feed at all trophic levels. As noted by Krebs (1996), this is one of the reasons that humans are so biologically successful. Omnivorous species are relatively rare, although there are species of fish that eat their way up the food chain as they grow. Organisms at the base of the food web tend to be more numerous than species at higher trophic levels, and are often represented as a pyramid of numbers and size (total biomass) to assess the relative distribution of biomass among trophic levels within an ecosystem (Fig. 8.3). Overall, system biomass is proportional to primary production (Pimm, 1982), and the proportion of species occupying top, intermediate and basal trophic levels appears to
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be constant across food webs (Cohen, 1978). There also appears to be a relatively constant ratio of 2–3 species of prey for every predator in an ecosystem (Martinez, 1991), although numbers of species of prey consumed by each species of predator tend to increase as the size of the food web increases. A number of biologists have noted that food chains typically are short (Elton, 1927; Hutchinson, 1959; Pimm and Lawton, 1977; Pimm, 1982; Ricklefs and Miller, 2000). However, the average number of links appears to be longer in marine ecosystems compared with freshwater communities, grasslands or wet tropical forests (Briand and Cohen, 1987; Ricklefs and Miller, 2000). Christensen and Pauly (1993) compared 41 aquatic ecosystems
Fig. 8.3. Trophic pyramids representing the distribution of biomass and energy flow in four ecosystems. The pyramids are scaled so that the volume at each trophic level corresponds to the sum of all flows at that level. The top angles are inversely proportional to the transfer efficiency (acute angle = high efficiency) (from Trites et al., 1999).
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and found average path lengths of 2–5 linkages. The longest average path lengths were in tropical estuaries (3.0–5.0), followed by tropical shelves (2.8–4.0) and oceanic upwelling areas (2.2–2.8) (Baird et al., 1991; Christensen and Pauly, 1993). Maximum chain lengths were eight in tropical shelves, seven in tropical estuaries and six in oceanic upwelling areas (Christensen and Pauly, 1993). This is supported further by a review of 75 aquatic food webs that found only three with maximum food chains longer than six (Schoener, 1989). It appears that the average number of feeding links per species (linkage density) increases as the size of the web (i.e. number of species) increases. This implies that the number of prey that a predator will eat increases in proportion to the total number of species in that community. However, the number of links relative to all possible links (connectance) decreases as the number of groups in a food web increases (Pimm, 1982; Christensen and Pauly, 1993). Martens (1987) suggested that this might reflect an increase in ecosystem stability. However, Christensen and Pauly (1993) conclude that any interpretation of connectance is ambiguous due to the binary nature of its scoring (i.e. either a link exists or it does not, irrespective of the fraction of diet it represents). The length of food chains may be a function of the amount of primary production at the base of the food pyramid, and the efficiency with which energy is transferred from one trophic level to the next (Hutchinson, 1959; Slobodkin, 1960). Higher rates of transfer efficiency presumably mean that more energy can be passed up the food chain to support more species. However, Pimm (1991) disputes this explanation, noting that some areas of high ocean productivity have short food chains. He argues instead that longer food chains are associated with stable environments, while shorter food chains are in less predictable environments. Species at the end of long food chains would be at risk of extinction if the abundance of species lower in the food chain fluctuated severely. In general, it appears that food webs in variable environments have fewer linkages than webs in more constant environments
(Briand, 1983). However, it has also been shown experimentally that food chains are longer in more productive environments (Pimm and Kitching, 1987; Jenkins et al., 1992), and that population dynamics are less stable in long food chains than in short ones (Lawler and Morin, 1993). Thus it appears that the length of food chains is a function of both environmental stability and energy transfer efficiency.
Limitations of food webs Food webs, such as those shown in Figs 8.1 and 8.2, are collages of species interactions that sometimes conceal more than they reveal (Paine, 1988; Raffaelli, 2000). For example, some of the interactions may not occur simultaneously, or they may change over seasons or years. They also tend to be oversimplifications. Similarly, some interactions, such as parasite/pathogen–host interactions or mutualistic interactions, may be critical to community dynamics but fail to be captured by food web depictions (Cohen, 1993; Paine, 1994; Huxham et al., 1995; Hall and Raffaelli, 1996). Food web diagrams are useful tools for conceptual understanding of ecosystems despite their shortcomings and the inadequacy of many of the food webs used in comparative analyses (see critiques by Paine, 1988; Polis, 1991; Hall and Raffaelli, 1993, 1996). With the advent of quantitative ecosystem modelling tools (e.g. Christensen and Pauly, 1992; Jørgensen, 1998), food web analysis is leading towards a better understanding of food web structure and the design of better management strategies for conservation. The structure of food webs has implications for community persistence (Pimm, 1991), and may provide insights into which systems can support additional species, and which are unstable and susceptible to species losses. This is particularly relevant to understanding the effects on food webs of culling (e.g. Yodzis, 1998), overfishing (e.g. Christensen, 1998; Hacquebord, 1999; Jackson et al., 2001) or introducing exotic species (e.g. Grosholz et al., 2000). Food web analysis may also help to
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identify which processes are critical to the ecosystem, and highlight which components need more research (Wooten, 1994).
Food Consumption Consumption can be expressed as the total amount eaten, or as the rate (i.e. velocity) that energy is ingested. For the purpose of this review, consumption is defined as the amount of food eaten in a fixed period of time (i.e. days or years). Organisms require energy for growth, reproduction, physical activity and maintenance of cells and organs, and must ingest sufficient calories to meet these fixed costs of living. However, organisms cannot assimilate all of the energy contained within the food they ingest due to differences in the nutrient content and digestibility of different types of prey. For example, an organism that ingests excess nitrogen in its diet will excrete it as energy in the form of nitrogen-containing organic waste. Thus, energy that is digested and absorbed (assimilated energy) is a function of digestive physiology and of the make up of the prey (e.g. proportion of bones, scales, exoskeleton, etc.). Digestibility of ingested prey may also vary by season. Generally, foods of animal origin are easier to digest than foods of plant origin, and vertebrates are easier to digest than invertebrates. Thus assimilation efficiency depends on the quality of the diet (particularly the amount
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of digestion-resistant structural material it contains) and varies from about 15 to 90%. Non-assimilated energy – excreted in the form of waste – contributes to detritus and is consumed by species specializing in digesting recalcitrant materials. Within the context of an ecosystem, no energy is wasted energy. Assimilated energy can be used by the organism for the synthesis of new biomass (production) through growth and reproduction, which can be consumed by organisms at higher trophic levels. The fraction of ingested food that is used for production is lowest in organisms whose costs of maintenance and activity are greatest, and highest in species with low maintenance and activity costs (e.g. 0.2–0.3% gross food conversion efficiency in mammals, 1.1–1.5% in birds, 15–30% in fish and 30–40% in cephalopods; see Table 8.1).
Estimating energy requirements There are a number of ways to estimate the amount of food consumed by marine organisms. Stomach content analysis is one method of determining daily ration, but involves making a number of assumptions about frequency of feeding and seasonal changes in energy requirements (e.g. Jarre et al., 1991). Another approach is to infer feeding rates from those of captive-fed individuals, or to measure the metabolism of free-ranging individuals (e.g. via doubly labelled water) or of captive individuals (via oxygen exchange)
Table 8.1. Approximate rates of consumption, growth, efficiency and turnover for six major species groups. Q/B is the ratio of ingested energy to biomass, and is expressed as a daily and annual rate of consumption. Population growth rate is expressed as the ratio of annual production to biomass (P/B), and gross food efficiency (P/Q) is the fraction of ingested gross energy that is converted into production (growth). Turnover rate is the average residency time of energy within each species group (expressed in years). Species Sea birds Mammals Fish Crabs and shrimp Squid Zooplankton
Consumption Q/B day−1
Consumption Q/B year−1
Growth rate P/B year−1
Gross efficiency P/Q
Turnover rate B/P years
0.15–0.20 0.03–0.05 0.01–0.02 0.02–0.05 0.01–0.04 0.04–0.15
55–73 11–18 3–8 8–20 4–15 15–55
0.80–50.0 0.02–0.06 0.60–2.50 1.50–3.00 1.50–4.50 2.50–6.50
0.011–0.015 0.002–0.003 0.150–0.300 0.150–0.200 0.300–0.400 0.120–0.170
1.25–50.00 17.00–50.00 0.40–1.70 0.30–0.70 0.20–0.70 0.15–0.40
Source: based on Christensen (1995) and Trites et al. (1999).
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(e.g. Winberg, 1956; Mann, 1978; Innes et al., 1987). An alternative, but practically impossible method in aquatic systems, is to derive consumption from estimates of biomass missing from lower trophic levels. A fifth approach is modelling. Models typically synthesize information (either measured or assumed from related species) about the costs of basal metabolism, activity, growth, reproduction, excreted waste and assimilation efficiencies – and range from simple to detailed accountings. The simplest approach estimates food consumption as a function of food conversion efficiency and mean individual weight (e.g. Kendeigh et al., 1977; Croxall et al., 1984; Pauly, 1986, 1989; Innes et al., 1987; Christensen, 1995; Guinet et al., 1996; Trites et al., 1997; Palomares and Pauly, 1998). A more detailed approach – but more difficult to parameterize – calculates the costs of Respiration (the energy used for work that is degraded to heat), Production (energy deposited into tissue growth, fat storage, eggs, sperm, etc.), and Faeces and Urine (energy excreted from the body) such that Consumption = Faeces + Urine + Respiration + Production
for an organism that is in energy balance (e.g. Klekowski and Duncan, 1975; Stenson et al., 1997; Winship et al., 2002). Multiplying the mean individual consumption by population density yields an estimate of total consumption by a group of organisms. Energy requirements are a function of body size and assimilation efficiency of different diets. Metabolic needs for maintenance are a function of body weight raised to the power of 0.70–0.75, and are lower per unit of body mass for larger species or individuals compared with smaller ones (Kleiber, 1975). However, young animals have significantly higher rates of consumption compared with mature individuals due to the high energetic cost of growth. Thus, estimating consumption requires estimates of animal density, assimilation efficiency and performance (growth, activity and maintenance), which in turn depend on animal physiology, and on the digestibility and other nutritional properties of the food. On average, consumption of marine organisms (expressed as a percentage of an
individual’s body weight per day) ranges from about 4–15% for zooplankton, to 1–4% for cephalopods, 1–2% for fish, 3–5% for marine mammals, and 15–20% for sea birds (calculated from Q/B ratios in Christensen, 1995; Trites et al., 1999, Table 8.1). However, consumption is not constant throughout the year, but varies with seasonal changes in growth and reproduction (e.g. Paul et al., 1993, 1998; Winship et al., 2002). Some organisms, such as baleen whales, may only feed for 6 months of the year, while other may fast for periods of days or weeks (Pauly et al., 1998). Finally, the nutritional quality and energy content of prey may also change seasonally (e.g. Paul et al., 1993, 1998).
Transfer efficiencies and turnover times The number of trophic levels in marine ecosystems averages between four and six, and appears to increase from coastal systems to reefs and shelves, and decline for upwelling systems (Christensen and Pauly, 1993). The fraction of energy that enters one trophic level and passes on to the next higher trophic level (transfer efficiency) typically is low (Lindeman, 1942), and decreases with higher trophic levels due to increased respiration (Burns, 1989). As a general rule of thumb, gross food conversion efficiency decreases at higher trophic levels (Table 8.1), and transfer efficiency (assimilation or gross production at level n divided by that at level n–1) remains constant between trophic levels, at about 10% (Kozlovsky, 1968; Pauly and Christensen, 1995). A comparison of aquatic ecosystems suggests ecological efficiencies of about 10% for herbivores and detritivores, 11% for the next trophic level, and lower efficiencies of 7.5–9.0% at higher levels (Christensen and Pauly, 1993). Most energy held within a trophic level is dissipated before organisms feeding at the next higher trophic level can consume it. Ultimately, the amount of energy reaching each trophic level depends on the net primary production at the base of the food chain and on the conversion efficiencies. Thus, high fish catches are associated with high
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primary productivity and with fishing at lower trophic levels (Christensen, 1996). The turnover or residency time of energy at each trophic level can be calculated by dividing the biomass (total stored energy) by net productivity (the rate at which energy is converted into biomass). This effectively is the time it takes energy to flow through the ecosystem (Ricklefs and Miller, 2000). Average turnover times in marine ecosystems range from about 6 days for phytoplankton to 3 months for zooplankton, to 5 months for cephalopods, 6 months for crabs and shrimp, 1 year for fish, 17 years for seals and 50 years for whales (calculated from the inverse of the P/B ratios in Table 8.1, Christensen, 1995; Trites et al., 1999). Upwelling systems generally have shorter turnover times than shelves, reefs and estuaries (Christensen and Pauly, 1993). Longer residence times reflect greater accumulations of energy. Turnover times for aquatic primary production are extremely short compared with terrestrial systems (Ricklefs and Miller, 2000).
Food Webs and Consumption in the Bering Sea A considerable amount of research currently is being focused on the eastern Bering Sea ecosystem, due largely to the decline of Steller sea lions (1977 to present) and their listing by the USA in 1997 as an endangered species (Alverson, 1992; Trites and Larkin, 1996; Loughlin, 1998). The cause of the population decline is not yet known, but may be related to a decline in prey abundance caused by fisheries or by natural changes in the ecosystem (Anon., 1993). Food web analysis and estimates of prey consumption are essential pieces of information needed to resolve this issue.
Bering Sea food web dynamics Stomach contents and biomass estimates of fish, invertebrates, birds and mammals were collected by the then Soviet Union during the 1950s, and by the US National Marine Fisheries Service since the 1970s. Food webs
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constructed for the 1950s and 1980s (before and after the Steller sea lion decline started) suggest that the linkages between species were the same in each era (Fig. 8.2), but that dramatic changes occurred in the biomass of each group and in the amount of energy passing from one group to another (Trites et al., 1999). The mass balance ecosystem models constructed by Trites et al. (1999) suggest that most of the top predators (Trophic level IV) declined from the 1950s to the 1980s, along with a significant number of mid-trophic level species (i.e. crabs, shrimp and forage fishes such as herring and sandlance – Trophic level III). Species that increased dramatically during this period (Fig. 8.4) included walleye pollock (level III) and large flatfish (level IV). Pollock appear to have contributed over 50% of the energy transferred at the mid-trophic levels during the 1980s compared with only 10% in the 1950s. In contrast, pelagic fishes contributed nearly 50% of the total Bering Sea energy flow in the 1950s. Stomach contents of Steller sea lions in the Gulf of Alaska suggest changes in their diet that are consistent with stock assessments and model predictions (Table 8.2). During the 1950s, Steller sea lion stomach samples contained mostly pelagic fishes (herring and sandlance), some gadids (pollock and cod) and no flatfish. From the 1960s to the 1990s, however, the dietary concentration of pelagic fishes fell, while gadids and flatfish became more prevalent (Table 8.2). Attempts to simulate the effects of commercial fishing on the Bering Sea ecosystem failed to explain the change in ecosystem dynamics between the two eras (Trites et al., 1999). Fishing could not explain the decline of forage fish species (most of which were never fished), nor could it explain the 60% increase in large flatfish or the 400% increase in pollock. Food web interactions could not account for the magnitude of changes that occurred in the eastern Bering Sea. Instead, it appears that the survival of a suite of species was favoured over another by a combination of natural environmental changes in water temperatures and ocean currents. This is commonly referred to as a regime shift (Francis and Hare, 1994; Anderson et al., 1997; Beamish et al., 2000; Benson and Trites, 2002)
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Fig. 8.4. Estimated trophic levels and relative biomass of species in the eastern Bering Sea during the 1980s. Black boxes indicate groups that had lower abundance in the 1980s relative to the 1950s, and shaded boxes show species that had higher abundance in the 1980s relative to the 1950s. Major flows of energy between the boxes are shown in Fig. 8.2 (from Trites et al., 1999). Table 8.2. Changes in the proportion (1970s–1980s) or probability (1990s) of major prey types occurring in diets of Steller sea lions in the Gulf of Alaska (%). Note that the sum of the percentages do not add up to 100%. Period 1950–1960s 1976–1978 1985–1986 1990–1993
Gadids
Flatfish
Pelagics
Few 32 60 85
None 0 5 13
Mostly 61 20 18
Source: from Merrick et al. (1997).
and suggests that the Bering Sea can exist in at least two alternative stable states that support suites of species at alternatively high and low population levels (Trites et al., 1999). Ecosystem indices from the models suggest that the Bering Sea was more ‘mature’ (sensu Odum, 1971) in the 1950s than in the 1980s, but is overall relatively resilient and resistant to perturbations (Trites et al., 1999). Pollock and/or Atka mackerel (a hexagramid related to lingcod) currently dominate
the diets of Steller sea lions in the declining populations that border the Bering Sea. This is in sharp contrast to the diets of growing sea lion populations in southeast Alaska and British Columbia that contain a more diverse array of prey (i.e. salmon, rockfish, forage fish, gadids and flatfish). Dietary diversity correlates with the rates of sea lion population change (Merrick et al., 1997), and suggests that the recovery of Steller sea lions is linked to consuming a more diverse range of species with higher fat (energy) contents than they currently are obtaining. Recent field observations of foraging Steller sea lions suggest that they might choose preferentially herring over pollock (Thomas and Thorne, 2001). This is consistent with predictions of the junk-food hypothesis, which proposes that Steller sea lions have declined because they have been consuming too much pollock, which contain fewer calories, and not enough of the fattier, highenergy fishes (Alverson, 1992; Rosen and Trites, 2000b).
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Estimating energy requirements Understanding how a change in diet could affect Steller sea lions comes from laboratory analysis of sea lions (digestive efficiencies and metabolism) and their prey (caloric and nutritional value). For example, bomb calorimetry analysis shows that herring (6–11 kJ g−1) has more energy per unit mass than do salmon (5–9 kJ g−1), pollock (3–5 kJ g−1) or squid (4–6 kJ g−1) (from data compiled by Winship and Trites, 2003). Feeding trials with captive Steller sea lions show that digestive efficiency (90–95% of gross energy intake) increases with prey energy density, while heat increment of feeding (10–20% of gross energy intake) decreases with increasing prey energy density (Rosen and Trites, 1997, 1999, 2000a). Digestive efficiency is the proportion of usable energy within a prey, and heat increment of feeding is the proportion of energy that is burnt during the mechanical and biochemical processes of digesting a prey item. Thus Steller sea lions can digest prey with higher fat content more easily than they can digest leaner prey. It also turns out that Steller sea lions have to burn more of the energy contained within larger prey to digest them than they do from smaller meal sizes (Rosen and Trites, 1997). Thus a Steller sea lion would have to eat an average of 56% more pollock than herring to obtain an equivalent amount of energy because pollock are bigger prey, contain fewer calories and require more energy to digest than do herring (Rosen and Trites, 2000b). Estimating the amount of prey that Steller sea lions require can be determined by incorporating estimates of diet composition, digestive efficiencies, heat increments of feeding, activity budgets, body growth, basal metabolism and population size into bio-energetics models (Winship and Trites, 2003; Winship et al., 2002). They indicate that Steller sea lions in southeast Alaska require more food in winter and spring than they do during summer and autumn (~ 45–60% more due primarily to seasonal changes in the energy density of the diet), and that the average sea lion requires about 17 kg of prey day−1, or 6000 kg year−1. Within
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different regions of Alaska, per caput food requirements differ by as much as 24%, depending upon the relative amounts of energy-poor prey (gadids) versus energy-rich prey (e.g. forage fish and salmon) that Steller sea lions consume (Fig. 8.5; Winship and Trites, 2003). In 1998, the biomass of pollock and cod in Alaska was estimated at 11.57 million t (8.98 million t pollock, 2.59 million t cod), with an annual natural mortality of 3.56 million t (2.68 million t pollock, 0.88 million t cod), and a total commercial catch of 1.52 million t (1.25 million t pollock, 0.27 million t cod) (see Winship and Trites, 2003). Total annual gadid consumption by Steller sea lions for all regions of Alaska was 0.18 million t (Winship and Trites, 2003). Thus, consumption of gadids by Steller sea lions represented about 2% of the stock size, or 5% of natural mortality or 12% of commercial landing. Gadid consumption by Steller sea lions appears to be small relative to total gadid natural mortality and stock size, which is consistent with conclusions drawn by Livingston (1993). Differences in the quality of prey available to Steller sea lions have consequences on the individual, and ultimately the population. One strategy an animal can invoke when faced with an energy shortage is to reduce the amount of prey they require by reducing their energy expenditures. Captive experiments have shown that Steller sea lions can reduce their metabolism by an average of 31% when food is withheld from them, which is typical of a fasting response (Rosen and Trites, 2002). However, sea lions do not appear to reduce their metabolism when fed smaller meals, despite losing body mass (Rosen and Trites, 2002). Instead, they exhibit a hunger response, which might lead them to increase their foraging effort in the wild. However, increased foraging effort has an increased energetic cost, as well as an increased risk of exposure to predation by sharks or killer whales. Recently weaned young are the segment of the population most likely to incur the greatest cost of reduced caloric intake. Energetically, a 1-year-old female must consume
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Fig. 8.5. Estimated annual food biomass requirements (thousands of t) for Steller sea lions in 1998 in seven study areas of Alaska, assuming that the summer diet was consumed all year long. Pie charts represent the proportions of diet biomass represented by each prey species category. Diameters of the pie charts are proportional to their respective mean food requirement estimates. Food requirements were calculated assuming digestive efficiency and the heat increment of feeding for maintenance varied with prey energetic density (from Winship and Trites, 2003).
about twice the relative quantity of prey compared with a mature female (13% of her body weight per day eating a mixed diet, compared with 6% for a mature female; Winship et al., 2002). This same yearling female would require 9% of her body mass per day if she ate only small schooling fish, or 17% if she ate
only gadids (Winship et al., 2002). These energy requirements could well be twice these values (i.e. 18–34% of body mass per feeding trip) considering that Steller sea lions do not eat every day, but typically fast for a day between their 1- or 2-day feeding trips (Trites and Porter, 2002).
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Synthesis Combining the individual energy requirements with information about the changes that occurred in the Bering Sea food web suggests that the population decline of Steller sea lions in Alaska was probably related to the inability of young animals to acquire sufficient energy from the low-quality prey available to them. There is no indication of there being a shortage of low quality prey. However, the energetic modelling and captive feeding studies suggest it may not be physically possible for young Steller sea lions to consume enough low-energy prey to meet their daily energetic needs. Consuming fewer calories can stunt growth and cause reproductive failure (i.e. abortions) – symptoms that have been observed in Alaska over the time that Steller sea lions have declined (Calkins et al., 1998; Pitcher et al., 1998). A lower nutritional plane may also increase the susceptibility of sea lions to disease, and increase their risk of being killed by predators – a factor that may account for the apparent high mortality of juvenile sea lions (York, 1994). Mathematical modelling suggests that killer whales could have been a significant contributing factor in the decline of Steller sea lions, and may now be preventing the population from recovering (Barrett-Lennard et al., 1995).
Conclusions The Bering Sea case study is an example of the importance of constructing food webs and estimating the energy requirements of marine organisms to understand ecosystem dynamics. This can only be achieved through a combination of fieldwork, captive studies and mathematical models – all of which are essential tools for the responsible management of fisheries and ecosystems. Food webs and energetic modelling shed light on the consequences of removing or adding organisms, and on the role that humans play in shaping the dynamics of marine ecosystems. More importantly,
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they are essential techniques for recognizing what our marine ecosystems once were, what they are currently and what they might be in the future.
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Martens, B. (1987) Connectance in linear and Volterra systems. Ecological Modelling 35, 157–163. Martinez, N.D. (1991) Artifacts or attributes? Effects of resolution on the Little Rock Lake food web. Ecological Monographs 61, 367–392. McConnaughey, T. and McRoy, P. (1976) Food-web structure and the fraction of carbon isotopes in the Bering Sea. In: Science in Alaska, 1976. Proceedings of the 27th Alaska Science Conference, Fairbanks, Alaska, 4–7 August 1976. Alaska Division of the American Association for the Advancement of Science, Fairbanks, pp. 293–316. Merrick, R.L., Chumbley, M.K. and Byrd, G.V. (1997) Diet diversity of Steller sea lions (Eumetopias jubatus) and their population decline in Alaska: a potential relationship. Canadian Journal of Fisheries and Aquatic Sciences 54, 1342–1348. Odum, E.P. (1971) Fundamentals of Ecology, 2nd edn. Saunders, Philadelphia. Paine, R.T. (1980) Food webs: linkage, interaction strength and community infrastructure. Journal of Animal Ecology 49, 667–685. Paine, R.T. (1988) Food webs: road maps of interactions or grist for theoretical development? Ecology 69, 1648–1654. Paine, R.T. (1994) Marine Rocky Shores and Community Ecology: an Experimentalist’s Perspective. Ecology Institute, Oldendorf. Palomares, M.L. and Pauly, D. (1998) Predicting food consumption of fish populations as functions of mortality, food type, morphometrics, temperature and salinity. Marine Freshwater Research 49, 447–453. Paul, A.J., Paul, J.M. and Smith, R.L. (1993) The seasonal changes in somatic energy content of Gulf of Alaska yellowfin sole, Pleuronectes asper. Journal of Fish Biology 43, 131–138. Paul, A.J., Paul, J.M. and Smith, R.L. (1998) Seasonal changes in whole-body energy content and estimated consumption rates of age 0 walleye pollock from Prince William Sound, Alaska. Estuarine, Coastal and Shelf Science 47, 251–279. Pauly, D. (1986) A simple method for estimating the food consumption of fish populations from growth data and food conversion experiments. Fishery Bulletin 84, 827–840. Pauly, D. (1989) Food consumption by tropical and temperate fish populations: some generalizations. Journal of Fish Biology 35 (Supplement A), 11–20. Pauly, D. and Christensen, V. (1995) Primary production required to sustain global fisheries. Nature 376, 279–279.
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A.W. Trites
Pauly, D., Trites, A.W., Capuli, E. and Christensen, V. (1998) Diet composition and trophic levels of marine mammals. Journal of Marine Science 55, 467–481. Pimm, S.L. (1982) Food Webs. Chapman and Hall, London. Pimm, S.L. (1991) The Balance of Nature? University of Chicago Press, Chicago, Illinois. Pimm, S.L. and Kitching, R.L. (1987) The determinants of food chain lengths. Oikos 50, 302–307. Pimm, S.L. and Lawton, J.H. (1977) Number of trophic levels in ecological communities. Nature 268, 329–331. Pitcher, K.W., Calkins, D.G. and Pendleton, G.W. (1998) Reproductive performance of female Steller sea lions: an energetics based reproductive strategy? Canadian Journal of Zoology 76, 2075–2083. Polis, G.A. (1991) Complex trophic interactions in deserts: an empirical critique of food-web theory. American Naturalist 138, 123–155. Raffaelli, D. (2000) Trends in research on shallow water food webs. Journal of Experimental Marine Biology and Ecology 250, 223–232. Ricklefs, R.E. and Miller, G.L. (2000) Ecology, 4th edn. Freeman, New York. Rosen, D.A.S. and Trites, A.W. (1997) Heat increment of feeding in Steller sea lions, Eumetopias jubatus. Comparative Biochemistry and Physiology 118A, 877–881. Rosen, D.A.S. and Trites, A.W. (1999) Metabolic effects of low-energy diet on Steller sea lions, Eumetopias jubatus. Physiological and Biochemical Zoology 72, 723–731. Rosen, D.A.S. and Trites, A.W. (2000a) Digestive efficiency and dry matter digestibility of Steller sea lions fed herring, pollock, squid and salmon. Canadian Journal of Zoology 78, 234–239. Rosen, D.A.S. and Trites, A.W. (2000b) Pollock and the decline of Steller sea lions: testing the junk-food hypothesis. Canadian Journal of Zoology 78, 1243–1258. Rosen, D.A.S. and Trites, A.W. (2002) Changes in metabolism in response to fasting and food restriction in the Steller sea lion (Eumetopias jubatus). Comparative Biochemistry and Physiology 132, 389–399. Santos, M.B., Clarke, M.R. and Pierce, G.J. (2001) Assessing the importance of cephalopods in the diets of marine mammals and other top predators: problems and solutions. Fisheries Research 52, 121–139. Schoener, T.W. (1989) Food webs from the small to the large. Ecology 70, 1559–1589. Slobodkin, L.B. (1960) Ecological energy relationships at the population level. American Naturalist 94, 213–236.
Stenson, G.B., Hammill, M.O. and Lawson, J.W. (1997) Predation by harp seals in Atlantic Canada: preliminary consumption estimates for Arctic cod, capelin and Atlantic cod. Journal of Northwest Atlantic Fishery Science 22, 137–154. Summerhayes, V.S. and Elton, C.S. (1923) Contributions to the ecology of Spitzbergen and Bear Island. Journal of Ecology 11, 214–286. Thomas, G.L. and Thorne, R.E. (2001) Night-time predation by Steller sea lions. Nature 411, 1013–1013. Tollit, D.J., Steward, M.J., Thompson, P., Pierce, G.J., Santos, M.B. and Hughes, S. (1997) Species and size differences in the digestion of otoliths and beaks: implications for estimates of pinniped diet composition. Canadian Journal of Fisheries and Aquatic Sciences 54, 105–119. Trites, A.W. and Larkin, P.A. (1996) Changes in the abundance of Steller sea lions (Eumetopias jubatus) in Alaska from 1956 to 1992: How many were there? Aquatic Mammals 22, 153–166. Trites, A.W. and Porter, B.T. (2002) Attendance patterns of Steller sea lions (Eumetopias jubatus) and their young during winter. Journal of Zoology (London) 256, 547–556. Trites, A.W., Christensen, V. and Pauly, D. (1997) Competition between fisheries and marine mammals for prey and primary production in the Pacific Ocean. Journal of Northwest Atlantic Fishery Science 22, 173–187. Trites, A.W., Livingston, P.A., Mackinson, S., Vasconcellos, M.C., Springer, A.M. and Pauly, D. (1999) Ecosystem change and the decline of marine mammals in the Eastern Bering Sea: testing the ecosystem shift and commercial whaling hypotheses. Fisheries Centre Research Reports 7(1). Wada, E., Mizutani, H. and Minagawa, M. (1991) The use of stable isotopes for food web analysis. Critical Reviews in Food Science and Nutrition 30, 363–371. Winberg, G.G. (1956) Rate of metabolism and food requirements of fishes. Fisheries Research Board of Canada Translation Series No. 194. Winemiller, K.O. (1990) Spatial and temporal variation in tropical fish trophic net-works. Ecological Monographs 60, 331–367. Winship, A. and Trites, A.W. (2003) Prey consumption by Steller sea lions in Alaska: how much do they require? Fishery Bulletin, in press. Winship, A., Trites, A.W. and Rosen, D.A.S. (2002) A bioenergetics model for estimating the food requirements of Steller sea lions (Eumetopias jubatus) in Alaska. Marine Ecology Progress Series 229, 291–312.
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9
Regional Assessments of Prey Consumption and Competition by Marine Cetaceans in the World Tsutomu Tamura The Institute of Cetacean Research, Tokyo, Japan
Abstract The total annual prey consumption by 45 species (12 baleen whale species; 33 toothed whale species) out of 84 species (13 baleen whale species; 71 toothed whale species) of marine cetaceans in the world was assessed. The assessment was based on: (i) recently available abundance estimates of cetaceans; (ii) daily prey consumption rates of cetaceans estimated using three methods; (iii) estimated biomass of cetaceans by use of average body weight and abundance; and (iv) composition of prey species of cetaceans. The annual prey consumption of cetaceans was estimated for three ocean regions: the southern hemisphere, including the Indian Ocean (120–242 million t); the North Pacific (63–85 million t); and the North Atlantic (55–107 million t). Total annual prey consumption by cetaceans in the world was estimated to be at least 249–434 million t, because the data did not cover all cetacean species nor their whole range. The fish consumption by cetaceans in the southern hemisphere including the Indian Ocean was estimated to be 18–32 million t and equated to 66–120% of commercial fisheries catches in 1996. In the North Pacific, fish consumption was estimated to be 21–31 million t, equivalent to 67–99% of commercial fisheries catches in 1996. In the North Atlantic, the fish consumption by cetaceans was 15–25 million t, equivalent to 87–144% of commercial fisheries catches in 1996. There was probably direct competition between cetaceans and commercial fisheries in the North Pacific and the North Atlantic. However, as the information on the abundance of cetaceans was not included for all species, the actual figures for annual prey consumption by all cetaceans are probably larger than the results presented. More information on abundance, body weight and prey composition of cetaceans in each region is necessary to consider competition between cetaceans and commercial fisheries in order to address a more realistic strategy for fisheries management and the conservation of cetaceans in future.
Introduction Eighty-four species of cetaceans (whales, dolphins and porpoises) inhabit the world, of which 79 species live in the sea (13 baleen whales and, 71 toothed whales; IWC, 2001). As cetaceans are mammals, they need a large amount of energy to maintain their body
temperature in the aquatic environment. Thus, they must consume large amounts of prey. The cetaceans are top predators of the food web, playing an important role in the marine ecosystem. It is important to elucidate the total prey consumption of cetaceans in the world, because these results can provide information useful to address the issue of
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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long-term sustainability of marine living resources. This kind of work has practical difficulties, because the data available are incomplete and many assumptions are needed for such a study. The quantitative data on feeding ecology of top predators are insufficient. Some researchers have estimated the food consumption of cetaceans based on techniques such as calculations of energyrequirements (Hinga, 1979; Lockyer, 1981; Innes et al., 1986; Sigurjónsson and Víkingsson, 1997; T. Tamura and Y. Fujise, 2000, unpublished data). Furthermore, in recent years, some scientists have tried to understand competition between fish stock and cetaceans in small areas using statistical simulation models. This study calculates the daily and annual prey consumption of cetaceans, based on available recent abundance estimates, average body weight of each species and three methods used previously (T. Tamura and S. Ohsumi, 2000, unpublished data). This study was carried out for three regions: the southern hemisphere including the Indian Ocean (SHIO), the North Pacific (NP) and the North Atlantic (NA). Furthermore, the competition between cetaceans and fisheries in each region is indicated. In addition, an assessment of the competition between cetaceans (common minke whales (Balaenoptera acutorostrata) and Bryde’s whales (B. edeni)) and fisheries in the Northwest Pacific is considered, based on JARPN (The Japanese Whale Research Program under Special Permit in the Western North Pacific, 1994–1999) and JARPN II (2000) results.
Materials and Methods Regional assessments of prey consumption by cetaceans
Available abundance estimates The Scientific Committee of the International Whaling Commission (IWC) uses estimates of current abundance based on direct methods such as sighting surveys. However, such data are limited for species and regions, and
some of them are out of date. In this study, the estimates of abundance are from a recent report by T. Tamura and S. Ohsumi (2000, unpublished data). In the SHIO, the figures on abundance of 15 species out of 56 species of cetaceans found in this region are shown in Table 9.1–1. For the NP, the figures on abundance of 31 species out of 39 species of cetaceans found in this region are shown in Table 9.1–2, and for the NA, figures on abundance of 16 species out of 39 species of cetaceans in this region are shown in Table 9.1–3.
Estimation of biomass of cetaceans based on abundance and average body weight Biomass estimates are based on recent abundance estimates on each cetacean species from published sources, and estimated average body weights by use of the formula of Trites and Pauly (1998). They estimated the mean body weight by sex in each species. For sperm whales (Physeter macrocephalus) in the southern hemisphere, the average male weight was used. Biomass of each cetacean species in three ocean regions was calculated by multiplying abundance by average body weight.
Estimation of daily prey consumption of cetaceans Daily prey consumption was calculated using the rate of prey intake per body weight of each cetacean species per day (feeding rate: percentage of body weight) and average body weight. This terminology of ‘the feeding rate’ was proposed by Sergeant (1969). In this study, the daily prey consumption and feeding rate for each species was estimated based on three methods. METHOD 1: ESTIMATION OF DAILY PREY CONSUMP-
Innes TION FROM AVERAGE BODY WEIGHT et al. (1986) proposed the following method to estimate daily prey consumption of a cetacean from its average body weight: I = 0.42M0.67
(9.1)
where I is daily prey consumption (kg day−1) and M is average body weight (kg).
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Table 9.1. Abundance data for marine cetaceans in each region. 1–1. Southern hemisphere including Indian Ocean. Species
Area
Blue whale
South of 30°S
Pigmy blue whale
Abundance
CV
1,255
0.36
5,000
N.e.
Fin whale
South of 30°S
85,200
N.e.
Sei whale
South of 30°S
10,8601
N.e.
Bryde’s whale Antarctic minke whale Humpback whale Southern right whale Pygmy right whale Sperm whale
South of 60°S South of 30°S
South of 30°S
89,000 761,000
N.e. 0.14–0.28
10,000 7,000
0.27 N.e.
N.D. 209,0002
95% CL
IWC (1996); Perry et al. (1999) Gambell (1976); Perry et al. (1999) IWC (1979); Perry et al. (1999) IWC (1980); Mizroch et al. (1984); H.W. Braham (1991, unpublished data); Perry et al. (1999) Ohsumi (1981) 510,000–1,140,000 IWC (1991a) 5,900–16,800
N.e. 0.44–0.46
Pygmy sperm whale Dwarf sperm whale Arnoux’s beaked whale Southern bottlenose whale South of 50°S Beaked whales3
N.D. N.D. N.D.
N.e. N.e. N.e.
N.D.
N.e.
599,000
0.15
Cuvier’s beaked whale Shepherd’s beaked whale Blainville’s beaked whale Gray’s beaked whale Ginkgo-toothed beaked whale Hector’s beaked whale Pygmy beaked whale True’s beaked whale Strap-toothed whale Andrew’s beaked whale Longman’s beaked whale
N.D.
N.e.
N.D.
N.e.
N.D.
N.e.
N.D. N.D.
N.e. N.e.
N.D.
N.e.
N.D.
N.e.
N.D. N.D. N.D.
N.e. N.e. N.e.
N.D.
N.e.
Source of abundance
IWC (2000) IWC (2000); Perry et al. (1999) D.S. Butterworth, D.L. Borchers, S. Chalis, J.B. DeDecker and F. Kasamatsu (1994, unpublished data) IWC (1995)
450,000–800,000
Kasamatsu and Joyce (1995)
continued
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1–1.
Continued.
Species
Area
Abundance
CV
Irrawaddy dolphin Killer whale
South of 60°S
N.D. 58,5004
N.e. 0.30
South of 60°S
86,5005
0.80–1.04
Long-finned pilot whale
Short-finned pilot whale False killer whale Pygmy killer whale Melon-headed whale Tucuxi Indo-Pacific humpbacked dolphin Rough-toothed dolphin Dusky dolphin Hourglass dolphin South of 50°S
N.D.
N.e.
N.D. N.D. N.D. N.D. N.D.
N.e. N.e. N.e. N.e. N.e.
N.D.
N.e.
N.D. 144,000
N.e. 0.17
Peale’s dolphin Risso’s dolphin Bottlenose dolphin Pantropical spotted dolphin Atlantic spotted dolphin Spinner dolphin Clymene dolphin Striped dolphin Common dolphin6 Fraser’s dolphin Southern right whale dolphin The Magellan Commerson’s strait dolphin Heavlside’s dolphin New Zealand Hector’s dolphin Black dolphin Spectacled porpoise Burmeister’s porpoise Finless porpoise
N.D. N.D. N.D. N.D.
N.e. N.e. N.e. N.e.
N.D.
N.e.
N.D. N.D. N.D. N.D. N.D. N.D.
N.e. N.e. N.e. N.e. N.e. N.e.
3,211
N.e.
N.D. 3,408 N.D. N.D. N.D.
N.e. N.e. N.e. N.e. N.e.
N.D.
N.e.
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95% CL
Source of abundance D.S. Butterworth, D.L. Borchers, S. Chalis, J.B. DeDecker and F. Kasamatsu (1994, unpublished data); IWC (1995) D.S. Butterworth, D.L. Borchers, S. Chalis, J.B. DeDecker and F. Kasamatsu (1994, unpublished data); IWC (1995)
100,000–200,000
Kasamatsu and Joyce (1995)
Leatherwood et al. (1988) Dawson and Slooten (1988)
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Continued (footnotes).
N.D., no data; N.e., not estimated; CV, coefficient of variation; CL, confidence limit. Estimates of abundance ranged from 9,720 to 12,000. 2 Estimates of abundance ranged from 128,000 to 290,000. 3 Arnoux’s beaked whale + Southern bottlenose whale. 4 Estimates of abundance ranged from 53,000 to 64,000. 5 Estimates of abundance ranged from 43,000 to 130,000. 6 Including long-beaked common dolphin. 1
1–2.
North Pacific.
Species
Area
Abundance
CV
3,300
0.24.
Fin whale
16,1251
N.e.
Sei whale
9,110
N.e.
Western Pacific
21,901
0.19
Eastern tropical Pacific Sea of Japan
13,023
0.20
7,600
0.40
Okhotsk Sea-West Pacific
25,000
0.30
7,0002
N.e.
3003
N.e.
Blue whale
Bryde’s whale
Common minke whale
Humpback whale
North Pacific right whale
Eastern Okhotsk Sea
14,781–32,450
12,800–48,600
404–2,108
900
Gray whale
Bering-ChukchiBeaufort Seas Eastern
26,300
0.03
Sperm whale
Western Pacific
102,112
0.16
Eastern tropical Pacific
22,666
0.22
Pygmy sperm whale Dwarf sperm whale Eastern tropical Pacific
N.D. 11,215
N.e. 0.29
Bowhead whale
95% CL
6,400–9,200
7,500
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21,900–32,400
Source of abundance Wade and Gerrodette (1993); Perry et al. (1999) H.W. Braham (1993, unpublished data); Perry et al. (1999) Tillman (1977); Perry et al. (1999) Shimada and Miyashita (1997); IWC (1997b) Wade and Gerrodette (1993) IWC (1984) IWC (1992b)
Calambokidis et al. (1997); Perry et al. (1999) IWC (1998); Perry et al. (1999) IWC (1998); Perry et al. (1999) Raftery and Zeh (1991); IWC (1992a) R.C. Hobbs and D.J. Rugh (1999, unpublished data) H. Kato and T. Miyashita (1998, unpublished data) Wade and Gerrodette (1993) Wade and Gerrodette (1993) continued
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1–2.
Continued.
Species
Area
Abundance
CV
Baird’s beaked whale
Western Pacific
3,950
0.28
Sea of Japan
1,260
0.45
Okhotsk Sea
660
0.27
20,000
0.27
N.D.
N.e.
N.D.
N.e.
N.D.
N.e.
N.D.
N.e.
Eastern tropical Cuvier’s beaked Pacific whale Blainville’s beaked whale Ginkgo-toothed beaked whale Hubbs’ beaked whale Stejneger’s beaked whale Eastern tropical Killer whale Pacific Western Pacific Short-finned pilot whale Western Pacific False killer whale Eastern tropical Pacific Pygmy killer whale Eastern tropical Pacific Eastern tropical Melon-headed Pacific whale Indo-Pacific humpbacked dolphin Eastern tropical Rough-toothed Pacific dolphin Pacific white-sided dolphin Western Pacific Risso’s dolphin Eastern tropical Pacific Bottlenose dolphin Western Pacific Eastern tropical Pacific Pantropical spotted Western Pacific dolphin Northeastern
Spinner dolphin
95% CL
T. Miyashita (1990, unpublished data) T. Miyashita (1990, unpublished data) T. Miyashita (1990, unpublished data) Wade and Gerrodette (1993)
8,500
0.37
53,608
0.22
34,723–82,756
16,668 39,800
0.26 0.64
10,034–27,689
38,900
0.31
45,400
0.47
N.D.
N.e.
145,900
0.32
83,289 289,300
0.18 0.34
58,764–118,049
168,791 243,500
0.26 0.29
102,000–279,044
438,064
0.17
312,285–614,503
730,900
0.14
Western/ southern Coastal
1,298,400
0.15
29,800
0.35
Eastern
631,800
0.24
1,019,300
0.19
Whitebelly
Source of abundance
Wade and Gerrodette (1993) Miyashita (1993a) Miyashita (1993a) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993)
Wade and Gerrodette (1993) 988,000 0.17–1.50 164,000–6,790,000 Miyashita (1993b)
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Miyashita (1993a) Wade and Gerrodette (1993) Miyashita (1993a) Wade and Gerrodette (1993) Miyashita (1993a) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) continued
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Continued.
Species
Area
Abundance
CV
95% CL
Striped dolphin
570,038 Western Pacific Eastern tropical 1,918,000 Pacific Northern 476,300
0.18 0.11
397,435–817,602
Central
406,100
0.38
2,210,900
0.22
289,300
0.34
Common dolphin4
Southern Fraser’s dolphin
Eastern tropical Pacific
Northern right whale dolphin Pacific Dall’s porpoise Okhotsk
Harbour porpoise Vaquita Finless porpoise White whale
0.37
308,000 0.31–1.13 1,186,000
0.09
554,000
Seto Inland Sea
N.D. N.D. 5,000
Alaska USSR
5,800 27,000
Source of abundance
Miyashita (1993a) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) Wade and Gerrodette (1993) 59,000–1,680,000 Miyashita (1993b) 991,000–1,420,000 Buckland et al. (1993) T. Miyashita (1991, unpublished data); IWC (1993)
N.e. N.e. Kasuya and Kureha (1979) IWC (1992c) IWC (1992c)
N.D., no data; N.e., not estimated; CV, coefficient of variation; CL, confidence limit. 1 Estimates of abundance ranged from 14,620 to 18,630. 2 Estimates of abundance ranged from 6,000 to 8,000. 3 Estimates of abundance ranged from 100 to 500. 4 Including long-beaked common dolphin. 1–3. North Atlantic. Species
Area
Blue whale
North Western Atlantic
Fin whale Sei whale
Bryde’s whale Common minke whale
Humpback whale North Atlantic right whale Bowhead whale Sperm whale
Abundance
CV
3301
N.e.
47,300 4,000
N.e.
N.D. North Eastern 118,299 Atlantic Central Atlantic 28,000 West Greenland 3,266 West of Iceland 10,600 North Western 4002 Atlantic 450 190,000
95% CL
Source of abundance
27,723–82,031
H.W. Braham (1991, unpublished data); Perry et al. (1999) IWC (1992d) H.W. Braham (1991, unpublished data); Perry et al. (1999)
96,681–44,750
IWC (1997a)
21,600–31,400 1,790–5,950 9,300–12,100
IWC (1991b) IWC (1991b) Smith et al. (1999) IWC (1986a); Perry et al. (1999) Zeh et al. (1993) Rice, (1989); Odell, (1992) continued
N.e.
0.07 N.e. N.e. N.e.
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1–3.
Continued.
Species
Area
Pygmy sperm whale Dwarf sperm whale Cuvier’s beaked whale Northern bottlenose Iceland whale Blainville’s beaked whale Sowerby’s beaked whale Gervais’ beaked whale True’s beaked whale Killer whale Iceland Long-finned pilot whale
Eastern
Short-finned pilot whale False killer whale Pygmy killer whale Melon-headed whale Atlantic humpbacked dolphin Rough-toothed dolphin White beaked dolphin Atlantic white sided Iceland dolphin Risso’s dolphin Bottlenose dolphin Pantropical spotted dolphin Atlantic spotted dolphin Spinner dolphin Clymene dolphin Striped dolphin Common dolphin1 Fraser’s dolphin Harbour porpoise Iceland White whale
Canada USSR
Abundance
CV
N.D. N.D. N.D.
N.e. N.e. N.e.
44,300
N.e.
N.D.
N.e.
N.D.
N.e.
N.D.
N.e.
N.D. 5,500
N.e. N.e.
778,000
0.30
N.D.
N.e.
N.D. N.D. N.D.
N.e. N.e. N.e.
N.D.
N.e.
N.D.
N.e.
13,420
N.e.
38,680
N.e.
N.D. N.D. N.D.
N.e. N.e. N.e.
N.D.
N.e.
N.D. N.D. N.D. N.D. N.D. 28,510
N.e. N.e. N.e. N.e. N.e. N.e.
45,700 9,500
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95% CL
Source of abundance
Sigurjónsson and Víkingsson (1997)
Sigurjónsson and Víkingsson (1997) 440,000–1,370,000 S.T. Buckland, K.L. Cattanach, Th. Gunnlaugsson, D. Block, S. Lens and J. Sigurjónsson (1992, unpublished data); IWC (1993)
Sigurjónsson and Víkingsson (1997) Sigurjónsson and Víkingsson (1997)
Sigurjónsson and Víkingsson (1997) IWC (1992c) IWC (1992c) continued
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Prey Consumption and Competition by Marine Cetaceans
1–3.
151
Continued.
Species
Area
Abundance
Narwhal
Canada-Green 28,000 land Northern 1,300 Hudson Bay
CV
95% CL 22,000–33,500
Source of abundance IWC (1992c) IWC (1992c)
N.D., no data; N.e., not estimated; CV, coefficient of variation; CL, confidence limit. 1 Estimates of abundance ranged from 100 to 560. 2 Estimates of abundance ranged from 300 to 500. 3 Including long-beaked common dolphin.
METHOD 2: ESTIMATION OF DAILY PREY CONSUMPTION FROM THE STANDARD METABOLISM
Sigurjónsson and Víkingsson (1997) proposed a method for estimation of daily prey consumption from the standard metabolism of each cetacean species. The daily prey consumption is given by: (9.2) D = 206.25M0.783; I = D/1110.3 (for baleen whales in the southern hemisphere) (9.3) D = 206.25M0.783; I = D/1300 (for baleen whales in the northern hemisphere and toothed whales in the world) where D is daily caloric value of prey intake (kcal day−1), M is average body weight (kg) and I is daily prey consumption (kg). The present writer assumed that estimated caloric values of prey were 1110.3 kcal kg−1 for baleen whales in the southern hemisphere (Clark, 1980) and 1300 kcal kg−1 for baleen whales in the northern hemisphere and toothed whales in the world (Steimle and Terranova, 1985). METHOD 3: ESTIMATION OF DAILY PREY CONSUMP-
Klumov TION FROM KLUMOV’S FORMULA (1963) proposed a method for estimating daily prey consumption from the average body weight of each cetacean species, deriving the daily prey consumption from the formula: I = 0.035M
(9.4)
where I is daily prey consumption (kg day−1), and M is average body weight (kg).
Estimation of annual prey consumption of cetaceans The author recalculated the annual prey consumption (C) by each cetacean species in three ocean regions from available abundance estimates (N) and daily prey consumption rates (I) obtained from the above three methods by applying the formula: C = 365NI
(9.5)
Composition of prey species The annual prey consumption in each prey category was calculated using the assumed prey composition (percentage of weight) in each region from published sources (e.g. Pauly et al., 1998). The categories of prey species were divided into three groups, namely fish (pelagic and mesopelagic), cephalopods (squids) and crustaceans (copepods, amphipods and krill).
Catch of marine organisms by fisheries worldwide To compare the amount of prey consumed by cetaceans with the amount of catch by fisheries in the world, the annual catch statistics available from the FAO for marine organisms captured by commercial fisheries were used. FAO divides the seas of the world into a number of major fishing areas for statistical purposes (Fig. 9.1), and for this study these statistical areas were aggregated into three ocean regions, namely SHIO (FAO areas 41 (Southwest Atlantic), 47 (Southeast Atlantic),
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Fig. 9.1. Map of major fishing areas for statistical purposes and ocean regions in this study (SHIO, NP and NA), based on FAO (FAO, 1997).
51 (Western Indian Ocean), 57 (Eastern Indian Ocean), 81 (Southwest Pacific), 87 (Southeast Pacific) and 48, 58 and 88 (Southern Oceans)); NP (61 (Northwest Pacific), 67 (Northeast Pacific), 71 (Western Central Pacific) and 77 (Eastern Central Pacific)); and NA (21 (Northwest Atlantic), 27 (Northeast Atlantic), 31 (Western Central Atlantic), 34 (Eastern Central Atlantic) and 37 (Mediterranean and Black Sea)). Figures for fisheries catches excluding inland fisheries and aquaculture in 1996 were taken from FAO (1998). Catch species in 1996 were divided into four groups (fish, cephalopods, crustaceans and others) based on FAO Fishery statistics (Table 9.2; FAO, 1998). Among them, the category ‘others’ that included seaweed and others was excluded from the analysis as it is not a prey organism of cetaceans.
Competition between whales and fisheries in the Northwest Pacific based on the results of JARPN and JARPN II
Research area and periods Japan conducted a whale research programme in the western North Pacific from 1994 to 1999 under Special Permit, as provided for by Article VIII of the International Convention for the Regulation of Whaling
(ICRW). Since some scientific issues remained outstanding following the 1994–1999 programme (JARPN – The Japanese Whale Research Program under Special Permit in the Western North Pacific, 1994–1999), a second phase of the research – a feasibility study for the years 2000 and 2001 – began in July 2000 (JARPN II). The main emphasis for this phase of the research is feeding ecology, involving studies on prey consumption by cetaceans, prey preferences of cetaceans, and ecosystem modelling. The JARPN research areas between 1994 and 1999 and those of JARPN II in 2000 were a part of subareas 7, 8 and 9, excluding the exclusive economic zone (EEZ) of foreign countries, which were established by the IWC (Fig. 9.2; IWC, 1994). The whales were sampled according to sampling procedures described by Kato et al. (1989). Sampled whales were transported immediately to a research base vessel, where biological measurements and sampling was carried out. A summary of the survey months, years and sample size of common minke, Bryde’s and sperm whales in each subarea is shown in Table 9.3.
Sampling of stomach contents and data analyses After the stomach contents of common minke, Bryde’s and sperm whales were
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Prey Consumption and Competition by Marine Cetaceans
Table 9.2.
153
Catch of each group by commercial fisheries in 1996 (FAO, 1998). Commercial fisheries catch in 1996 (t)
Study region
FAO area
Fish
Southern hemisphere (including Indian Ocean)
48,58,41 48,58,47 48,58,51 48,58,57 48,58,81 48,58,87 48,58,88 Subtotal 48,58,61 48,58,67 48,58,71 48,58,77 Subtotal 48,58,21 48,58,27 48,58,31 48,58,34 48,58,37 Subtotal
1,667,613 1,012,052 3,602,001 3,423,452 16,550,967 16,761,837 1661,9,963 27,027,885 19,716,050 2,750,809 7,505,751 1,240,702 31,213,312 1,013,674 10,467,611 1,275,317 3,112,271 1,277,218 17,146,091 75,387,288
North Pacific
North Atlantic
Total
Cephalopods Crustaceans 1,716,652 1,907,996 1,094,367 1,084,656 1,073,554 1,051,705 1,028,928 1,028,958 1,156,571 1,02,9627 1,282,374 1,188,331 1,627,903 1,038,880 1,058,435 1,031,143 1,192,083 1,060,480 1,381,021 3,037,882
1,081,745 1,010,933 1,323,099 1,267,892 1,907,203 1,063,651 1,101,212 1,855,735 2,329,317 1,098,640 1,675,117 1,089,941 3,193,015 1,392,028 1,261,751 1,257,865 1,049,843 1,045,193 1,006,680 5,055,430
Others
Total
1,908,061 1,028,735 1,011,336 1,081,473 1,906,395 1,150,918 1,028,190 1,258,918 1,764,243 1,030,454 1,375,276 1,049,702 2,219,675 1,586,150 1,240,684 1,132,342 1,908,970 1,111,126 1,079,272 3,557,865
2,474,071 1,031,716 4,030,803 3,857,473 17,638,119 17,028,111 17,111,203 29,171,496 24,966,181 2,880,530 8,838,518 1,568,676 38,253,905 2,030,732 11,028,481 1,696,667 3,363,167 1,494,017 19,613,064 87,038,465
contents were removed and frozen for later analyses. In the laboratory, prey species in the subsamples were identified to the lowest taxonomic level possible, and the relative frequency of occurrence of each prey species and the relative prey importance by weight of each prey species were calculated.
Results Regional assessments of prey consumption by cetaceans Fig. 9.2. Subareas surveyed by the JARPN from 1994–1999. Subareas were based on IWC (1994), excluding the exclusive economic zone of foreign countries. Furthermore, subarea 7 was divided into east (7E) and west (7W).
sampled, each stomach’s contents (with and without liquid) were weighed to the nearest 0.1 kg. Then, for common minke and Bryde’s whales, a subsample (3–4 kg) of forestomach contents was removed and frozen for later analyses. For sperm whales, stomach
Southern hemisphere including the Indian Ocean Estimated annual prey consumption of cetaceans is given in Table 9.4–1. Baleen whales consume 72–148 million t of crustaceans (mainly Euphausia superba). Cephalopods (mainly squids) were consumed only by toothed whales, and amounted to 27–56 million t. The annual crustacean consumption by Antarctic minke whales (B. bonaerensis) accounted for 41–55% of total annual crustacean consumption by baleen whales. The
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154
annual cephalopod consumption by sperm whales accounted for 76–90% of total annual cephalopod consumption by cetaceans. Total fish consumption by cetaceans was 18–32 million t, equivalent to 66–120% of commercial fisheries fish catch (27 million t) in 1996. Total cephalopod consumption by cetaceans was 27–56 million t, which was one order of magnitude greater than the commercial catch of cephalopods (1 million t) in 1996.
Table 9.3. Subareas, months and years of surveys and sample size used in this study. Subarea
Survey month
Year
Sample size
1998 1997 1996 1999 1996 2000 1996 2000 1998 1998 1996 1997 1996 1997 1995 1997 1994 1995 1994 1995 2000 1994 1999 1996
56 2 1 50 15 6 15 18 8 36 11 31 5 27 14 40 8 61 9 25 16 4 50 30 538
Bryde’s whale 7 August September Total
2000 2000
24 19 43
Sperm whale 7 August September Total
2000 2000
4 1 5
Common minke whale 7E May June July June 7W August September
North Pacific 8
Estimated annual prey consumption of cetaceans is given in Table 9.4–2. Baleen whales consumed 2–4 million t of fish and 14–35 million t of crustaceans (copepods and krill). Fin whales (B. physalus) accounted for 9–14% and common minke whales for 47–58% of total annual fish consumption by baleen whales. Consumption by toothed whales was 17–29 million t of fish and 22–28 million t of cephalopods (squids). Sperm whales consumed 12–44% and common dolphins 13–22% of total annual fish consumption by toothed whales. Sperm whales accounted for 36–75% of total annual cephalopod consumption by toothed whales. Total fish consumption by cetaceans was 21–31 million t, equivalent to 67–99% of the commercial fisheries fish catch (31 million t) in 1996. Total cephalopod consumption by cetaceans was 22–28 million t, one order of magnitude greater than the commercial cephalopod catch (1.6 million t) in 1996.
North Atlantic Estimated annual prey consumption of cetaceans is given in Table 9.4–3. Baleen whales consumed 6–11 million t of fish and 15–41 million t of crustaceans (copepods and krill). Fin whales accounted for 5–9% and common minke whales 68–78% of total annual fish consumption by baleen whales. Toothed whales consumed 9–14 million t of fish and 24–39 million t of cephalopods (squids). Sperm whales consumed 60–81% and long-finned pilot whales (Globicephala melas) 15–31% of the total annual fish
9
May June July August May June July August
11
September July August
Total
consumption by toothed whales. Sperm whales accounted for 62–82% of total annual cephalopod consumption by toothed whales. Total fish consumption by cetaceans was 15–25 million t, equivalent to 87–144% of the commercial fisheries catch (17 million t) in 1996. Total cephalopod consumption by cetaceans was 24–39 million t, a value two orders of magnitude greater than the commercial fisheries cephalopod catch (0.4 million t) in 1996.
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Method 1 Species
Biomass (t)
Fish
Method 2
Cephalopods Crustaceans
Fish
Method 3
Cephalopods Crustaceans
Fish
Cephalopods Crustaceans 1,647,144 4,400,988 60,203,295 2,327,635 9,727,719 63,833,175
Blue whale Pigmy blue whale Fin whale Sei whale Bryde’s whale Antarctic minke whale Humpback whale Southern right whale Baleen whales total
128,935 344,500 4,736,268 182,567 1,436,727 4,996,726
0 0 98,651 2,257 4,230,381 0
0 0 0 0 0 0
438,574 1,336,962 19,631,460 1,126,290 4,770,429 42,122,913
0 0 149,965 2,997 5,592,243 0
0 0 0 0 0 0
714,615 2,082,294 29,842,936 1,495,707 6,306,146 50,301,128
0 0 302,529 4,665 8,626,468 0
0 0 0 0 0 0
304,080 163,681
0 0 4,331,288
0 0 0
1,545,767 907,410 71,879,805
0 0 5,745,205
0 0 0
2,194,957 1,250,816 94,188,600
0 0 8,933,662
3,884,622 0 2,091,025 0 0 148,115,601
Sperm whale Beaked whales1 Killer whale Long-finned pilot whale Hourglass dolphin Commerson’s dolphin Hector’s dolphin Toothed whales total
5,630,251 842,194 133,439 73,612
7,447,046 4,722,578 797,282 304,475
20,851,729 5,312,900 159,456 913,424
1,489,409 1,770,967 0 0
8,908,785 4,046,717 721,575 246,510
24,944,597 4,552,556 144,315 739,530
1,781,757 1,517,519 0 0
17,981,614 4,303,611 852,338 235,097
50,348,520 4,841,563 170,468 705,290
3,596,323 1,613,854 0 0
4,896 90 228
117,210 2,295 4,370 13,395,256
117,210 1,377 3,933 27,360,029
0 918 437 3,261,731
65,951 1,263 2,655 13,993,455
65,951 758 2,389 30,450,096
0 505 265 3,300,046
31,273 574 1,458 23,405,967
31,273 345 13,13 56,098,771
0 230 146 5,210,553
17,726,544
27,360,029
75,141,536
19,738,660
30,450,096
97,488,647
32,339,628
56,098,771 153,326,154
Cetaceans total
Prey Consumption and Competition by Marine Cetaceans
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Table 9.4. Estimated the annual prey consumption (tonnes) based on three methods in each region. 4–1. Southern hemisphere including the Indian Ocean.
1
Arnoux’s beaked whale + Southern bottlenose whale.
155
North Pacific. Method 1
Species
Biomass (t)
Fish
Method 2
Cephalopods Crustaceans
Fish
Method 3
Cephalopods Crustaceans
Fish
Cephalopods Crustaceans
339,032 896,389 153,148 563,778 214,052
0 186,707 32,187 392,047 1,263,134
0 63,480 11,360 0 0
1,153,222 3,483,945 903,143 3,139,911 541,343
0 242,407 36,507 442,631 1,288,266
0 82,418 12,885 0 0
1,604,869 4,523,319 1,024,354 3,545,038 552,114
0 572,568 66,520 799,451 1,914,156
0 194,673 23,478 0 0
4,331,135 10,684,125 1,866,471 6,402,814 820,353
Humpback whale North Pacific right whale Bowhead whale Gray whale Baleen whales total
212,856 28,060
186,110 0
0 0
895,927 155,556
225,709 0
0 0
1,086,554 183,136
467,708 0
0 0
2,251,527 358,461
233,070 404,284
0 128,700 2,188,885
0 0 74,841
1,176,327 2,445,293 13,894,667
0 144,504 2,380,025
0 0 95,303
1,430,124 2,745,571 16,695,079
0 258,236 4,078,641
0 0 218,151
2,977,469 4,906,487 34,598,842
2,310,764 1,133 18,414 16,580
3,458,788 3,786 69,329 83,009
9,684,606 30,292 108,945 166,018
691,758 3,786 19,808 27,670
3,966,122 2,409 65,046 67,007
11,105,140 19,276 102,215 134,015
793,224 2,409 18,585 22,336
7,380,002 1,447 82,334 63,543
20,664,005 11,576 129,383 127,086
1,476,000 1,447 23,524 21,181
19,389 34,470
115,844 250,226
23,169 375,340
0 0
104,844 196,274
20,969 294,411
0 0
123,844 176,141
24,769 264,212
0 0
Sperm whale Dwarf sperm whale Baird’s beaked whale Cuvier’s beaked whale Killer whale Short-finned pilot whale
T. Tamura
Blue whale Fin whale Sei whale Bryde’s whale Common minke whale
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4–2.
1
306,765 38,613 47,197 277,643
306,765 64,355 110,127 138,821
0 0 0 46,274
237,742 24,487 30,166 174,822
237,742 40,812 70,386 87,411
0 0 0 29,137
208,478 14,610 18,270 102,886
208,478 24,350 42,629 51,443
0 0 0 17,148
77,064
1,823,542
981,907
0
1,127,000
606,846
0
639,920
344,572
0
83,460 77,511 162,316
214,505 1,583,044 3,137,698
1,823,296 527,681 3,137,698
107,253 0 0
149,354 1,080,623 1,899,642
1,269,511 360,208 1,899,642
74,677 0 0
106,620 742,649 1,036,791
906,271 247,550 1,036,791
53,310 0 0
67,695 288,612 247,464 27,484 32,340
1,828,235 5,530,155 6,253,635 562,492 533,654
1,218,823 3,225,923 2,680,129 328,121 533,654
0 460,846 0 46,874 0
1,050,698 3,574,558 3,875,993 355,468 341,079
700,465 2,085,159 1,661,140 207,356 341,079
0 297,880 0 29,622 0
518,883 2,212,214 2,212,947 210,661 206,572
345,922 1,290,458 948,406 122,886 206,572
0 184,351 0 17,555 0
106,140 205 10,266
2,304,757 4,614 165,397 28,592,930
1,676,187 3,691 23,628 27,169,178
209,523 923 47,256 1,661,972
1,385,379 2,652 119,599 19,830,964
1,007,549 2,121 17,086 22,270,539
125,944 530 34,171 1,428,515
745,766 1,309 91,807 16,897,696
542,375 1,048 13,115 27,553,897
67,797 262 26,231 1,888,806
30,781,815
27,244,018
15,556,639
22,210,989
22,365,843
18,123,594
20,976,337
27,772,048
36,487,648
Including long-beaked common dolphin.
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Cetaceans total
32,639 3,812 4,967 13,423
Prey Consumption and Competition by Marine Cetaceans
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False killer whale Pygmy killer whale Melon-headed whale Rough-toothed dolphin Pacific white-sided dolphin Risso’s dolphin Bottlenose dolphin Pantropical spotted dolphin Spinner dolphin Striped dolphin Common dolphin1 Fraser’s dolphin Northern right whale dolphin Dall’s porpoise Finless porpoise White whale Toothed whales total
157
North Atlantic. Method 1
Species
Biomass (t)
Fish
Method 2
Cephalopods Crustaceans
Fish
Method 3
Cephalopods Crustaceans
Fish
Cephalopods Crustaceans
33,903 2,629,407 67,244 978,334 322,325 9,353
0 328,604 8313 4,884,451 983,108 0
0 0 0 0 0 0
115,322 10,624,850 407,358 3,394,279 655,405 51,852
0 426,637 9429 4,981,637 1,192,285 0
0 0 0 0 0 0
160,487 13,794,588 462,029 3,461,815 794,857 61,045
0 1,007,720 17,181 7,401,910 2,470,620 0
0 0 0 0 0 0
433,114 32,582,954 841,861 5,143,700 1,647,080 119,487
13,984
0 6,204,476
0 0
70,580 15,319,646
0 6,609,988
0 0
85,807 18,820,629
0 10,897,430
0 0
178,648 40,946,844
Sperm whale Northern bottlenose whale Killer whale Long-finned pilot whale White beaked dolphin Atlantic white sided dolphin Harbour porpoise White whale Narwhal Toothed whales total
3,518,610 74,823
5,266,711 148,098
14,746,792 691,124
1,053,342 148,098
6,039,231 129,560
16,909,845 604,615
1,207,846 129,560
11,237,561 143,379
31,465,170 669,102
2,247,512 143,379
12,546 662,078 1,906 3,559
74,958 2,738,511 42,696 79,741
14,992 8,215,534 11,386 30,669
0 0 2,846 12,268
67,840 2,217,165 28,236 50,210
13,568 6,651,494 7,530 19,311
0 0 1,882 7,725
80,134 2,114,512 18,258 29,549
16,027 6,343,535 4,869 11,365
0 0 1,217 4,546
884 17,278 9,523
32,720 278,352 75,760 8,737,547
8,725 39,765 108,228 23,867,215
2,181 79,529 32,469 1,330,733
18,220 201,276 55,015 8,806,752
4,859 28,754 78,593 24,318,569
1,215 57,507 23,578 1,429,314
8,468 154,505 42,577 13,828,944
2,258 22,072 60,825 38,595,223
565 44,144 18,247 2,459,611
14,942,023
23,867,215
16,650,379
15,416,740
24,318,569
20,249,942
24,726,374
38,595,223
43,406,455
Cetaceans total
T. Tamura
Blue whale Fin whale Sei whale Common minke whale Humpback whale North Atlantic right whale Bowhead whale Baleen whales total
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4–3.
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Total. Method 1
Region
Fish
Method 2
Cephalopods Crustaceans
Fish
Method 3
Cephalopods Crustaceans
Fish
Cephalopods Crustaceans
Southern hemisphere including Indian Ocean Baleen whales 4,331,288 0 Toothed whales 13,395,256 27,360,029 Cetacean total 17,726,544 27,360,029
71,879,805 3,261,731 75,141,536
5,745,205 13,993,455 19,738,660
0 30,450,096 30,450,096
94,188,600 3,300,046 97,488,647
8,933,662 23,405,967 32,339,628
0 56,098,771 56,098,771
148,115,601 5,210,553 153,326,154
North Pacific Baleen whales Toothed whales Cetacean total
2,188,885 28,592,930 30,781,815
74,841 27,169,178 27,244,018
13,894,667 1,661,972 15,556,639
2,380,025 19,830,964 22,210,989
95,303 22,270,539 22,365,843
16,695,079 1,428,515 18,123,594
4,078,641 16,897,696 20,976,337
218,151 27,553,897 27,772,048
34,598,842 1,888,806 36,487,648
North Atlantic Baleen whales Toothed whales Cetacean total
6,204,476 8,737,547 14,942,023
0 23,867,215 23,867,215
15,319,646 1,330,733 16,650,379
6,609,988 8,806,752 15,416,740
0 24,318,569 24,318,569
18,820,629 1,429,314 20,249,942
10,897,430 13,828,944 24,726,374
0 38,595,223 38,595,223
40,946,844 2,459,611 43,406,455
Total Baleen whales Toothed whales Cetacean total
12,724,649 50,725,733 63,450,382
74,841 78,396,422 78,471,262
101,094,119 6,254,436 107,348,554
14,735,218 42,631,172 57,366,389
95,303 77,039,205 77,134,508
129,704,308 6,157,875 135,862,183
23,909,733 54,132,607 78,042,339
218,151 122,247,891 122,466,042
223,661,287 9,558,969 233,220,256
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4–4.
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Competition between whales and fisheries in the Northwest Pacific (Results of JARPN and JARPN II)
these fishing grounds; they feed mainly on Japanese anchovy.
Geographical and seasonal changes in dominant prey species in forestomach contents From the 1994– 1999 results of JARPN, Japanese anchovy (Engraulis japonicus) was the most important prey species in May and June on the Pacific side, while Pacific saury (Cololabis saira) was the most important one in July and August in subareas 7, 8 and 9. However, in JARPN II, in subarea 7 during August and September, walleye pollock (Theragra chalcogramma) was the most important prey species. Furthermore, Japanese anchovy and Japanese common squid (Todarodes pacificus) were also important prey species. At the same time, Pacific saury consumed by common minke whales was low in proportion. In subarea 9, Japanese anchovy was the most important prey species in August in 2000 (Table 9.5).
Discussion Southern hemisphere including the Indian Ocean
COMMON MINKE WHALE
BRYDE’S WHALE In JARPN II, Japanese anchovy was most important prey species; krill were also important prey species in subarea 7 (Table 9.5). SPERM WHALE In JARPN II, deep-sea squid was the most important prey species in subarea 7 (Table 9.5).
Competition between common minke, Bryde’s whales and fisheries COMMON MINKE WHALE Tamura and Fujise (2002) showed the relationship between common minke whales and fishing grounds of Pacific saury in summer near the Pacific side of Hokkaido. Common minke whales feed mainly on Pacific saury in August and September (Fig. 9.3). BRYDE’S WHALE Figure 9.4 shows the fishing grounds of skipjack tuna (Katsuwonus pelamis) and the positions of Bryde’s whale sightings in subarea 7 in August. Most of the Bryde’s whales sightings occurred close to
The total fish consumption by cetaceans was 18–32 million t, equivalent to 66–120% of current commercial fisheries catch of fish. Northridge (1984) reported the distribution and prey species of cetaceans in SHIO, noting that the population and prey species of most cetaceans were unknown. Most baleen whales (blue whale (B. musculus), fin whale, sei whale (B. borealis), Antarctic minke whale and humpback whale (Megaptera novaeangliae)) feed mainly on krill (mainly E. superba), and their feeding grounds are in the Antarctic (Kawamura, 1980a, 1994). Nemoto (1959) categorized the feeding types of baleen whales in the southern hemisphere as follows:
• • • •
Euphausiid feeder: blue whale, fin whale, humpback whale, Antarctic minke whale and sei whale. Amphipod feeder: sei whale. Copepod feeder: southern right whale (Eubalaena australis) and sei whale. Fish feeder: Bryde’s whale.
Antarctic minke whale is considered to be an euphausiid feeder (Horwood, 1990; Ichii and Kato, 1991; Tamura, 1998). Northridge (1984) stated that these baleen whales, excluding Bryde’s whale, did not feed mainly on fish, and no interactions with fisheries were likely. Bryde’s whales tends to stay in warmer waters and are known to prey on some commercial fish species. Best (1967) reported that they consumed 47% fish and 53% euphausiids in South Africa. The fish prey included some commercial species, such as pilchard (Sardinops ocellata), anchovy (Engraulis capensis) and mackerel (Trachurus spp.). Best (1977) found that the inshore form of Bryde’s whale eats mainly fish, especially pilchard, anchovy and horse mackerel (T. capensis).
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Subarea 7 Species
Research periods
Prey species
Common minke whale
JARPN
Japanese anchovy Krill Walleye pollack Others Pacific saury Krill Walleye pollack Others Japanese anchovy Walleye pollack Japanese common squid Krill Pacific saury Japanese anchovy Krill Others Deep-sea squids Deep-sea fishes
May–June
July–September
JARPNII Aug–September
Bryde’s whale
JARPNII Aug–September
Sperm whale
JARPNII Aug–September
Subarea 8
FO (%) WC (%) Prey species 92.56 1.15 5.22 1.07 49.89 35.56 6.57 7.97 36.36 35.55 15.00
92.09 1.16 5.39 1.37 51.68 37.33 9.03 2.06 34.81 37.15 14.99
8.64 4.45 71.74 28.26 0.00 99.38 0.62
6.94 6.10 71.60 28.40 0.00 98.4 1.60
Japanese anchovy Pacific saury Krill Others Pacific saury Japanese anchovy Krill Others
Subarea 9
FO (%) WC (%) Prey species 91.80 4.97 2.86 0.37 85.21 10.50 2.14 2.14
92.67 2.21 2.86 2.33 82.43 8.59 4.43 4.55
Japanese anchovy Pacific saury Krill Others Pacific saury Japanese anchovy Krill Others Japanese anchovy Krill Mackerel Pacific saury
FO (%) WC (%) 81.63 12.63 3.29 2.46 76.36 8.41 9.67 5.55 92.42 5.83 0.42
81.53 12.79 3.17 2.56 76.30 9.27 7.59 6.83 89.18 7.77 1.48
1.25
1.42
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Table 9.5. Prey composition (% of frequency of occurrences (FO) and % of weight composition (WC)) of common minke and Bryde’s whales revealed by Japan’s Whale Research (JARPN and JARPN II) in the western North Pacific (1994–2000).
JARPN: Japanese Whale Research Program under Special Permit in the Western North Pacific (1994–1999). JARPN II: Japanese Whale Research Program under Special Permit in the Western North Pacific: Phase II (2000). FO: relative frequency of occurrence of each prey species. WC: the relative prey composition by weight of each prey species.
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Fig. 9.3. Relationship between common minke whale sightings and the fishing grounds of Pacific saury during summer in the Pacific side of Hokkaido (a part of subarea 7W) in 1996. The information on the fishing grounds was obtained from the telex series Nos 27–33 on fishing grounds off the Pacific coast of eastern Hokkaido by the Japanese Fisheries Information Center (JAFIC) (redrawn from Tamura and Fujise, 2000b). Pie charts show the prey composition of common minke whales in the Pacific side of Hokkaido in August and September in 1996.
Fig. 9.4. Relationship between Bryde’s whale sightings (triangles) and the fishing grounds of skipjack tuna (hatched areas) in the Northwest Pacific during summer 2000. The position of the sightings of Bryde’s whales and the locations of the commercial fishing grounds for skipjack tuna in August (2–29) based on the thermal distribution during 23–29 August (Telex No. 1697 on fishing grounds in the North Pacific from JAFIC). The later information was also obtained from the telex series Nos 1697–1699 of the JAFIC. Pie charts show the prey composition of Bryde’s whales in the Northwest Pacific (latitude: 36–41°N; longitude: 144–150°E) in 2000.
Kawamura (1980b) reported that they feed mainly on krill (e.g. Euphausia diomedeae, E. recurva and Thysanoessa gregaria) in the South Pacific and Indian Ocean. Northridge (1984)
stated that this species would seem to have some degree of competition with fisheries, but no conflict was evident at that time. For almost all small cetaceans, the population sizes and
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prey species are unknown. However, their main prey seemed to be pelagic and mesopelagic fish and squid. Hence, there is some anecdotal information of interactions between capture fisheries and cetaceans such as killer whale (Orcinus orca), false killer whale (Pseudorca crassidens) and common dolphin (Delphinus delphis) in the region. The sperm whale biomass is large (5630 thousand t), occupying 29.4% of the total biomass of cetaceans in the region. However, this value refers only to the Antarctic, while sperm whales are found in all oceans, from equatorial waters to polar regions. Their prey species spectrum is dominated by mesopelagic squids, for which there is no commercial fishery. However, they eat mainly fish in regions such as New Zealand (Kawakami, 1980). Even though the share of commercial fish is small in their prey consumption, the absolute quantity of fish consumption is very large because of their huge biomass. There may be some interactions between sperm whales and capture fisheries. Furthermore, if either the commercial fisheries for squid or bottom fish or the abundance of sperm whales expand in the future, there may be some interactions between sperm whales and squid or bottom-fish fisheries. In the Antarctic, baleen whales, excluding Bryde’s whales, feed mainly on krill during austral summer, where the krill fisheries decreased recently because of diminishing markets. Of these cetaceans, the Antarctic minke whales play a most important role in the food web in the Antarctic. Armstrong and Siegfried (1991) indicated that the Antarctic minke whales consume 95% of the total biomass of krill that is consumed by baleen whales in the Antarctic. This study showed that the annual crustacean consumption by Antarctic minke whales was 42–64 million t, and that this amounted to 40–54% of total annual crustacean consumption by cetaceans in the southern hemisphere. T. Tamura, T. Ichii and Y. Fujise (1997, unpublished data) estimated the prey consumption of krill by Antarctic minke whales around the Ross Sea in the Antarctic to be an order of magnitude greater than the estimated consumption by Adelie penguins (Phgoscelis adeliae) and crabeater seals (Lobodon carcinophagus). While
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there may have been direct competition for krill among cetaceans, seals and seabirds in austral summer in the Antarctic, krill fishery appears to be of minor importance now. However, any development of this fishery could lead to increased competition between cetaceans and fisheries. For better understanding of this phenomenon, it will be necessary to monitor the abundance of cetaceans and obtain quantitative information on prey species to assess the interaction between fisheries and cetaceans. This applies particularly to the Indian Ocean, where there is no available abundance information for cetaceans.
North Pacific The total fish consumption by cetaceans was estimated at 21–31 million t, equivalent to 67–99% of commercial fish catches in recent years. In contrast to the southern hemisphere, many species of baleen whales feed on various pelagic prey species of zooplankton, squid and fish (Kawamura, 1980a). Nemoto (1959) categorized the feeding types of baleen whales in the North Pacific as:
• • • •
Euphausiid feeders: blue whale, fin whale, Bryde’s whale, humpback whale and common minke whale. Copepod feeders: North Pacific right whale (Eubalaena japonica), sei whale and fin whale. Fish feeders: fin whale, Bryde’s whale, humpback whale and common minke whale. Squid feeders: fin whale and sei whale.
North Pacific right whales and bowhead whales (Balaena mysticetus) feed on small copepods (e.g. Calanus glacialis and C. hyperboreus), and blue whales feed on krill (e.g. Euphausia pacifica, Thysanoessa inermis and T. longipes). Fin whales feed on many kinds of fish, mostly small, schooling fish such as Japanese anchovy, Pacific saury, chub mackerel (Scomber japonicus), Pacific herring (Clupea pallasii) and walleye pollock, and they also eat a variety of pelagic zooplankton and even some squid, such as Japanese common squid. Their prey species overlap with some
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commercial species, and their biomass is larger than other baleen whales, so there may be interactions to some extent in the North Pacific. Humpback whales also feed on many kinds of fish, mostly small, schooling fish such as capelin (Mallotus villosus), chum salmon (Oncorhynchus keta), sand lance (Ammodites hexapterus and A. personatus), Pacific herring and walleye pollock, as well as a variety of pelagic krill (e.g. E. pacifica) (Nemoto, 1959; Kawamura, 1980a). Northridge (1984) stated that no interactions with fisheries were apparent. However, their prey species also overlap with some commercial catch, and thus there may be interactions to some extent in the North Pacific. Sei whales feed on copepods (Calanus spp.), but they also feed on some small, schooling fish such as Japanese anchovy, Japanese pilchard (Sardinops melanostictus), Pacific saury and Japanese common squid. Bryde’s whales feed on krill, but they also feed on some small, schooling fish such as Japanese anchovy and Japanese pilchard (Nemoto, 1959; Kawamura, 1980a). There is at least one report of a Bryde’s whale that had been feeding upon penaeid shrimp in the South China Sea (Persons et al., 1999). Prey species of sei whales and Bryde’s whales also varied both geographically and temporally in the North Pacific. Northridge (1984) stated that there appears to be no reported conflicts with fisheries. However, their prey species overlap with some commercial fisheries catch, so there may be an interaction to some extent in the North Pacific. In JARPN II, most of the Bryde’s whales sightings occurred close to these fishing grounds; they feed mainly on Japanese anchovy but do not feed on skipjack tuna. The skipjack tuna is reported to feed on the Japanese anchovy (Kawasaki, 1965). There appears to be indirect competition between Bryde’s whales and skipjack tuna in summer in the western North Pacific. Common minke whales feed on various pelagic prey species of zooplankton, squid and fish (Nemoto, 1959; Kawamura, 1980a; Tamura and Fujise, 2002). Prey species varied both geographically and temporally. For example, Kasamatsu and Tanaka (1992)
reported that the prey composition of common minke whales caught off the Sanriku–Hokkaido area changed greatly between 1965 and 1987. They suggested that chub mackerel was the dominant prey species in 1968–1976, but Japanese pilchard was then the dominant prey species after 1977. Tamura et al. (1998) examined in detail the stomach contents of common minke whales caught under JARPN, and noted that dominant prey species have changed to Pacific saury and Japanese anchovy in recent years. The common minke whales feed on both pelagic zooplankton and pelagic schooling fish. Tamura and Fujise (2002) reported that most of the common minke whale sightings occurred close to Pacific saury fishing grounds; they feed mainly on Pacific saury (Fig. 9.3). T. Tamura and Y. Fujise (2000, unpublished data) estimated the seasonal consumption of Pacific saury by common minke whales along the Pacific side of Japan during August and September to be equivalent to 10–21% of the catch of Pacific saury in Japan. Northridge (1984) stated that there was some interaction between this species and fisheries in the North Pacific. There exists direct competition between common minke whales and commercial Pacific saury fisheries from summer to autumn in the western North Pacific. However, in JARPN II, Pacific saury consumed by common minke whales was low in proportion. The common minke whales feed mainly on walleye pollock and Japanese common squid, which are very important target species of fisheries in Japan. This seems to suggest a relationship between common minke whales and walleye pollock and Japanese common squid in summer in the western North Pacific, in addition to the relationship between common minke whale and Pacific saury fishery. Wade and Gerrodette (1993) and T. Miyashita (1990 and 1991, unpublished data, 1993a,b) estimated the population sizes of small cetaceans in the North Pacific, where the main prey species seemed to be fish and squids, both pelagic and mesopelagic. There are at least some reports of interactions between fish fisheries and cetaceans such as killer whale, false killer whale, Dall’s porpoise (Phocoenoides delli) and common dolphin in
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the region. Abundance estimate of small cetaceans and quantitative data of prey species are necessary to assess the interaction between fisheries and small cetaceans in the future. As for sperm whale, their biomass is large (2.3 million t), occupying 34.9% of total biomass of cetaceans. Their prey species is dominated by mesopelagic squid, for which there seems to be no commercial capture fishery. However, they eat some commercial pelagic fish, such as sardines, pink salmon (Onchorhynchus gorbuscha), Pacific saury and chub mackerel in the western North Pacific (Kawakami, 1980). Furthermore, Rice (1989) reported that sperm whales fed on black cod (Anoplopoma fimbria), or sablefish, from longlines being retrieved by fishermen in the eastern Gulf of Alaska. Although the fish quantity is a relatively small part of their consumption, it is conceivable that the total quantity of fish consumption is very large, because their biomass is so large. Therefore, there may be some interactions between sperm whales and fisheries. Furthermore, if either the commercial fisheries of squid or demersal fish intensify, or the abundance of sperm whales increases, such interactions may become more severe. In the Bering Sea, Lowry and Frost (1985) tried to clarify the biological situation by assessing potential interactions between marine mammals and commercial fisheries. They calculated ranked value based on diet composition, feeding strategy, importance, population size, etc. However, there was some question that competition with fisheries occurred, because of insufficient knowledge as to how marine mammals eat their prey (especially in relation to geographical, seasonal and yearly changes of prey species) and how the energy obtained from feeding relates to growth, maturation, reproduction and survival. Trites et al. (1997) tried to assess the degree of competition between fisheries and marine mammals in the Pacific Ocean (FAO areas 61, 67, 71, 77, 81, 87 and 88). They calculated the total annual prey consumption of marine mammals as 150 million t, equivalent to roughly three times the commercial fisheries catch. However, as the prey consisted primarily of mesopelagic squid and fish, they proposed that the most important consumers
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of fish and competitors of commercial fisheries were probably other predator fish, not marine mammals. However, they considered that there was indirect competition between fisheries and cetaceans in the Pacific Ocean. Furthermore, since the available primary production for sustaining fish and marine mammals is reduced, they suggested that the commercial fisheries could not continue to expand as previously. Hunt et al. (2000) reported on prey consumption by marine birds and mammals during summer in the PICES region (30°N to the Bering Strait). Their estimates of consumption exceed 2.5 million t by marine birds and 13 million t by marine mammals. However, as information on the abundance of marine birds and mammals was not included for all species, the actual figures of annual prey consumption by all marine birds and mammals are most probably larger than these results. They consider that the estimates must be re-examined in the near future.
North Atlantic The total fish consumption by cetaceans was estimated at 15–25 million t, equivalent to 87–144% of the commercial fish catch in recent years. In contrast to the southern hemisphere, many species of baleen whales feed on various pelagic prey species of zooplankton and fish similar to those in the North Pacific (Kawamura, 1980a). North Atlantic right whales (Eubalaena glacialis) and bowhead whales feed on small copepods, while blue whales feed on krill (e.g. Meganyctiphanes norvegica and Thysanoessa inermis). Northridge (1984) stated that these species did not feed on fish, and were unlikely to be affected by commercial fisheries. However, fin whales consume various pelagic prey species of zooplankton, squid and fish. They feed on many kinds of fish, such as capelin, sand lance, mackerel, herring, cod and lantern fish (Nemoto, 1959; Kawamura, 1980a). Prey species of fin whales vary both geographically and temporally in the North Atlantic. Perkins and Beamish (1979) reported that the fin whales feed
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mainly on capelin in Newfoundland. Northridge (1984) stated that there were no confirmed interactions with fisheries. However, there is the possibility of interactions with fisheries, because their biomass is larger than that of any other baleen whale. Sei whales feed on copepods (Calanus spp.), but they also feed on some small, schooling fish and squids in other regions. Bryde’s whales feed mainly on krill, but they also feed on some small, schooling fish in the North Pacific and southern hemisphere (Nemoto, 1959; Best, 1967, 1977; Kawamura, 1980a,b). Northridge (1984) stated that there appear to be no reported conflicts with fisheries. However, since sei whales consume some small, schooling fish in other regions, there exists the possibility of interaction with fisheries. Humpback whales feed on many species of fish, mostly small, schooling fish such as capelin, as well as a variety of pelagic krill (e.g. M. norvegica and T. inermis). Perkins and Beamish (1979) reported that the humpback whale feeds mainly on capelin in Newfoundland. Northridge (1984) stated that no interactions with fisheries were apparent. However, as their prey species overlap with some commercial catches, there may be an interaction to some extent in the North Atlantic. Common minke whales eat various pelagic prey species of zooplankton and fish in the North Atlantic. Prey species of common minke whales varied both geographically and temporally. In this region, feeding ecology of common minke whales has already been studied, and their diet varies according to season, geographical area and prey availability. Although krill are the important prey species, a wide range of fish species are preyed upon, among which capelin, herring and sand lance are predominant. In the North Sea, mackerel (Scomber scombrus) and sand lance are thought to be the dominant prey species. In the Northeast Atlantic and in the Barents Sea, a variety of prey is consumed, the most important of which are krill, capelin and herring, but gadoids, especially cod (Gadus morhua), saithe (Pollachius virens) and haddock (Melanogrammus aeglefinus), are also significant prey items (Haug et al., 1995, 1996). In recent years, increased attention has been paid to
interactions between commercial fisheries and common minke whales in the North Atlantic. For example, consumption of Atlantic herring by common minke whales was estimated to be 633 000 t year−1 in a part of the Northeast Atlantic. This is more than half of the total Norwegian catch of herring (L.P. Folkow, T. Haug, K.T. Nilson and E.S. Norødy 1997, unpublished data). Furthermore, Schweder et al. (2000) calculated using a simulation model that the net loss to the herring and cod fishery is some 5 t of herring and cod due to direct and indirect effects from the catches of an extra common minke whale in the Barents Sea. Off Iceland, Sigurjónsson and Vikingsson (1997) and Stefansson et al. (1997) have studied interactions of common minke, humpback and fin whales with ongoing cod and capelin fisheries, and found significant impacts of the whale stocks on the fish yields. In Icelandic waters, common minke whales feed predominantly on fish, i.e. capelin, sand lance, cod and herring, but frequently on krill (M. norvegica and T. inermis) (Sigurjónsson et al., 2000). There seems to be clear evidence of direct competition between common minke whales and commercial fisheries in the North Atlantic. The population and prey species of almost all small cetaceans are unknown (Northridge, 1984). Their main prey species seem to be pelagic and mesopelagic fish and squids. However, as for the North Pacific, there are some interactions between fisheries and cetaceans such as killer whale, false killer whale, harbour porpoise (Phocoena phocoena), bottlenose dolphin (Tursiops truncatus), white beaked dolphin (Lagenorhynchus albirostris), white whale (Delphinapterus leucas) and common dolphin in the region. There is need for more abundance estimates of small cetaceans, as well as quantitative data of prey species, in order to assess the interaction between fisheries and small cetaceans in the future. Sperm whale have a large biomass (3.52 million t), occupying 42.9% of total biomass of cetaceans, similar to other regions. Their dominant prey species is mesopelagic squid, for which there is no commercial fishery, although it has been reported that they eat mainly fish in Iceland waters (Sigurjónsson
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and Víkingsson, 1997). There may be some interactions between sperm whales and fish fisheries. Furthermore, if either the commercial fisheries of squid or demersal fish expand or the population of sperm whales increases, there will be interactions between sperm whales and the respective fisheries. There seems to be evidence enough that there is direct competition between some cetaceans and commercial fisheries in the North Pacific and North Atlantic. However, it will be necessary to have more available abundance estimates of cetaceans and quantitative information on prey species to assess the interaction between fisheries and cetaceans. Furthermore, there is a need to understand the potential for cetaceans to have an impact on commercial fisheries, either directly (by consuming commercial species such as herring, Pacific saury and walleye pollock), or indirectly (by competing for prey resources) using simulation models for specific geographical regions. Growing concerns about the possible consequences of competition between marine mammals and fisheries make this an increasingly important issue in fisheries management and conservation in the future. For this purpose, comparative research on the seasonal, local and annual distribution and abundance of cetaceans and their prey should be extended. This should make it possible to develop a blanket, multi-species management plan for marine organisms that also involves marine mammals such as whales, dolphins, porpoises and pinnipeds, in order to allow a more realistic fisheries management strategy, aiming for both short- and long-term sustainability of marine organisms, including marine mammals and their conservation in the world.
References Armstrong, A.J. and Siegfried, W.R. (1991) Consumption of Antarctic krill by minke whales. Antarctic Science 3(1), 13–18. Best, P.B. (1967) Distribution and feeding habits of baleen whales off the Cape Province. Investigational Report, Division of Sea Fisheries, South Africa 57, 1–44.
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Best, P.B. (1977) Two allopatric forms of Bryde’s whale off South Africa. Report of the International Whaling Commission (Special Issue 1), 10–38. Buckland, S.T., Cattanach, K.L. and Hobbs, R.C. (1993) Abundance estimates of Pacific whitesided dolphin, northern right whale dolphin, Dall’s porpoise and northern fur seal in the North Pacific, 1987–1990. Bulletin of the International North Pacific Fishing Commission 53, 387–407. Calambokidis, J., Steiger, G.H., Straley, J.M., Quinn, T., Herman, L.M., Cerchio, S., Salden, D.R., Yamaguchi, M., Sato, F., Urban, J.R., Jacobsen, J., von Ziegesar, O., Balcomb, K.C., Gabriele, C.M., Dahlheim, M.E., Higashi, N., Ford, J.K.B., Miyamura, Y., de Guevara, P.L., Mizroch, S.A., Schlender, L. and Rasmussen, K.R. (1997) Abundance and population structure of humpback whales in the North Pacific basin. Final contr. report conducted by Cascadia Research Collective under Contr. 50ABNF500113 for NMFS Southwest Fisheries Science Center, La Jolla, California. Clark, A. (1980) The biochemical composition of krill, Euphausia superba Dana from South Georgia. Journal of Experimental Marine Biology and Ecology 43, 221–236. Dawson, S.M. and Slooten, E. (1988) Hector’s dolphin, Cephalorhynchus hectori: distribution and abundance. Report of the International Whaling Commission (Special Issue 9), 315–324. FAO (1997) Review of the state of world fishery resources: marine fisheries. FAO Fisheries Circular No. 920. FAO (1998) Fishery Statistics Capture Production. FAO Yearbook, Vol. 82. Gambell, R. (1976) World whale stocks. Mammal Review 6(1), 41–53. Haug, T., Gjøæter, H., Lindstrøm, U., Nilssen, K.T. and Røttingen, I. (1995) Spatial and temporal variation in northeast Atlantic minke whale Balaenoptera acutorostrata feeding habits. In: Blix, A.S., Walløe, L. and Ulltang, Ø. (eds) Whales, Seal, Fish and Man. Elsevier, Amsterdam, pp. 225–239. Haug, T., Lindstrøm, U., Nilssen, K.T., Røttingen, I. and Skaug, H.J. (1996) Diet and food availability for northeast Atalantic minke whales, Balaenoptera acutorostrata. Report of the International Whaling Commission 46, 371–382. Hinga, K.H. (1979) The prey requirements of whales in the Southern Hemisphere. Deep-Sea Research 26A, 569–577. Horwood, J. (1990) Biology and Exploitation of the Minke Whale. CRC Press, Boca Raton, Florida.
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Hunt, G.L., Kato, H. and Mckinnell, S.M. (2000) Predation by marine birds and mammals in the subarctic North Pacific Ocean. PICES Science Report 14, 1–165. Ichii, T. and Kato, H. (1991) Food and daily food consumption of southern minke whales in the Antarctic. Polar Biology 11, 479–487. Innes, S., Lavigne, D.M., Eagle, W.M. and Kovacs, K.M. (1986) Estimating feeding rates of marine mammals from heart mass to body mass ratios. Marine Mammal Science 2, 227–229. IWC (International Whaling Commission) (1979) Report of the Scientific Committee on protected species. Annex G, Appendix I. Report of the International Whaling Commission 29, 84–86. IWC (1980) Report of special meeting on Southern Hemisphere sei whales. Report of the International Whaling Commission 30, 493–511. IWC (1984) Report of the Scientific Committee, Annex E2. Report of the sub-committee on Northern Hemisphere minke whales. Report of the International Whaling Commission 34, 102–111. IWC (1986a) Report of the workshop on the status of right whales. Report of the International Whaling Commission (Special Issue 10), 1–34. IWC (1991a) Report of the Scientific Committee, Annex E. Report of the sub-committee on Southern Hemisphere minke whale. Report of the International Whaling Commission 41, 113–131. IWC (1991b) Report of the Scientific Committee, Annex F. Report of the sub-committee on North Atlantic minke whales. Report of the International Whaling Commission 41, 132–171. IWC (1992a) Report of the Scientific Committee, Annex E. Report of the Bowhead Whale Assessment Meeting. Report of the International Whaling Commission 42, 137–155. IWC (1992b) Report of the Scientific Committee, Annex F. Report of the sub-committee on North Pacific minke whale. Report of the International Whaling Commission 42, 156–177. IWC (1992c) Report of the Scientific Committee, Annex G. Report of the sub-committee on small cetaceans. Report of the International Whaling Commission 42, 178–233. IWC (1992d) Report of the comprehensive assessment special meeting on North Atlantic fin whales. Report of the International Whaling Commission 42, 595–606. IWC (1993) Report of the Scientific Committee, Annex G. Report of the sub-committee on small cetaceans. Report of the International Whaling Commission 43, 130–145. IWC (1994) Report of the Working Group on North Pacific minke whale management trials. Report
of the International Whaling Commission 44, 120–144. IWC (1995) Report of the Scientific Committee, Annex E. Report of the sub-committee on southern hemisphere baleen whales. Report of the International Whaling Commission 45, 120–141. IWC (1996) Report of the Scientific Committee, Annex E. Report of the sub-committee on Southern Hemisphere baleen whales. Report of the International Whaling Commission 46, 117–131. IWC (1997a) Report of the Scientific Committee. Report of the International Whaling Commission 47, 59–112. IWC (1997b) Report of the Scientific Committee, Annex G. Report of the sub-committee on North Pacific Bryde’s whales. Report of the International Whaling Commission 47, 163–168. IWC (2000) Report of the Scientific Committee, Annex G. Report of the sub-committee on the Comprehensive Assessment of Other Whale Stocks. Journal of Cetacean Research Management 2(suppl.): 167–208. IWC (2001) Appendix 3. Classification of the order Cetacea (whales, dolphins and porpoises). Journal of Cetacean Research Management 3, xi–xii. Kasamatsu, F. and Tanaka, S. (1992) Annual changes in prey species of minke whales taken off Japan 1948–87. Nippom Suisan Gakkaishi 58, 637–651. Kasamatsu, F. and Joyce, G. (1995) Current status of odontocetes in the Antarctic. Antarctic Science 7, 365–379. Kasuya, T. and Kureha, K. (1979) The population of finless porpoise in the Inland Sea of Japan. Scientific Reports of the Whales Research Institute 31, 1–44. Kato, H., Hiroyama, H., Fujise, Y. and Ono, K. (1989) Preliminary report of the 1987/88 Japanese feasibility study of the special permit proposal for Southern Hemisphere minke whales. Reports of the International Whaling Commission 39, 235–248. Kawakami, T. (1980) A review of sperm whale prey. Scientific Reports of the Whales Research Institute 32, 199–218. Kawamura, A. (1980a) A review of prey of Balaenopterid whales. Scientific Reports of the Whales Research Institute 32, 155–197. Kawamura, A. (1980b) Food habits of the Bryde’s whales taken in the South Pacific and Indian Oceans. Scientific Reports of the Whales Research Institute 32, 1–23. Kawamura, A. (1994) A review of baleen whale feeding in the Southern Ocean. Report of the International Whaling Commission 44, 261–271.
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Kawasaki, K. (1965) Katsuo no seitai to shigen I. Suisan kenkyu sousho 8–1. Suisan sigen hogokyoukai. (in Japanese). Klumov, S.K. (1963) Feeding and halminth fauna of whalebone whales (Mystacoceti). Trudy Instituta Okeanol 71, 94–194. Leatherwood, S., Kastelein, R.A. and Hammond, P.S. (1988) Estimate of numbers of Commerson’s dolphins in a portion of the northeastern Strait of Magellan, January–February 1984. Report of the International Whaling Commission (Special issue 9), 93–102. Lockyer, C. (1981) Growth and energy budgets of large baleen whales from the Southern Hemisphere. FAO Fisheries Series (5) (Mammals in the Seas) 3, 379–487. Lowry, L.F. and Frost, K.J. (1985) Biological interactions between marine mammals and commercial fisheries in the Bering Sea. In: Beddington, J.R., Beverton, R.J.H. and Lavigne, D.M. (eds) Marine Mammals and Fisheries. George Allen & Unwin, London, pp. 41–61. Miyashita, T. (1993a) Abundance of dolphin stocks in the western North Pacific taken by the Japanese drive fishery. Report of the International Whaling Commission 43, 417–437. Miyashita, T. (1993b) Distribution and abundance of some dolphins taken in the North Pacific driftnet fisheries. Bulletin of the International North Pacific Fishing Commission 53, 435–460. Mizroch, S.A., Rice, D.W. and Breiwick, J.M. (1984) The sei whale. Balaenoptera borealis. Marine Fisheries Review 46(4), 25–29. Nemoto, T. (1959) Prey of baleen whales with reference to whale movements. Scientific Reports of the Whales Research Institute 14, 149–290. Northridge, S.P. (1984) World review of interactions between marine mammals and fisheries. FAO Fisheries Technical Paper No. 251. Odell, D.K. (1992) Sperm whale (Physeter macrocephalus), family Physeteridae, order Cetacea. In: Humphrey, S.R. (ed.) Rare and Endangered Biota of Florida. Florida University Press Gainesville, Florida, pp. 168–175. Ohsumi, S. (1981) Further estimation of population sizes of Bryde’s whales in the South Pacific and Indian Ocean using sighting data. Report of the International Whaling Commission 31, 407–415. Pauly, D. Trites, A.W., Capuli, E. and Christensen, V. (1998) Diet composition and trophic levels of marine mammals. ICES Journal of Marine Science 55, 467–481. Perkins, J.S. and Beamish, P.C. (1979) Net entanglements of baleen whales in the Inshore fishery of Newfoundland. Journal of the Fisheries Research Board, Canada 36, 521–528.
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Persons, E.C.M., Chan, H.M. and Kinoshita, R. (1999) Trace metal and organochlorine concentrations in a pygmy Bryde’s whale (Balaenoptera edeni) from the South China Sea. Marine Pollution Bulletin 38, 51–55. Perry, S.L., DeMaster, D.P. and Silber, G.K. (1999) The great whales: History and status of six species listed as endangered under the U.S. endangered species act of 1973. Marine Fisheries Review (Special issue) 61(1), 1–74. Raftery, A.E. and Zeh, J.E. (1991) Bayes empirical Bayes estimation of bowhead whale population size based on the visual and acoustic census near Barrow, Alaska, in 1986 and 1988. Paper SC/43/PS8 presented to the IWC Scientific Committee, May 1991 (unpublished). Rice, D.W. (1989) Sperm whale – Physeter macrocephalus Linnaeus 1758. In: Ridgway, S.H. and Harrison, R. (eds) Handbook of Marine Mammals, Vol. 4. River Dolphins and the Larger Toothed Whale. Academic Press, London, pp. 177–233. Schweder, T., Hagen, G.S. and Hatlebakk, E. (2000) Direct and indirect effects of minke whale abundance on cod and herring fisheries: a scenario experiment for the Greater Barents Sea. In: Vikingsson, G.A. and Kapel, F.O. (eds) Minke Whales, Harp and Hooded Seals. NAMMCO Scientific Publications, Vol. 2, pp. 120–132. Sergeant, D.E. (1969) Feeding rates of cetacean. Fiskeridirektoratets Skrifter, Serie Havundersokelser 15, 246–258. Shimada, H. and Miyashita, T. (1997) Population abundance of the western North Pacific Bryde’s whale estimated from the sighting data collected from 1988 to 1996. Paper SC/49/NP4 presented to the IWC Scientific Committee, Sep. 1997 (unpublished). (Available from the author). Sigurjónsson, J. and Víkingsson, G.A. (1997) Seasonal abundance of and estimated prey consumption by cetaceans in Icelandic and adjacent waters. Journal of Northwest Atlantic Fish Science 22, 271–287. Sigurjónsson, J., Galeu, A. and Vikingsson, G.A. (2000) A note on stomach contents of minke whales (Balaenoptera acutorostrata) in Icelandic waters. In: Vikingsson, G.A. and Kapel, F.O. (eds) Minke Whales, Harp and Hooded Seals. NAMMCO Scientific Publications, Vol. 2, pp. 82–88. Smith, T.D., Allen, J., Clapham, P.J., Hammond, P.S., Katona, S., Larsen, F., Lien, J., Mattila, D., Palsbøll, P.J., Sigurjónsson, J., Stevick, P.T. and Øien, N. (1999) An ocean-basin-wide mark–recapture study of the North Atlantic
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humpback whale (Megaptera novaeangliae). Marine Mammal Science 15, 1–32. Stefansson, G., Sigurjónsson, J. and Vikingsson, G.A. (1997) On dynamic interactions between some fish resources and cetaceans off Iceland based on a simulation model. Journal of Northwest Atlantic Fish Science 22, 357–370. Steimle, F.W. and Terranova, R.J. (1985) Energy equivalents of marine organisms from the continental shelf of the temperate Northwest Atlantic. Journal of Northwest Atlantic Fish Science 6, 117–124. Tamura, T. (1998) The study of feeding ecology of minke whales in the Northwest Pacific and the Antarctic. PhD Thesis. Hokkaido University. Tamura, T. and Fujise, Y. (2002) Geographical and seasonal changes of the prey species of minke whale in the Northwestern Pacific. ICES Journal of Marine Science 56, 516–528. Tamura, T., Fujise, Y. and Shimazaki, K. (1998) Diet of minke whales Balaenoptera acutorostrata in the northwestern part of the North Pacific in the summer, 1994 and 1995. Fisheries Science 64, 71–76.
Tillman, M.F. (1977) Estimates of population size for the North Pacific sei whale. Report of the International Whaling Commission (Special Issue 1), 98–106. Trites, A.W., Christensen, V. and Pauly, D. (1997) Competition between fisheries and marine mammals for prey and primary production in the Pacific Ocean. Journal of Northwest Atlantic Fish Science 22, 173–187. Trites, A.W. and Pauly, D. (1998) Estimating mean body masses of marine mammals from maximum body length. Canadian Journal of Zoology 76, 886–896. Wade, P.R. and Gerrodette, T. (1993) Estimates of cetacean abundance and distribution in the eastern tropical Pacific. Report of the International Whaling Commission 43, 477–493. Zeh, J.E., Clark, C.W., George, J.C., Withrow, D., Carroll, G.M. and Koski, W.R. (1993) Current population size and dynamics. In: Burns, J.J., Montague, J.J. and Cowles, C.J. (eds) The Bowhead Whale. The Society for Marine Mammalogy, Lawrence, Special Publication No. 2, pp. 409–489.
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Multi-species and Ecosystem Models in a Management Context Gunnar Stefansson University of Iceland, Science Institute, Reykjavik, Iceland
Abstract The final decades of the 20th century saw the emergence and first applications of multi-species models of marine ecosystems along with a general recognition of the potential importance of taking into account multi-species interactions when managing fisheries. Multi-species effects can include biological and technical interactions. Technical interactions frequently are of concern, for example when discards of certain species are believed to be a consequence of the management system. Biological interactions may fundamentally change the perspective of how to utilize an ecosystem, since a fishery or a moratorium on a predator may completely change the survival of a prey and, conversely, fishing on a prey may affect the growth of a predator. Modern research on multi-species modelling is highly multidisciplinary in nature, drawing on expertise from fishery science, fish biology, ecology, hydrography, mathematics, statistics, economics, operations research and computer science. As the models become more detailed and complex, they are able to address more issues that are of concern to managers, but at the same time it becomes ever more difficult to interpret results. Fundamental issues are raised in the multi-species context, and particularly so when fishing is viewed in the light of the precautionary approach. Some multi-species research has indicated that heavier fishing with smaller mesh sizes may lead to more profits for the fishing industry, whereas most earlier single-species research has indicated that low fishing pressure, particularly on juveniles, would be beneficial for the resource and the fishery. Conclusions from other research have indicated that economic considerations such as maximum economic yield may not be applicable, and have failed to lead to sustained utilization, whereas the traditional view has been that long-term economic views generally will lead to sustainable use of the resources. This chapter seeks to resolve some of these apparent conflicts, drawing on the multi-disciplinary nature of fishery science. It is seen that almost all points of view lead to the conclusion that fishing with low fishing pressure is not only sustainable but in accordance with the precautionary approach. Further, almost all multi-species concerns strengthen the need for reduced fishing pressure. It is also argued that simple management measures such as quotas, effort control or areal closures alone may not suffice to maintain viable fisheries in multi-species ecosystems.
Introduction This chapter endeavours to describe recent developments in multi-species modelling
approaches and how the results from those developments will or are likely to affect management decisions. To this end, the sections following the Introduction describe common
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current single-species methods of assessment and prediction, together with a description of multi-species issues that must be taken into account. Multi-species effects tend to be classified into two different types: biological and technical interactions. The chapter discusses the importance of each of these effects, modelling approaches and how these effects affect the possible utilization of the resources. It turns out that quite a few important management questions can only be addressed through the use of complex models, which include several species, areas and fine temporal scales. Such questions include the effects of closed areas, multi-species effects of a moratorium on fishing for a predator, and so on. Finally, given the current state of many of the world’s commercial fisheries and problems recently found in management advice given in many regions, it seems clear that tools are needed to evaluate the ecosystems in a more comprehensive manner than previously. This is particularly important in light of recent observations, which imply that entire ecosystems have collapsed primarily (and quite commonly) due to overfishing (Jackson et al., 2001). Models that include several species and their interactions have existed for quite some time, starting with the Lotka–Volterra models and later the emergence of models that can incorporate very many species. The first true applications of multi-species models of marine ecosystems were, however, seen closer to the end of the 20th century (e.g. ICES, 1991). Since the multi-species models need to be disaggregated spatially, they must contain a migration component in addition to biological and technical interactions. These models are therefore much more complex internally than previous single-species models. This chapter describes some of these issues, together with a related problem, that of using difficult and complex data sets to estimate unknown parameters of the models. When combined with prediction, the modelling approach requires a conglomerate of expertise from a variety of subject areas. This multidisciplinary nature of modern research on multi-species models is detailed in a separate section of this chapter.
Some fundamental issues are raised in the multi-species context, and particularly so when fishing is viewed in the light of the precautionary approach. In fact, some conclusions appear to be in conflict and there is a need to resolve these conflicts in order to pave the way for reasonable management. Some of these conflicts can be resolved by drawing on the multidisciplinary nature of fishery science. Generally speaking, fishing with low fishing pressure is sustainable, economic and in accordance with the precautionary approach. Most multi-species considerations further strengthen the need for reduced fishing pressure. Along with the application of the models comes a general recognition of the potential importance of taking into account multispecies interactions when managing fisheries. Thus, some of the first applications by advisory bodies immediately implied that fundamental understanding of the effects of fisheries could, at least in principle, be seriously affected by multi-species considerations. One such fundamental issue in managing marine resources is the overall level of fishing mortality to be exerted on the fish stock(s). The decisions on overall levels of harvest need to be based on all aspects of knowledge, first biological, but no less economic and social. Decisions on sensible fishing pressures pave the way for what control systems can be implemented, since they must be designed to achieve pre-defined goals. It will be seen that simple management measures such as quotas, effort control or areal closures alone will not in general suffice to maintain viable fisheries in multi-species ecosystems. Rather, combinations of these measures are needed to safeguard against the various issues raised in the multi-species context.
The Single-species Models of Assessment and Prediction The most common single-species models include recruitment, growth, maturation and mortality due to fishing and natural causes. In their simplest forms (Beverton and Holt,
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1957), these models commonly assume constant natural mortality, constant growth, a constant maturation pattern by age and a constant fishing pattern by age. Analyses of the effects of fishing may use these assumptions in order to evaluate the likely development of a yearclass and its possible utilization. Even the simplest such analyses need to consider the effects of incorrect assumptions. These computations subsequently provide first indications of potential yield from the resource, but need some estimate of yearclass size as input. In order to obtain such typical yearclass sizes, some assessment of the resource is needed. Classical methods include those which track individual yearclasses, from the initial virtual population analysis VPA (Gulland, 1965) to more recent statistical methods that incorporate more appropriate statistical assumptions. Whatever the methods, the outcome will be some stock estimate, typically in terms of the historical number of fish by age and year, up through the last data year. In order to evaluate the effects of management policies, the effects of these need to be modelled. Usually this is done by predicting the stock forward in time (i.e. from the assessment year), under the given harvest policy. In order to undertake predictions, some stock–recruitment relationship also needs to be used. This can be estimated from data, assumed to be a constant or to be of some general form. It turns out that the stock– recruitment relationship is of crucial importance when estimating the sustainability of a harvest policy or probability of stock collapse. The natural next steps involve the addition of factors such as cannibalism and density-dependent growth, if any of these are believed to be important. These single-species techniques, still fairly simple, have been used extensively to estimate medium- and long-term consequences of management actions. Alternative approaches to assessments and predictions have been developed. These include aggregate models that require fewer data but also provide less output. Such simpler models may be very useful and even prove better than the more disaggregate models in some circumstances (Butterworth
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et al., 1990). It is clear, however, that overly simple models cannot answer any of the more complicated issues in multi-species research. The above model classes range from very simple static biomass production models, through dynamic total biomass models, to age-disaggregated dynamic stock production models. Unknown parameters (e.g. recruitment) in the last model class are usually estimated using statistical methods of fitting models to data (but see below). Although given different names, such as HITTERFITTER (see, for example, Punt and Butterworth, 1991) or ADAPT (Gavaris, 1990), resulting techniques are all of the general form of adapting an internal model to data. All these single-species assessment models are of the form of an internal black box, which simulates an ecosystem based on some parameters. Results from different parameter values are compared with data and the parameters are estimated by finding the best fit to the observed measurements. In order to estimate the various unknown parameters of the models, statistical methods are used. During most of the last century, this step was skimmed over by using simple assumptions (such as independent lognormal errors), though in rare cases these were augmented by using known statistical distributions believed to describe the sampling process better. Subsequent analyses indicate that these assumptions are far from correct and that the sources of variation in the measurements are sufficiently complex to warrant the development of completely new statistical distributions to describe the data sets. This was noted early on for abundance data (Pennington, 1983; Lo et al., 1992; Stefansson, 1996), but for other biological data, such as length distributions, simpler assumptions have been used (MacDonald and Pitcher, 1979), sometimes extended to multi-nominal distributions (Methot, 2000). Recent research has indicated that these various extensions still suffer from being highly inadequate descriptions of reality (Hrafnkelsson and Stefansson, 2001). Although these statistical issues may seem esoteric, the results of incorrect statistical assumptions can, unfortunately, have devastating effects on overall conclusions. This is best seen by considering the best currently
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applied single-species assessment methods, which can not only provide stock estimates but can also estimate uncertainty (Patterson et al., 2001). One way of describing the uncertainty is to provide intervals that describe probability. A statistical method will, for example, provide a biomass level below which it is highly unlikely that the true biomass can lie. In particular, such a 1% lower confidence bound is designed in such a fashion that the true biomass should only be below it in 1% of all assessments. The best illustration of the problem involved is that recent research has indicated that when the standard statistical assessment methods report a 1% lower bound on a biomass value, in reality the true probability of being below that value can easily be 30%, and the most commonly used assessment methods give a corresponding underestimate of uncertainty (Gavaris et al., 2000; Patterson et al., 2000; Restrepo et al., 2000). The result of the statistical issues above is that even if management has been aiming for a low fishing mortality, for example, to be 99% certain of the stock staying above a depletion level, the actual probability of falling below that critical point may have been 30% (Gavaris et al., 2000). For predictions, it is essential to take into account the high degree of uncertainty involved in predictions of the development of marine species. Most management bodies need to know not only immediate and future yields but also probability of stock collapse, interannual variation in yield, likely rebuilding time, etc. The fact that recent work in this area has indicated that previous estimates of uncertainty may have been severely underestimated and that new and sophisticated statistical methods are required raises serious issues of reliability of predictions in general. In spite of their problems, single-species models have provided guidance on methods for rational utilization of fish stocks. In a nutshell, general results from these models are the following:
•
A low fishing mortality will generally decrease the probability of stock collapse.
• • • •
Low-to-medium fishing mortalities will not usually lead to reduced harvests. High fishing mortalities may lead to reduced harvests. High fishing mortality may lead to stock collapse. Economic considerations tend to imply a need for even lower fishing mortality than implied by biological models alone.
These results may not be completely universal conclusions from all single-species models, but very nearly so. These resulting points of view will be termed the singlespecies basic premises as they have resulted in fundamentals used in fisheries management worldwide. This is not meant to imply that the single-species models are correct or provide an adequate description of the ecosystem, but merely that the tenet of a low fishing mortality remains applicable as more complex situations are analysed. Apparently, in many cases, the broad results of severe management actions have been predicted adequately using fairly simple models. Thus, for example, a number of stocks have shown reduced total mortality during reduced fishing pressure and have even regained earlier levels following such management action (e.g. herring; Jakobsson, 1980). This is not true in all cases, however, as in some cases mortality does not appear to decrease following a stock collapse, even if fishing is reduced to a moratorium. For stocks that have not recovered under moratoria, reports are available on apparent increases in natural mortality (Sinclair, 2001). In other cases, revised models have indicated that a change in natural mortality is not required in order to explain available data on stock collapse, but a better statistical model is needed (Myers and Cadigan, 1994).
Multi-species Effects Although there are several examples of fisheries which can be classified simply as cases of overfishing and the importance of reducing fishing intensity is clear, there are
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quite a number of instances where the questions raised are somewhat more complex than these. Simple examples of questions such as the predicted effects of a closed area or the effect of an increase in a predator stock on its prey involve a necessary deviation from the simple models. It is simply not enough, for many (if not most) purposes, to have simple eye-glasses. Rather, the analyses and interpretations commonly must take into account the fact that species do not live in isolation.
Biological interactions In order to model multi-species effects, it is necessary first to develop a list of effects that may be important. This is the difficult step in modelling, as the mathematical and statistical models will follow naturally once a conceptual biological model has been developed. Following the single-species models in the previous section, the next natural steps in model extensions involve the biological interactions between species. Typically, these interactions involve predation and the resulting primary effect of predation on the mortality of the prey (Helgason and Gislason, 1979). The second factor to be taken into account is the effect of the predation on the growth of the predator. Depending on the ecosystem and species combinations, one or both factors may be important (ICES, 1991; Stefansson et al., 1998). As for single-species models, approaches to the multi-species models vary from simple model extensions through holistic approaches where the main processes in the system are cast in a unified mathematical framework. Even in the holistic approach, however, there is considerable scope for choice, ranging from the very simple ECOPATH approach (Christenesen and Pauly, 1990), which starts as a simple equilibrium biomass flow model, through models such as MSVPA, which are dynamic and age disaggregated for all (or most) species. These first models are useful in determining the main multi-species effects in the
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systems. In particular, ECOPATH is designed to indicate whether most important players in the system are included, and MSVPA similarly will indicate the most important sources of predation mortality for each species. Although in principle spatial factors may be important even when just considering one species, these factors become crucial when biological interactions are considered. The reason for this is of course the question of spatial overlap between the predator and prey species, which has been demonstrated in many ecosystems to be highly variable, resulting in widely varying predation mortality (e.g. Bogstad and Tjelmeland, 1990; Bogstad et al., 1994). The decision to take spatial variation into account has several important consequences, the obvious one being the requirement for a more realistic and much more complex model. For example, migration typically depends on the maturity stage of the fish, thus further implying that the model must take into account the difference in behaviour between mature and immature fish. In variable ecosystems where some species are tightly coupled with a predator–prey relationship of considerable importance for both species, it therefore becomes important to incorporate fishing, predation mortality, maturation, migration and growth as dependent on consumption (Stefansson and Palsson, 1998). These models inevitably include a large number of parameters, values of which can only be estimated using statistical techniques. Although, in principle, standard statistical methods can be used, fisheries data are very difficult to handle, and highly specialized methods are required. Although it will not be known in advance how complex the models need to be, it is clear that testing the effects of complexity can only be done using highly detailed models. A possible conclusion from such model tests may be that the increased complexity is not needed, but this cannot be known in advance. Recent work has indicated that highly complex models can indeed be evaluated and tested using advanced statistical techniques (Helu et al., 2000).
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Technical interactions Technical interactions frequently are of concern, such as when discards of certain species are believed to be a consequence of the management system. Further, different components of a fleet may target different components of a stock, most notably some components of the fishery may target the spawning stock in a spawning area during spawning time, whereas another part of the fishery may target smaller juvenile fish. For this reason, when technical interactions are concerned, the need quickly arises to take into account fine temporal and spatial scales. In terms of modelling, once multi-species interactions have been included, few issues or added complexity tend to arise due to the inclusion of technical interactions. In contrast, if fleet behaviour is added as a model component, together with the fleet’s response to economic issues, then a considerably different emphasis may develop (Olafsson et al., 1991).
Model Complexity As mathematical models become more complex, there is increased potential for serious issues of confounding to appear. The net effects of such confounding can become quite serious. One of the simplest forms of such confounding appears in disputes over whether mortality is due mainly to fishing or to natural causes. These simple confounding issues and resulting debates can, in many cases, be easily resolved. In most fields of science, it is standard practice to use designed experiments to verify what models are incorrect. Contrary to popular belief, large-scale experiments are also common in fisheries, although rarely designed explicitly to answer specific questions. Examples of such experiments include complete closures of fisheries due to stock collapses or wars, and implementations of strict management measures based on model predictions. Such experiments have demonstrated repeatedly that the basic premises of single-species fish population
dynamics as mentioned earlier are fundamentally correct (Jakobsson and Stefansson, 1998), though there are noteworthy exceptions (Sinclair, 2001). It is of course exceedingly useful to have such experimental results that verify model predictions. In more complicated models, unfortunately the confounding between factors can be of such a nature that it is impossible, using current methods, to verify true relationships.
Multi-species Modelling Approaches Modern research on multi-species modelling is highly multi-disciplinary in nature, drawing on expertise from fishery science, fish biology, ecology, hydrography, mathematics, statistics, economics, operations research and computer science. Naturally, the more extensive the inclusion of such factors, the more complex the models. It must be noted, however, that some of the most important conclusions regarding fisheries and overfishing do not depend on complex models. In particular, some particularly simple techniques can be used to demonstrate serious overfishing. If the sole purpose of analysis is to find such effects, then there often is no need to go into excessively complex models. In some cases, relatively simple extensions to single-species models can be used to verify effects of individual multi-species interactions. Thus the effect of a predator or a prey species can sometimes be entered as a simple regression variable (Pope and Knights, 1982; Stefansson et al., 1998). When developing models of the highly complex type considered here, the first step needs to be to define the biological factors to be taken into account. Having done this, the next step involves defining the corresponding mathematical model of the processes involved, followed by implementing the models in a computer program using statistical estimation techniques. The statistical aspects of the models become even more important in the multispecies models than in the single-species case.
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This is due to the increase in the number of data sets that must be used. In the singlespecies case, these will be only a few data sets, but even there the weight given to each data set may be quite important. In the multispecies case, a combination of either incorrect weights or inappropriate statistical assumptions may completely invalidate the output from the models (Stefansson, 1998). Having obtained the basic framework, the most promising current direction appears to be to build models of increasing complexity by comparing them with data in a stepwise fashion (Helu et al., 2000). This does, however, require appropriate statistical assumptions. Some of the current modelling work attempts to address all of these issues (Anon., 2001). Economic considerations must be taken into account if it is of interest to compare different fishing strategies, since they may lead to a shift in catches from one species to another and thus the regimes can only be compared using costs and income rather than simple biological yield. When attempting to find optimal harvest strategies or only to compare different strategies, methods of operations research, including maximization of utility functions, or at least comparisons of utility, commonly are used (Danielsson et al., 1997). These issues are important and certainly may affect results in individual situations, but are unlikely to change the principle of low fishing mortality and are outside the scope of the present chapter.
The Precautionary Approach in the Multi-species Context Traditional economic analysis would imply that fishing should be in such a manner as to ensure maximum long-term profits (or, more generally, maximum utility). Depending on what factors are taken into account, this sometimes has been simplified to maximizing total yield of a species in the long term, leading to maximum sustainable yield (MSY), corresponding fishing mortality (FMSY) or other biological measures (e.g. F0.1), which do not take economic considerations directly into account but aim for effort
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slightly lower than that giving maximum yield, as would happen if a cost function were used. The fundamentals of these approaches have come under considerable fire in recent years, particularly due to the (near-) collapse of many fish stocks (Mangel et al., 1996). It is, however, clear that most of the major stock collapses have occurred due to a combination of several factors, one or more of which led to considerable overfishing. Thus, fishing from most currently collapsed stocks simply was not in accordance with MSY or any other similar criterion, so these case studies tend not to affect the MSY principle per se. This does not alleviate the problem, however. The fact remains that stocks collapse and do so even when official policy is to maintain moderate harvests from the stocks. The reasons are usually not a policy of overfishing or a policy of fishing over MSY (there are, of course, exceptions where management aims directly for high fishing mortality, but these will not be addressed here). Rather, the official policies tend to be of moderate fishing, but the problem becomes one of a failure to attain this goal. The question becomes how to revise policy in order to ensure that harvests are sustainable despite the considerable uncertainty involved both in the science and in the implementation. In particular, it would usually be quite adequate to maintain a policy of MSY as a target, if it could be ensured that this would rarely be exceeded. In order to suggest remedies, it is of some importance to recognize a few causes rather than just the symptoms. Direct and documented causes of stock collapse or problems (serious and unexpected declines) include the following: incorrect advice on stock status; fishing well over advised levels; and lack of advice on danger levels and multi-species or environmental effects. The precautionary approach, stated in its simplest form, implies that care needs to be taken to ensure that fishing is undertaken in a sustainable manner and that when uncertainty is present, this should be taken into account by reducing fishing mortality. In implementing the precautionary approach, reference points have been defined. Loosely, they are defined in order to set rules that
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satisfy the criterion that as long as fishing is within bounds defined by the reference points, fishing mortality will not exceed specified harvest rates. Now, considering the present framework, these things become a bit more complicated. The easiest example involves a prey species that has been reduced to a very low level. Overfishing a predator species may then re-instate the prey to previous levels much faster than any other measure, but this would clearly violate the precautionary approach as regards the predator. At present, there is no system in place to address issues such as this. In the multi-species context, it is possible, in principle, that heavier fishing with smaller mesh sizes may lead to more profits for the fishing industry, whereas most singlespecies research has indicated that low fishing pressure, particularly on juveniles, would be beneficial for the resource and the fishery. Examples where quite different results are obtained from multi-species research include the North Sea (ICES, 1991). It is indeed easy to envisage how this can happen, simply through a reduction in the abundance of a predator leading to an increase in prey species. The specific North Sea results were obtained from forward projections that included some multi-species interactions, notably predation mortality. At the same time, there is considerable information to the effect that heavy fishing of juveniles can dramatically increase the probability of stock collapse. It is not clear in any general sense what the net or overall effect of, for example, heavy fishing on a predator would be in the long term. Clear results, however, include the effect of heavy fishing quickly drawing stocks to stock collapse, and of no fishing, in which case species can survive for millions of years. It would seem, therefore, that very strong evidence indeed is required to conclude that high fishing mortality of juveniles is beneficial. In examples where economic concerns have been included, at least one case study exists where it was predicted that reducing fishing mortality on a predator would lead to more than a 50% reduction in catches of a prey (Danielsson et al., 1997). In that particular case, economic analyses indicated that it was none
the less beneficial to the fishing industry to accept those reductions since the predicted total profits more than outweighed the negative aspects. Interestingly, in this case, the prey species did indeed collapse subsequent to an increase in the predator biomass.
Management in the Multi-species Context Initially, the inclusion of multi-species interactions, technical interactions, advanced mathematical and statistical models leads to considerable obfuscation. Thus, it is no longer uniformly clear whether mesh sizes should be increased or reduced, or whether fishing pressure needs to be reduced or increased to obtain sustainable fishing mortality. Upon some reflection, however, it is clear that the emerging figure is not as muddy as might appear at first. It must be noted at the outset that decisions on utilization need to take note of the precautionary approach. This implies that any uncertainty needs to be interpreted in favour of reduced fishing pressure. Thus, the fact that some new issues and questions are raised has no effect at all on principles such as a need to maintain low fishing mortality. Until clearly understood, such issues and questions merely urge more caution than before. Only in the case when it has been shown clearly that the complex models are a demonstrably better description of the system and it has been demonstrated how species should be utilized under the new system, should they be used for management purposes. For example, multi-species models incorporating predation mortality by adults alone may indicate a need for increased fishing pressure but, if the models do not include potential recruitment failure, then their results cannot be used to draw a conclusion of increasing fishing mortality. In practice, most multi-species and technical interaction models lead to conclusions that further emphasize the need for low fishing mortality:
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Results that indicate that fluctuations in a predator species may have adverse effects on survival of a prey imply that
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fishing effort on the prey must be reduced even further than previously thought. Results that indicate that the growth of a predator is positively influenced by the growth of a prey will imply that more care needs to be exercised in the prey harvest than before. Estimates of uncertainty will tend to be higher (and better) since more factors are included than before, leading to more aversion from high fishing mortality.
Multi-species results of a different nature include:
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In a three-species system, reduced pressure on a top predator may adversely affect its immediate prey, an intermediate-level species. This species may have its own prey (or competitor) which will become successful due to the reduced predation (or competition) pressure. Examples of such systems appear to exist (Bogstad et al., 1992). Effects of predation on stock–recruit relationships of prey appear to be very complex, and the resulting effects on, for example, biological reference points are even more difficult to interpret, though initial results indicate that such reference points can be developed (Gislason, 1999). Some examples of multi-species results also exist where predicted effects of mesh changes contradict earlier singlespecies results (ICES, 1991). This is to be expected, and in particular these may alleviate some of the unlikely optimistic biomass predictions which result from applying a suite of single-species models (with low and fixed natural mortality) and summing species-specific biomass values.
Finally, there are instances where a species may suffer very high or total mortality after spawning (Vilhjalmsson, 1994). In these cases, there is considerable incentive to fish up the stock before natural mortality occurs. In a few cases, a holistic approach has been taken, i.e. considerations of the effect of prey biomass on predator growth have been taken into
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account. In such situations, it sometimes has been found that the lack of catches of the prey species is offset by an increase in the growth of the predator, even to the extent of matching the loss. In the spatially explicit multi-species context, these factors crystallize even further, since it is clear that some of the dying prey will provide food for the predator. There is, therefore, even less incentive to fish hard on the prey. The full results of such analyses depend, however, on the economic importance of the predator and prey. Such predator–prey price ratios may differ greatly from one ecosystem to another (e.g. the different price of anchovy compared with capelin). It is seen that there may indeed be examples where the inclusion of multi-species effects implies that fishing pressure should be increased in order to obtain higher yields and even to obtain a more stable or sustainable fishery. As these findings appear to be exceptions, what remains, however, is the need to demonstrate this in individual situations. It is therefore not a valid argument to point to these exceptions and argue that this justifies increases in fishing mortality. Such justification must be clearly demonstrated based on data and models for the given situation. The default methodology under the precautionary approach needs to be the prudent one of low fishing pressure since this appears to be the general situation, and lack of knowledge of interactions simply qualifies as any other reduced knowledge and implies a need for low fishing mortality. This conclusion is even more important in the light of results that imply that simple control rules that ensure low fishing mortality will perform well even in situations of considerable variation in the true biological parameters of the populations (Walters and Parma, 1996).
A Missing Component in the Models Notably absent from most, if not all, current single- and multi-species biological and economic models is the concept of maximum potential effort. In a system with some form of limited entry, this can be very different from the effort as intended by
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management. Basically, in a system with limited entry, there is a possibility of an enormous dormant effort. The inclusion of such a concept would immediately bring forward the following model components, which currently are not implemented in models of marine ecosystems:
depleted species cannot economically sustain individual fishing trips and a more abundant species justifies the trips and sailing time, but the catches of the depleted species can be taken at minimal additional costs during the trip. This only occurs as a consequence of the combination of multi-species issues and an oversized fleet.
1. A large dormant effort results in a constant political pressure to increase realized effort. A model to take this into account should place a probability of a political decision to increase fishing mortality over a sustainable threshold, simply due to political pressure. This applies to all control systems. 2. At any given point in time, an increase in total allowable catch (TAC) or effort allocation can always be realized, even if this is erroneous and leads to a major increase in fishing mortality. This applies not only to TAC and effort control systems but also to systems based on areal closures (for mobile species, the effect of an areal closure may thus be negated by a large fleet fishing in adjacent areas). 3. An estimation error towards a low TAC (or effort) is unlikely to be realized as a low death rate symmetric to an overestimate. In addition to the political pressure, this is also a result of high grading, discarding of the species that will occur with other species that have not been underestimated, and an unknown slippage mortality due to excessive fishing activity on other species or size groups. 4. For a small fleet size, the effect of quota variation and discordance among species is negligible since the individual vessels will not be able to catch species that are not abundant. Basically this is due to the maximum possible excess effort in a small fleet. Thus there is a built-in guard against overfishing simply in the fleet size. With excessive dormant effort, however, vessels will find ways to fill all quotas, resulting in all overpredictions directly realized in mortality and excessive slippage in other species. 5. An oversized fleet can subsidize the fishery of certain overfished species through the catch of a more abundant species. This can happen, for example, when the fishery for a
The potential effects of dormant capacity become particularly clear in light of multispecies (biological or technical) interactions, and the concept is important enough to warrant inclusion in multi-species models and to be addressed by management systems.
Control Systems in the Multi-species Context Management systems typically depend on one or more of quotas, effort control, areal closures or other technical measures, such as mesh size changes. Considerable scientific and empirical evidence has been provided for the performance of each of these systems. These control mechanisms can now be viewed in the light of knowledge gained from the development of the multi-species models. Some problems affect all of these system, most notably problems of discards or high grading and catchability variation. High grading can be a general problem, particularly at high catch rates. Regardless of the control system chosen, it may be economically viable for a vessel crew to decide to discard low-value fish for high-value catches under any limitation whatsoever. In particular, the limit put by the size of the hold in the vessel may be enough to warrant discards under high catch rates. It can be beneficial to the operations of a freezing trawler to discard an entire hold full of frozen fillets if they are of a low-value species, should the vessel find a spot with another species of high value. This potentially can occur under any control system (including free fishing). Environmental changes can result in considerable changes in catchability which, when not taken into account, will lead to incorrect predictions (Stefansson and Eiriksson, 1998).
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These variations have an effect on the uncertainty of estimates of stock sizes and thus on appropriate effort, size of areal closure or TAC, thus affecting all control systems. A few examples suffice to show that each of these systems, when implemented alone, suffers from deficiencies.
Failure of a well-designed quota system Quota (or TAC) systems are based on deciding an annual TAC for each species. A well-designed quota system is one where the catches taken are in accordance with the quota set, which again is according to some system that aims to provide sustainable catches for the species involved. In principle, a quota system should not need to include other issues such as effort control or fleet size regulation, since the primary issue of fishing mortality is addressed directly by setting the TAC to achieve a pre-specified goal. A quota system can thus, in principle, limit fishing mortality inflicted on a given species. Without any further limitations, however, a fleet can move its effort towards areas of high abundance of spawning fish or of high abundance of juveniles. Such an increased effort towards certain age groups can easily lead to very high fishing mortality on certain age groups. The following model is an example where an initial design of a quota system will fail badly through perturbations not covered by the quota system. Suppose the quotas are intended to be set so that the TAC appropriate for each species is according to a sustainable fishing mortality. Several of the following problems have been recorded with such a set up, whereas others are plausible explanations for existing situations. 1. Uncertainty in the estimate of the fishing mortality may give a considerable (e.g. 30%) overestimate of the desired quota of some species, resulting in an increase in fishing mortality from the target. These effects tend to become hangover effects for several years, exacerbating the situation (Rivard and Foy, 1987). The true uncertainty in the population estimates has only recently been investigated systematically, and found to be considerably
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larger than commonly estimated previously (Gavaris et al., 2000). 2. Re-allocation of effort between areas can change the fishing pattern for a given species so that the juveniles or some other component get twice the intended effort, leading to high probability of stock collapse (Rose, 1993). This problem cannot be addressed in models unless spatial effects are modelled directly. 3. A species whose stock size has been overestimated can get a quota which is so high that it is virtually impossible to catch, leading to serious difficulties in a fleet which searches for this target species but catches only other species whose quota has already been taken. The net result can be a serious discard problem. 4. A species whose stock size has been underestimated may get discarded since it appears much more frequently in the catches than predicted. 5. A species with very low tolerance to fishing can be overfished even when taken only (or mainly) as by-catch in a fishery for another species that is fished sustainably (Walker and Hislop, 1998). This is an important example of a multi-species effect not normally taken into account when a TAC system is designed. 6. The fishery for a prey species may well affect the growth of a predator in such a fashion that the total economic outcome of the combined fishery is worse than that of not fishing for the prey at all. This scenario has not been addressed in the precautionary approach, but seems to be plausible in the light of some case studies (e.g. Magnusson and Palsson, 1991; Danielsson et al., 1997). It is seen that, from a modelling and advisory point of view, there are problems involved in evaluating the effects of management actions in a TAC system, problems that are not addressed using the models in common use around the world. In order to evaluate these effects, new and more detailed models are needed, taking into account spatial effects, multi-species effects and different statistical design. It is thus seen that, from an implementational point of view, there are several issues that may not be addressed in any detail within a quota system. In particular, a quota system based on TAC allocations for individual species may not lead to a
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sustainable fishery for all species involved, even the target species, unless of course the TAC for each given species is set far below each target.
Failure of a well-designed effort-control system An effort-control system is defined by some measures designed to limit the total effort that a fleet can exert. A well-designed system will attain the effort limitation intended, and the intended effort level corresponds to some sustainable fishing mortality for certain target species under a given scenario. The primary problem with this method is that the fleet is free to target its effort to any species, species group or species size class within the system. The total effort reduction typically will be set to be adequate to harvest the system according to a sustainable fishing mortality under a specified harvest regime. In most fisheries, this implies that the fleet has been fishing in several areas and on several species. The effort system should lead to sustainable use of the resource if there are no changes in how the fleet proportionally targets each part of the species complex. This design completely misses the multi-species viewpoints and spatial variation in species or age composition. The net effect of this omission can be arbitrarily devastating. In the simplest example, the fleet initially consists of two discrete components, each of which fishes for its own target species. In the typical scenario, the effort controls are designed to bring fishing mortality down to just below a collapse fishing mortality level, but in the best of worlds the target may be about half of the collapse mortality. In either case, if the price of one species increases sufficiently, the two fleets will both go for that species, leading to stock collapse. Given the first collapse, the fleets will target the second species. A price change is not even needed for this to happen. Natural variation in stock size usually will be sufficient for a change in fleet behaviour. Thus, if the size of one stock goes down sufficiently due to natural variation, the fleet will target another species.
Finally, in no known cases of effort limitations has any attempt been made to account for the increase in catchability inherent in most fleets. Examples are available of long-term catchability increases of 4.7% per year (Stefansson, 1998). In an effort-control system, the incentive for increasing efficiency is much greater than under free fishing or a quota system, and thus catchability increases could be considerably greater than this. In such a system, with limitations on the total number of fishing days, the number of days allowed per year would therefore probably need to be reduced by, for example, 10% per year every year simply to ensure that fishing mortality would not be guaranteed to increase steadily. Interestingly, this even happens for thesmallest vessel classes, sometimes termed artisanal vessels or owner-operated vessels, typically with 1–2 crew members. Thus, there are examples of owner-operated vessels with four computerized winches and GPS positioning equipment. No formal estimates of catchability exist in this case, but from total catch figures it is clear that it is possible to maintain considerable catches with such configurations. In some countries, these vessel classes tend to receive different treatment from the rest of the fishing fleet. However, the catches have exactly the same effect on fishing mortality, regardless of the political status of the fishery, and recent research even indicates that human interventions may have had considerable effects on marine populations for hundreds and even thousands of years (Jackson et al., 2001), further driving home the importance of taking into account the artisanal fisheries. It is seen that the usual single-area, singlespecies models of assessment do not take into account the likely variation due to species switching or spatial re-allocation within an effort-control system. Advice based on these models is therefore unlikely to capture much of the variation due to the system itself. In addition to the advisory problem, a pure effort-limitation system does not in general guarantee conservation of fish stocks in any sense. It is, however, clear that reducing
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effort to zero will work. It follows that the only way in which effort limitations will work is if the limitations are such that the fleet cannot induce high fishing mortality even with complete re-targeting of total potential effort, and the effort is reduced further every year to account for possible efficiency increase.
Failure of a well-designed areal closure Areal closures are designed to protect a certain collection of stock components. A well-defined areal closure succeeds in eliminating fishing from the area in question. Areal closures are sometimes temporary closures of small areas. These clearly will have little effect in a general overfishing situation. Similarly, closures that are only temporary (e.g. short seasonal closures) cannot provide any guarantee against overfishing, which can take place in other areas at other times. For example, common closures of spawning grounds during spawning time provide little protection for spawning fish since the spawning stock can be reduced to arbitrarily small levels through fishing on immature fish or on mature fish outside the spawning season. The net effect of fishing 50% of a yearclass before it matures is exactly the same as fishing 50% of the yearclass on the spawning grounds as the yearclass is preparing to spawn for the first time. Closures of major portions of the fishing grounds have, at the same time, apparently been seen to affect fishing mortality considerably (Murawski et al., 2000), and such measures have been found to affect the abundance in adjacent areas positively (McClanahan and Mangi, 2000; Roberts et al., 2001). Some existing closures of entire juvenile grounds are also likely to have an effect on the survival of juveniles (e.g. Vilhjalmsson, 1994) and thus on the survival of the stock (Myers and Mertz, 1998), but this does not seem to have been demonstrated through any evaluations of the effects of such closures. In fact, when interpreting evaluations of closures, it must be noted that many positive reports assume isolated adult stock components, either in modelling assumptions (Nowlis and Roberts, 1998) or in case studies that
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refer explicitly to reef fisheries (e.g. Bohnsack, 1998). In general, this may not hold, however. Suppose an areal closure is implemented in order to protect a given species. The simplest example where this will not suffice consists of a single species which has a migration pattern between certain areas, one of which is taken to be the closed area. The crucial factors in determining the effect of a permanent areal closure will be the rate of emigration from the area and the fishing mortality outside the area. This is because, if no other restrictions are implemented, then there is no intrinsic upper bound on the fishing mortality that can be implemented outside the closed area. Thus, the only upper bound on mortality due to fishing is simply the emigration rate. If the closed area is increased, the emigration rate is reduced and in the limit the areal closure will provide full protection. In some situations, knowledge may be available about the migration rate and it may also be possible to estimate fishing mortality with reasonable reliability. In these cases, it is, in principle, possible to estimate the effects of the areal closure. It is clear that such computations are essential if areal closures are to be generally useful as management tools. This implies a need for models that incorporate migration explicitly and provide estimates not only of the effect of fishing in the open areas but also of the associated uncertainty. In cases when such data are not available, it is very difficult to make any sensible statements except of a generalist nature as to the effect of areal closures. In particular, it is clear that there are many scenarios where unlimited fishing activity outside a closed area may lead to stock collapse. This will certainly be possible under several known migration patterns. Examples include such diverse species as tuna and cod. Thus it is seen that current single-area assessment models do little to predict the effects of an areal closure and, in the usual absence of scientific data on migration rates (including their uncertainty), it is not possible to provide advice on the net effect of these closures. Calling such areas sanctuaries does not in any way change the basic problem that
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the full effect of these on the population dynamics and sustainability is unknown and will, in some cases, be negligible. It follows that the only situation when there is any sort of guarantee that an areal closure suffices to provide sustainability for a stock is when the area is so large that most of the stock is protected.
Failures of other technical measures Other technical measures tend to be aimed at protecting certain age groups, length classes or maturity stages. Typically, these involve mesh size increases or other changes in fishing gear. If the technical measures are not combined in any way with overall fishing mortality limitations, then fleet development can continue to increase total fishing mortality without any specified upper limit. Thus, in general, the technical measures cannot be expected to guarantee a sustainable fishery. Notably, the results that high juvenile mortality may be linked with high adult (fishing) mortality in a documented case study (Myers et al., 1997) must be taken as a warning that hidden mortality can be considerable when total exerted effort is not limited. This hidden mortality may be due to slippage, discards or other (unknown) sources, but as long as it is linked to the total realized effort, the only way to reduce it is to reduce effort. The multi-species effects of, for instance, mesh size changes are quite contradictory. Some available research indicates that mesh increases thought to be beneficial in a single-species scenario may not lead to catch increases in a multi-species scenario. These results to date have not taken into account spatial variation in species composition, which in some cases is known to completely change the outcome of the models. The only instance when technical measures alone can be expected to provide a sustainable fishery is when they result in a complete termination of fishing on juvenile fish (Myers and Mertz, 1998). Even in this case, however, the mortality due to slipping through meshes is completely unknown,
and may be arbitrarily high unless some other measure is included to reduce total fishing mortality. Slipping mortality is not included in any standard assessment models. It can therefore be seen that these technical measures are unlikely to be sufficient in general to provide a sustainable fishery. It can be seen further that the state of the art in population dynamics models is unable to provide adequate advice on the (multi-species) effects of these technical measures.
Combined Control Systems It is seen that the usual control measures are not guaranteed to control fishing mortality and lead to sustainability. Some combinations are more likely than others to work, however.
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A combination of a TAC system with effort controls should reduce both the multi-species problem of re-allocation of effort between species (under effort-only control) and multi-species discard issue in the mis-specified TAC (in the TAC system). A combination of major closed areas for juveniles combined with a TAC system should reduce the (single-species) problems of fishing to unsustainable levels due to either re-allocation of effort to juveniles or overfishing in the open areas. A formal fleet reduction system in combination with any known control system will reduce all problems with every system. As with effort controls, however, a fleet reduction system is not enough on its own. An effort-control system along with large areal closures is much more likely to provide sustainable utilization than either system alone, since the combination can both provide a refuge and ensure that total effort is limited outside the closed area.
Any of these combinations would have to be designed in such a manner as to aim for
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an adequate definition of each component. Naturally there would be no use in adding an effort system to anything else, unless the effort control was designed truly to control effort to sustainable levels. The detailed implementation of such combined systems is outside the scope of the present chapter. It is, however, clear that such combinations are quite possible, though they may become somewhat complex. For example, an effort-control system can, in principle, easily be added onto a quota system with individually transferable quotas. Initially, this could be done by allocating each vessel its historical effort, subsequently allowing transfer of effort between vessels and reducing effort year by year sufficiently to guarantee more than compensation of efficiency increase. Naturally, the effort of large vessels needs to count more than the effort of small vessels in such a system, but the precise numbers are largely irrelevant in order to see some of the benefits of the combination. The only important issue in this case is to reduce effort enough annually to guarantee a true reduction in potential fishing mortality. As seen earlier, all of the systems are likely to fail in the case of dormant effort in the fleet. The usual exclusion of this effort from models leads directly to a bias in the predicted effects of all management action. Including multi-species and technical interactions in the prediction models may possibly alleviate the assessment problem somewhat but will not eliminate it completely. No current models are able to take into account the full potential effects of the overcapacity currently available in many of the world’s fisheries.
Conclusions This chapter has indicated the directions which current multi-species models have taken, how they have been developed in attempts to answer some of the questions raised by management and take into account various important biological issues. In terms
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of the utilization of resources, it is seen that the basic premises of classical single-species analyses hold in most instances, namely that maintaining low fishing pressure remains a prudent policy, is in accordance with the precautionary approach, is likely to provide sustainable catches, and will result in good yields in the long term. However, concerns raised in modern statistical, spatial and multi-species models indicate that the maintenance of low fishing pressure is much more difficult than previously believed. This is due to a combination of many factors, from management issues in the multi-species context through estimation problems due to biological and statistical issues raised in complex models of ecosystems. As a result, there is a much greater need to reduce fishing mortality further than ever considered previously. It is seen that when these more complex models are developed, several practical issues arise in the interpretation of results, as well as in the development of the models. There are at present fundamental unsolved issues in the very model definitions (not to mention implementation in real situations or predicting the effects of management action). Simply put: functioning holistic models are not yet available. The most important result from developing these models, however, is the potential to view the system as a whole and the fisheries as a whole. As these models are developed, it becomes obvious that species interactions, spatial patterns and technical interactions can have a devastating effect on predicted outcomes of traditional methods of fisheries management. In particular, it follows from the analyses above that the use of any of the common regulatory systems alone may not suffice to maintain viable fisheries in multi-species ecosystems. In order to facilitate sustainable use of the resources, it is highly likely that a combination of most, if not all, systems is needed, including formal fleet reduction mechanisms. The models required to illustrate the problems are sufficiently simple that a formal model evaluation is superfluous. Unfortunately, the reverse is not true, since to
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illustrate that certain combined measures will actually work requires much more attention to detail. Complex models are needed in order to evaluate the effects of complex regulatory measures. These models need extensive data, which in many cases is not available, such as data on consumption or migration rates. The lack of data is not an indication that the models are too complex, but rather that the effect of the management measures cannot be predicted. If management is to be in accordance with the precautionary approach and data are lacking, there is a need to implement control measures that will work in spite of the added uncertainty. Interestingly, the temptation by management not to take such conservative action often goes contrary to economically rational utilization, which would advocate (very) low fishing mortality. In cases where data are available, the models can be used to evaluate the effects of control measures. In such cases, it may be possible to reduce the size of a closed area or to demonstrate that a relaxation of the effort control will lead to greater catches without increasing the probability of overfishing. The models can also be used to evaluate the need for extensive data. Thus, the increased prediction accuracy obtained through more surveys or increased tagging can only be evaluated using corresponding models. The use of the models is therefore not only to advise management on control measures but also to advise on the data needed in order to be able to predict the effects of the measures. Finally, the basic tenets of single-species fish population dynamics probably need to be re-worded somewhat in the light of developments worldwide. In particular, rather than fishing at any (or the maximum) level that appears to be sustainable, it appears from the considerations in this chapter that an appropriate theme is: Marine resources should be harvested using the minimum fleet size possible, at that minimum level of fishing mortality that does not demonstrably lead to a serious long-term loss of catch.
References Anon. (2001) Development of structurally detailed statistically testable models of marine populations (dst2). Progress Report 1. MRI (Marine Research Institute, Reykjavik) Technical Report, No. 78. Beverton, R.J.H. and Holt, S.J. (1957) On the Dynamics of Exploited Fish Populations. Fascimile reprint, 1993. Chapman and Hall, London. Bogstad, B. and Tjelmeland, S. (1990) Estimation of Predation Mortalities on Capelin using a CodCapelin Model for the Barents Sea. Institute of Marine Research, Norway. Bogstad, B., Tjelmeland, S., Tjelda, T. and Ulltang, O. (1992) Description of a Multispecies Model for the Barents Sea (MULTSPEC) and a Study of its Sensitivity to Assumptions on Food Preferences and Stock Sizes of Minke Whales and Harp Seals. SC/44/O 9. Bogstad, B., Lilly, G.R., Mehl, S., Palsson, O.K. and Stefansson, G. (1994) Cannibalism and year-class strength in Atlantic cod (Gadus morhua L.) in Arcto-boreal ecosystems (Barents Sea, Iceland, and eastern Newfoundland). ICES Marine Science Symposium 198, 576–599, 1994–1021. Bohnsack, J.A. (1998) Application of marine reserves to reef fisheries management. Australian Journal of Ecology 23(3), 298–304. Butterworth, D.S., Hughes, G. and Strumpfer, F. (1990) VPA with ‘ad hoc’ tuning: implementation for disaggregated fleet data, variance estimation, and application to the Namibian stock of Cape horse mackerel (Trachurus trachurus capenis). South African Journal of Marine Science 9, 327–357. Christensen, V. and Pauly, D. (1990) The Ecopath II Model. ICES CM. 1990/L:68 Sess.Q. Danielsson, A., Stefansson, G., Baldursson, F. and Thorarinsson, K. (1997) Utilization of the Icelandic cod stock in a multispecies context. Marine Research Economics 12, 329–344. Gavaris, S. (1990) An Adaptive Framework for the Estimation of Population Size. Canadian Atlantic Fisheries Science Advisory Committee. CAFSAC Research Document 88/29. Gavaris, S., Patterson, K.R., Darby, C.D., Lewy, P., Mesnil, B., Punt, A.E., Cook, R.M., Kell, L.T., O’Brien, C.M., Restrepo, V.R., Skagen, D.W. and Stefansson, G. (2000) Comparison of Uncertainty Estimates in the Short Term Using Real Data. ICES CM 2000/V:03. Gislason, H. (1999) Single and Multi-species Reference Points for Baltic Fish Stock. ICES Working Group
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on Comprehensive Fishery Evaluations, 99/10. Gulland, J. (1965) Estimation of mortality rates. Annex to Rep. Arctic Fish. Working Group. Cons. Explor. Mer. C.M. 1965. Helgason, T. and Gislason, H. (1979) VPA-analysis with Species Interaction due to Predation, ICES CM. 1979/G:52. Helu, S.L., Sampson, D.B. and Yin, Y.S. (2000) Application of statistical model selection criteria to the Stock Synthesis assessment program. Canadian Journal of Fisheries and Aquatic Sciences 57(9), 1784–1793. Hrafnkelsson, B. and Stefansson, G. (2001) Likelihood functions for length distribution. In: MRI (Marine Research Institute, Reykjavik) Technical Report, No. 78. ICES (1991) Report of the Multispecies Assessment Working Group. ICES CM. Assess: 7. Jackson, J.B.C., Kirby, M.X., Berger, W.H. et al. (2001) Historical overfishing and the recent collapse of coastal ecosystems. Science 293 (5530), 629–638. Jakobsson, J. (1980) Exploitation of the Icelandic spring- and summer-spawning herring in relation to fisheries management, 1947–1977. Rapports et Proces-Verbaux 177, 23–42. Jakobsson, J. and Stefansson, G. (1998) Rational harvesting of the cod-capelin-shrimp complex in the Icelandic marine ecosystem. Fisheries Research 37, 7–21. Lo, N.C.-H., Jacobson, L.D. and Squire, J.L. (1991) Indices of relative abundance from fish spotter data based on delta-lognormal models. Canadian Journal of Fisheries and Aquatic Sciences 49(12), 2515–2526. MacDonald, P.D.M. and Pitcher, T.J. (1979) Age-groups from size-frequency data: a versatile and efficient method of analyzing distribution mixtures. Journal of the Fisheries Research Board, Canada 36, 987–1001. Magnusson, K.G. and Palsson, O.K. (1991) Predator–prey interactions of cod and capelin in Icelandic waters. ICES Marine Science Symposium 193, 153–170. Mangel, M. et al. (1996) Principles for the conservation of wild living resources. Ecological Applications 6(2), 338–362. McClanahan, T.R. and Mangi, S. (2000) Spillover of exploitable fishes from a marine park and its effect on the adjacent fishery. Ecological Applications 10(6), 1792–1805. Methot, R.D. (2000) Technical Description of the Stock Synthesis Assessment Program. US Department of Commerce, NOAA Technical Memo, NMFS-NWFSC-43.
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Murawski, S.A., Brown, R., Lai, H.L., Rago, P.J. and Hendrickson, L. (2000) Large-scale closed areas as a fishery-management tool in temperate marine systems: the Georges Bank experience. B. Marine Science 66(3), 775–798. Myers, R.A. and Cadigan, N.G. (1994) Was an increase in natural mortality responsible for the collapse of northern cod? Canadian Journal of Fisheries and Aquatic Sciences 52, 1274–1285. Myers, R.A. and Mertz, G. (1998) The limits of exploitation: a precautionary approach. Ecological Applications 8 (Supplement 1): S165–S169. Myers, R.A., Hutchings, J.A. and Barrowman, N.J. (1997) Why do fish stocks collapse? The example of cod in Atlantic Canada. Ecological Applications 7(1), 91–106. Nowlis, J.S. and Roberts, C.M. (1999) Fisheries benefits and optimal design of marine reserves. Fishery Bulletin 97(3), 604–616. Olafsson, S., Wallace, S.W. and Helgason, Th. (1991) Nordic Fisheries Management Model. Patterson, K.R., Cook, R.M., Darby, C.D., Gavaris, S., Restrepo, V.R., Punt, A.E., Mesnil, B., Skagen, D.W., Stefansson, G. and Smith, M. (2000) Validating Three Methods for Making Probability Statements in Fisheries Forecasts. ICES CM 2000/V:06. Patterson, K., Cook, R., Darby, C., Gavaris, S., Kell, L., Lewy, P., Mesnil, B., Punt, A., Restrepo, V., Skagen, D.W. and Stefansson, G. (2001) Estimating uncertainty in fish stock assessment and forecasting. Fish and Fisheries 2, 125–157. Pennington, M. (1983) Efficient estimators of abundance for fish and plankton surveys. Biometrics 39, 281–286. Pope, J.G. and Knights, B.J. (1982) Simple models of predation in multi-age multispecies fisheries for considering the estimation of fishing mortality and its effects. Canadian Special Publication in Fisheries and Aquatic Sciences 59, 64–69. Punt, A.E. and Butterworth, D.S. (1991) HITTERFITTER – bootstrap user’s guide version 2.0 (April 1991). SC/43/O 9. Restrepo, V.R., Patterson, K.R., Darby, C.D., Gavaris, S., Kell, L.T., Lewy, P., Mesnil, B., Punt, A.E., Cook, R.M., O’Brien, C.M., Skagen, D.W. and Stefansson, G. (2000) Do Different Methods Provide Accurate Probability Statements in the Short Term? ICES CM 2000/V:08. Roberts, C.M., Bohnsack, J.A., Gell, F., Hawkins, J.P. and Goodridge, R. (2001) Effects of marine reserves on adjacent fisheries. Science 294, 1920–1923.
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Rivard, D. and Foy, M.G. (1987) An analysis of errors in catch projections for Canadian Atlantic fish stocks. Canadian Journal of Fisheries and Aquatic Sciences 44, 967–981. Rose, G.A. (1993) Cod spawning on a migration highway in the north-west Atlantic. Nature 366, 458–461. Sinclair, A.F. (2001) Natural mortality of cod (Gadus morhua) in the Southern Gulf of St Lawrence. ICES Journal of Marine Science 58, 1–10. Stefansson, G. (1996) Analysis of groundfish survey data: combining the GLM and delta approaches. ICES Journal of Marine Science 53, 577–588. Stefansson, G. (1998) Comparing Different Information Sources in a Multispecies Context. Fishery Stock Assessment Models. Alaska Sea Grant College Program. AK-SG–98–01. 741–758. Stefansson, G. and Eiriksson, H. (1998) Assessment of Nephrops in Icelandic waters incorporating environmental factors. WP to EU funded SAP meeting in Barcelona, October, 1998.
Stefansson, G. and Palsson, O.K. (1998) Points of view. A framework for multispecies modelling of Arcto-boreal systems. Reviews in Fish Biology and Fisheries 8, 101–104. Stefansson, G., Skuladottir, U. and Steinarsson, B.A. (1998) Aspects of the ecology of a Boreal system. ICES Journal of Marine Science 55, 859–862. Vilhjalmsson, H. (1994) The Icelandic Capelin Stock: Capelin, Mallotus villosus (Müller), in the Iceland–Greenland–Jan Mayen area. Rit Fiskideildar 13, 1–281. Walker, P.A. and Hislop, J.R.G. (1998) Sensitive skates or resilient rays? Spatial and temporal shifts in ray species composition in the central and north-western North Sea between 1930 and the present day. ICES Journal of Marine Science 55, 392–402. Walters, C. and Parma, A.M. (1996) Fixed exploitation rate strategies for coping with effects of climate change. Canadian Journal of Fisheries and Aquatic Sciences 53, 148–158.
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11
Multiple Uses of Marine Ecosystems Andrew A. Rosenberg
College of Life Sciences and Agriculture, University of New Hampshire, Hampton, USA
Abstract The ocean is used for a very wide range of human activities, from recreation to food production to transportation. Each of these uses has the potential to affect fishery ecosystems and hence affect fisheries production, ecosystem health, stability and biodiversity. Here, impacts are categorized as direct, indirect or complex. Direct effects relate to changes in the mortality rate of ecosystem components. Indirect effects relate to changes in the productive capacity of ecosystem components, while complex effects are combinations of factors that change both mortality rates and productive capacity. Examples are given for each of these types of effects. Impacts also have three dimensions: spatial, temporal and complexity. To develop a coherent policy for addressing the impacts of multiple uses of marine ecosystems, it is important to consider how impacts occur in time and space, as well as how different factors inter-relate. A fourth dimension, detectability or quantifiability, could be added because many impacts of ocean use are very difficult to quantify. However, the lack of full scientific information on the magnitude of the impact of an activity in or on the ocean cannot be used as a reason for delaying policy action. The precautionary approach, widely included in international agreements on ocean management, applies and serves as a guide for policy making.
Introduction Incorporation of ecosystem considerations into fisheries management policy requires an understanding, at least conceptually, of how other concurrent ocean uses influence ecosystem properties. Ocean uses include disposal of contaminants, marine transportation, oil and gas exploitation, undersea mining for sand and gravel, cables for communication, eco-tourism, aquaculture, recreational activities, as well as fishing and conservation and preservation. The term fisheries ecosystem is used here to describe the biological, oceanographic and physical environment that supports exploited species within a specific area. The concept of
large marine ecosystems (LMEs) (AAAS, 1990) has been used in recent years to describe regional features of the marine environment in coastal areas. LMEs have distinct bottom topography, oceanographic features such as currents or water circulation, biological productivity and biodiversity, and are usually areas encompassing 200,000 km2 or more. The LME can extend from riverine and estuarine environments out into the coastal ocean, and even far offshore. Examples include the North Sea, the Gulf of Mexico, the Benguela Current, the Iceland Shelf, and so forth. Sherman (1994) identifies 49 LMEs globally and estimates that 95% of fishery yields comes from these LMEs. Here, the impacts of multiple uses on fisheries ecosystems can occur at the scale of LMEs or
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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can be more localized in scope. However, it is important to recognize that even a seemingly localized impact can have wider ranging effects. The LME concept is helpful for thinking of the linkages of biological, chemical and physical factors across large areas of the coastal ocean. Affecting any one part of the LME potentially can have repercussions throughout the region. The impacts of multiple uses of the ocean on fisheries ecosystems can be categorized as direct effects, indirect effects and complex effects. As with most categorizations, the distinctions between categories are imperfect, but provide a framework for thinking about potential impacts. In addition to thinking about the type of impact, three dimensions for impacts should be considered: spatial, temporal and complexity. These dimensions are important considerations for policy making with regard to the extent, duration and evaluation of any particular management action designed to address the impacts on fishery ecosystems of multiple uses of the ocean.
Direct, Indirect and Complex Effects A direct affect of a non-fishing activity on fisheries (or of a fishery on itself) occurs when that activity results in changes in mortality of
fish stocks in the ecosystem. In terms of the population dynamics of a fish stock, fewer young fish or recruits to the stock survive to spawn as mature fish. This is illustrated in Fig. 11.1, using a conventional picture of spawning stock and recruitment. As the slope of the line relating recruits to spawners increases, mortality of fish of pre-spawning age decreases. In other words, more young fish survive to reproduce. Conversely, if the slope of the line decreases, mortality increases and fewer young fish survive to reproduce. By and large, the dominant humancaused direct effect on fisheries ecosystems is fishing itself (Jackson et al., 2001). A large number of fisheries worldwide can be classified as overfished (FAO, 1999), meaning that the level of fishing activity in recent years is greater than the level that would give the maximum sustainable yield. Many of these fisheries are in fact exploited at a level that is not sustainable and stocks are continuing to be depleted. It is not uncommon in overfished fisheries for the exploitation rate to exceed 50%, i.e. 50% of the standing stock of the target species is removed each year. This is a huge increase in mortality, given that the exploitation rate that is likely to give the maximum sustainable yield for many stocks is of the order of 15–20% per annum and natural mortality is of a similar order or lower. This does not mean that the other types of impacts
Fig. 11.1. Simple population dynamics graph relating young fish or recruits to the number of mature fish or spawners they produce.
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described in this chapter are unimportant, but that overfishing, as a worldwide problem, must be prevented in every fishery ecosystem. By-catch is also a major problem for fisheries worldwide (Alverson et al., 1994) and is a direct effect of fishing on fisheries. By-catch can cause mortality of young stages of commercially important fish species. Increased mortality in young stages simply translates into fewer commercially valuable or mature fish some time later, a decreasing slope of the line in Fig. 11.1. There may be longer term effects of increased mortality on productivity, though these are considered here to be indirect. This illustrates an important point: any particular activity may have – is even likely to have – all three types of effects: direct, indirect and complex. Exploitation of one component of the ecosystem certainly can directly affect another. In the Gulf of Mexico, the USA shrimp fishery has a direct impact on the fishery for red snapper because of the by-catch of juvenile fish during the course of shrimping (NOAA, 1999). Fishing activities can also directly affect other, non-commercially important species both through by-catch, or the direct impacts of fishing gear on bottom or other organisms. In terms of the fishery, direct impacts on non-commercially valuable species may indirectly affect the commercially important stocks. Conservation and preservation efforts may affect fisheries directly by lowering the fishing mortality rate. In this case, the effect may be positive or negative with regard to fish yield. Reduced mortality for an overfished resource may allow rebuilding and increased yields. Conversely, efforts to protect a large part of an ecosystem may result in a smaller proportion of the commercially important fish stocks available for exploitation. In the groundfish fishery in the northeastern USA, a recovery programme put in place to rebuild overfished stocks of cod, haddock, flounders and other species reduced fishing mortality significantly in the late 1990s. In simple terms, this resulted in an increasing slope of the lines in the relationship shown in Fig. 11.1 for those fish stocks. The results of the conservation actions taken were initial reductions in yield from the fishery, followed by significant
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increases in yield, though recovery is not yet complete (Murawski et al., 1999). Chemical or nutrient contaminants may result in large-scale die-offs of marine life through their respective toxicity. This increased mortality of organisms can occur at several levels simultaneously, with a release of toxic substances, or the impacts may be different for different components of the ecosystem. For example, nutrient pollution can result in harmful algal blooms, such as brown tide (Laroche et al., 1997) and increased mortality of zooplankton and larval fish (Whitledge et al., 1999). This example illustrates that a non-toxic contaminant can result in direct effects as categorized here, namely increased mortality rates of organisms through secondary production of toxic substances. Of course there are also direct toxic impacts from contaminant releases such as oil spills (Rice et al., 1996). Finally, other competing uses, such as mining exploration and operation, transportation or aquaculture, may result in direct impacts on the ecosystem. Mining operations can disturb the bottom mechanically, killing organisms directly. Exploration using explosive charges produces shock waves that can kill even at some distance from the site. Transport vessels or aquaculture activities may introduce new species to an ecosystem that prey on indigenous ecosystem components (Carleton and Geller, 1993). All of these types of impacts are easy to describe, but often difficult to quantify. However, direct impacts are more likely to be quantifiable than indirect or complex effects. An indirect effect occurs when an activity results in changes in the productivity, from reproduction or somatic growth, of commercially important fish populations. In population dynamics terms, productivity changes due to reproduction can be depicted through the relationship between the number (or biomass) of spawners and the young fish or recruits that they produce. Figure 11.2 illustrates this, again using a conventional plot of spawners versus recruitment (note the axes are reversed from Fig. 11.1). As the height of the curve decreases, the spawning stock is less productive. In other words, a given number of spawning fish produce fewer progeny.
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Fig. 11.2. Simple population dynamics graph relating number of mature fish or spawners to the number of young fish or recruits they produce.
Productivity due to growth can also be affected by activities in the ecosystem. For example, if food resources are reduced through direct effects on prey species or competition with other organisms such as introduced species, fish may grow more slowly and the size of a spawning fish at the same age may be less. This, in turn, may result in fewer eggs or less viable eggs being produced by that mature fish than prior to the impact. For example, habitat destruction due to mining operations can result in reduced productivity because loss of habitat can reduce growth during young stages or reduce reproductive success. Climatic changes may reduce forage fish availability, thereby reducing productivity of commercially important species. This seems to have occurred in the Bering Sea with the reduction in stocks for herring and capelin due to a climatic regime shift (NRC, 1996). Conservation and preservation efforts may increase productivity through increased abundance of prey or increased availability of high quality habitat. The USA northeast fishery recovery efforts may again provide an example of this effect. One of the primary management measures used for this fishery was the establishment of large, year-round closed areas in 1995 that have remained closed for the last 6 years. These areas were established in primary habitat for many species on
George’s Bank, one of the most productive fishing grounds in the world. The recovery of commercially important stocks of fish and scallops is clearly due to reducing the direct impact of fishing mortality, but is also probably due to the indirect effect of habitat recovery (Murawski et al., 2000). Indirect effects of chronic contaminants are also an increasingly important and worrying phenomenon. Nutrient loads from the Mississippi River, emptying into the Gulf of Mexico, have resulted in the loss of a large amount of bottom habitat in the Gulf through the creation of a large hypoxic (low dissolved oxygen) zone of up to 18,000 km2 (Rabalais, et al., 1999). While mobile organisms can move away from the area, the habitat is no longer available to them, and productivity reductions are highly likely, though difficult to document. Other examples of indirect effects on fisheries ecosystems may arise through mining, transportation or aquaculture activities. Mining operations may chronically disturb habitat, reducing its productivity for some organisms; ship noise can disturb the habitat of animals such as whales that may be important components of a fishery ecosystem; and introduced species from ship ballast water or aquaculture can compete with indigenous species, reducing their productivity (Carleton and Geller, 1993). These effects are difficult to quantify, though it often is possible to determine that an impact is
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occurring, even if its magnitude is unknown (Rose, 2000). A complex effect occurs when the combined impacts of three or more factors affect the marine ecosystem upon which fisheries depend. To combine indirect and direct effects, again in simple population dynamics terms, we can overlay the graphs of productivity and mortality, as shown in Fig. 11.3. Now, because the axes must coincide, increasing slope of the straight lines representing the relationship between the number of recruits and the spawners they produce implies increasing direct mortality. Reduced height of the curve relating the number of spawners back to the number of recruits they produce implies reduced productivity. In combination, higher mortality and reduced productivity often result in reduced fishery yield over the long term, except for very lightly exploited fish stocks. For example, heavy fishing and habitat loss combine in a complex effect. Overfishing increases mortality and may reduce productivity, due to reduced spawning stock (e.g. Fig. 11.3), and a reduction in productive habitat due to fishing gear impacts, mining, contaminants or climatic changes exacerbates the loss in productivity. An important point here
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is that complex effects mean that factors interact such that it is not possible to conserve or recover ecosystem productivity simply by addressing one of the factors, even if that factor was responsible for the initial decline in production. For example, if climatic shifts such as those in the Bering Sea reduce the productivity of the ecosystem, fishing impacts must still be reduced to prevent further productivity losses, even if the initial decline was not due to overfishing. The plight of Steller sea lions in the North Pacific is a case in point. The western population of sea lions has declined over 80% over the last two decades due to a variety of factors, such as climatic regime shifts and incidental take by fishermen. Though fisheries may compete for prey with sea lions, it is unlikely that the initial large decline was due to fishing. However, recovery of the sea lion populations, initially affected by other factors, may now be hindered by localized intensive fishing for their prey species (Loughlin, 1998). Habitat loss and contaminants may combine to reduce the productivity of ecosystems. Habitat loss reduces the available areas for feeding, growth, spawning or nursery grounds, and contamination may reduce the suitability of the remaining habitat, even if
Fig. 11.3. Simple population dynamics graph relating number of mature fish or spawners to the number of young fish or recruits they produce (curved lines). The number of young fish are then related to the number of spawners they produce (straight lines). The latter are referred to as replacement lines because they indicate how the stock replaces itself from young fish to spawning fish.
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large-scale mortality does not occur. Such an effect occurs in nearshore or estuarine areas with the combined effects of filling of the estuary and runoff from developed land (Chesney et al., 2000). In this case, the complex effect is the combination of two indirect impacts. Similarly, aquaculture may cause habitat degradation and competitive interactions between farmed and wild fish, which in combination reduce the productivity of the ecosystem and hence fisheries. These categories are not exclusive, and the lines between them are somewhat blurred. Nevertheless, they provide a useful classification of competing uses of the ocean. However, there are only a few classes of interactions that can be quantified with regard to the extent of the effect. It is sometimes possible to quantify direct effects by estimating mortality rates over time. Less frequently, productivity can be quantified over time. Rarely are we able to understand enough about complex interactions to quantify their impact. The scientific challenge, then, is to improve our understanding, and, ultimately, our ability to enumerate and quantify the impacts of competing uses of the oceans.
Spatial, Temporal and Complexity Dimensions The impacts of multiple uses should also be considered in three dimensions in order to develop a coherent policy. For example, overfishing can be either localized, with depletion of resources in a small area but not necessarily throughout an LME, or region-wide. The short-term impacts of the direct effect of overfishing may be corrected by immediate conservation actions, but longer term impacts require more comprehensive planning for adjusting fishing capacity to the productive capacity of the resource. This is in part because it generally is true that fishing capacity increases more rapidly than it can be decreased for political reasons. Once the capacity exists, there is great pressure to keep that capacity fishing, even if long-term productivity decreases due to indirect or
complex effects as described above. For example, if overfishing is occurring in areas of degrading habitat, then the complexity of the impacts requires additional policy considerations that will adjust the fishing capacity of the fleet downward as the productivity of the resource declines. Alternatively, the habitat degradation must be reversed before the fishing pressure becomes too much for the resource to sustain. The spatial dimension is about more than just the scale of impacts on a fishery ecosystem. Distance from the coast is also a critical feature of many types of impacts, such as those due to contaminants. As a generalization, most contaminants are introduced near to shore and disperse and are diluted moving away from the source. Of course, an oil spill can occur offshore, but most oil production, and even shipping resulting in accidents, occurs relatively close to the coast. The types of factors resulting in impacts described in this chapter tend to be concentrated in nearshore areas. Therefore, the encounter rate with potentially impacting factors will increase. This is even true of most conservation and preservation efforts, though conservation in the Antarctic Southern Ocean is a counter example. Temporally, the distinction between short- and long-term impacts is straightforward, even if policy making for long-term impacts is not. This dimension may also be thought of as the gradient from acute to chronic effects. Direct effects are acute and often may be addressed with short-term policy decisions. Indirect effects are chronic by and large and require long-term solutions. Contaminant inputs such as nutrients from the Mississippi River result from a very large area, so the spatial dimension is large, but are chronic and require long-term policy actions. It can also be expected that the recovery of a fishery ecosystem from chronic impacts will take a long time. Even if nutrient loads in the Mississippi River were reduced quickly, the hypoxic zone in the Gulf of Mexico would probably persist for many years. The complexity of impacts cannot be ignored in policy development. Impacts
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on fisheries ecosystems are unlikely to be independent of one another in the sense that reversing the impact depends on addressing only a single factor. While overfishing may be a dominant direct effect in many ecosystems, other factors, such as habitat loss due to contamination or direct modification, also need to be considered. An illustration of ecosystem impacts of high complexity is the current situation with Pacific salmon in the USA Northwest (Lichatowich, 2001). Salmon have been subject to overfishing for many stocks, but their habitat has been very broadly affected. Now, fishing has been greatly reduced, but habitat for many life stages of the various salmon species is damaged or unavailable due to damming of rivers, water withdrawal for agriculture, contaminant impacts and other factors. Most salmon stocks in the Northwest are now considered threatened or endangered. The complexity of the impacts is very great and, consequently, the management actions needed are very complex, to say nothing of the political dimensions of the problems. The types and dimensions of effects are summarized, in a simplified way, in Table 11.1.
Policy Implications An additional dimension could be added: detectability or quantifiability. Direct effects are the easiest (though not necessarily easy) to quantify. Complex effects are hard to detect and very hard to quantify. This is not to say that they do not exist, or that there is no scientific evidence for them. Most importantly, the difficulty of quantifying
Table 11.1. complexity.
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effects is not a reason to ignore them in policy development. The policy challenge is to address these impacts in the absence of complete or quantitative information. Here, the precautionary approach to resource management can serve as a guide (Bodansky, 1991; FAO, 1995; Jordan and O’Riordan, 1999; Rosenberg, 2002). Fundamentally, fishery management policy should be cautious if a negative interaction is reasonably likely to occur, even if the extent of that interaction is unknown. In practice, this means restraining competing uses that may damage fisheries irrevocably, particularly in highly sensitive areas. For example, mining and drilling activities should be viewed very sceptically if they are proposed near or in areas of high fisheries production. Even if there is a lack of conclusive evidence that such activities are harmful, caution should be exercised, particularly when the potential risks to the ecosystem are high. In addition, if it is clear that an indirect effect has occurred or is unlikely to be avoided in future, fisheries should be restrained so that the, now reduced, productive capacity of the ecosystem is accounted for. If the productivity of a fish stock has been compromised because of, for example, habitat loss, that stock will be unable to withstand the same fishing pressure as before the loss. Regardless of whether the loss in productivity was due to fishing or other causes, it is important to reduce fishing pressure so as not to compound the error of habitat loss with the error of overfishing. Overall, competing uses of the oceans are likely to be complex from a management policy perspective. They are also likely to have a major, even dominant role in fisheries management in the near future.
Characterization of effects of various types over dimensions of space, time and Direct
Indirect
Complex
Spatial
Smaller, greater nearshore
Temporal Complexity
Short term, acute Low
Larger, can extend throughout LME Long term, chronic Higher
Variable, both local and LME-wide Short to long term Highest
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References AAAS (American Association for the Advancement of Science) (1990) Large Marine Ecosystems: Patterns Processes and Yields. AAAS Press, Washington, DC. Alverson, D.L., Freeberg, M.H., Murawski, S.A. and Pope, J.G. (1994) A global assessment of fisheries by-catch and discards. FAO Fisheries Technical Paper No. 339. Bodansky, D. (1991) Scientific uncertainty and the precautionary principle. Environment 33(7), 4–5, 43–44. Carleton, J. and Geller, J. (1993) Ecological roulette: the global transport of non-indigenous marine organisms. Science 261, 78–80. Chesney, E.J., Baltz, D.M. and Thomas, R.G. (2000) Louisiana estuarine and coastal fisheries and habitats: perspectives from a fish’s eye view. Ecological Application 10, 350–366. FAO (1995) Precautionary approach to capture fisheries and species introductions. FAO Fisheries Technical Paper No. 350 (Part 1). FAO (1999) The State of World Fisheries and Aquaculture. FAO, Rome. Jackson, J.B.C. et al. (2001) Historical overfishing and the recent collapse of coastal ecosystems. Science 293, 629–638. Jordan, A. and O’Riordan, T. (1999) The precautionary principle in contemporary environmental policy and politics. In: Raffensperger, C. and Tickner, J.A. (eds) Protecting Public Health and the Environment: Implementing the Precautionary Principle. Island Press, Washington, DC, pp. 15–35. Laroche, J.L., Nuzzi, R., Waters, R., Wyman, K., Falkowski, P.G. and Wallace, D.W.R. (1997) Brown Tide blooms in Long Island’s coastal waters linked to interannual variability in groundwater flow. Global Change Biology 3, 397–411. Lichatowich, J. (2001) Salmon Without Rivers. Island Press, Washington, DC. Loughlin, T.R. (1998) Steller sea lion: a declining species. Biosphere Conservation 1, 91–98. Murawski, S.A., Brown, R.W., Cadrin, S.X., Mayo, R.K., O’Brien, L., Overholtz, W.J. and
Sosebee, K.A. (1999) New England Groundfish. NOAA Technical Memo. NMFS-F/SPO–41, 71–80. Murawski, S.A., Brown, R., Lai, H.-L., Rago, P.J. and Hendrickson, L. (2000) Large-scale closed areas as a fishery management tool in temperate marine systems: the Georges Bank experience. Bulletin of Marine Science 66(3), 775–798. NOAA (National Oceanic and Atmospheric Administration) (1999) Our Living Oceans: Report on the Status of U.S. Living Marine Resources. NOAA Technical Memo. NMFS-F/SPO–41. NRC (National Research Council) (1996) The Bering Sea Ecosystem. National Academy Press, Washington, DC. Rabalais, N.N., Turner, R.E. and Wiseman, W.J., Jr. (1999) Hypoxia in the northern Gulf of Mexico: linkages with the Mississippi River. In: Kumpf, H., Steidinger, K. and Sherman, K. (eds) The Gulf of Mexico Large Marine Ecosystem: Assessment, Sustainability and Management. Blackwell, London, pp. 297–322. Rice, S.D., Spies, R.B., Wolfe, D.A. and Wright, B.A. (eds) (1996) Proceedings of the Exxon Valdez Oil Spill Symposium. American Fisheries Society Symposium 18. American Fisheries Society, Bethesda, Maryland. Rose, K.A. (2000) Why are quantitative relationships between environmental quality and fish populations so elusive? Ecological Applications 10, 367–385. Rosenberg, A.A. (2002) The precautionary approach from a manager’s perspective. Bulletin of Marine Science (in press). Sherman, K. (1994) Sustainability, biomass yields, and health of coastal ecosystems: an ecological perspective. Marine Ecology Progress Series 112, 277–301. Whitledge, T.E., Stockwell, D.A., Buskey, E.J., Dunton, K.C., Holt, G.J., Holt, S.A. and Montagna, P.A. (1999) Persistent brown tide bloom in Laguna Madre, Texas. In: Kumpf, H., Steidinger, K. and Sherman, K. (eds) The Gulf of Mexico Large Marine Ecosystem: Assessment, Sustainability and Management. Blackwell, London, pp. 338–359
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12
Impacts of Fishing Gear on Marine Benthic Habitats
Michel J. Kaiser1, Jeremy S. Collie2, Stephen J. Hall3, Simon Jennings4 and Ian R. Poiner5 1School
of Ocean Sciences, University of Wales-Bangor, Anglesey, UK; School of Oceanography, University of Rhode Island, Narragansett, USA; 3Australian Institute for Marine Science, Townsville, Australia; 4The Centre for Environment, Fisheries and Aquaculture Science, Lowestoft, UK; 5CSIRO Division of Marine Research, Cleveland, Australia 2Graduate
Abstract Fishing affects seabed habitats worldwide. However, these impacts are not uniform and are affected by the spatial and temporal distribution of fishing effort, and vary with the habitat type and environment in which they occur. Different fishing methodologies vary in the degree to which they affect the seabed. Towed bottom-fishing gears and hydraulic harvesting devices re-suspend the upper layers of the sedimentary habitat and hence re-mobilize contaminants and fine particulate matter into the water column. The ecological significance of these fishing effects has not yet been determined. Structurally complex habitats (e.g. sea-grass meadows, biogenic reefs) and those that are relatively undisturbed by natural perturbations (e.g. deep-water mud substrata) are more adversely affected by fishing than unconsolidated sediment habitats that occur in shallow coastal waters. Structurally complex and stable habitats also have the longest recovery trajectories in terms of the re-colonization of the habitat by the associated fauna. Comparative studies of areas of the sea bed that have experienced different levels of fishing activity demonstrate that chronic fishing disturbance leads to the removal of high-biomass species that are composed mostly of emergent seabed organisms. These organisms increase the topographic complexity of the seabed and have been shown to provide shelter for juvenile fishes, reducing their vulnerability to predation. Conversely, scavengers and small-bodied organisms, such as polychaete worms, dominate heavily fished areas. Such a change in habitat may lead to changes in the composition of the resident fish fauna. Fishing also has indirect effects on habitat through the removal of predators that control bio-engineering organisms such as algal-grazing urchins on coral reefs. However, such effects are only manifested in those systems in which the linkages between the main trophic levels are confined to less than ten species. Management regimes that aim to incorporate both fisheries and habitat conservation objectives can be achieved through the appropriate use of a number of approaches, including total and partial exclusion of towed bottom fishing gears, and seasonal and rotational closure techniques. Different management regimes can only be formulated and tested once objectives and criteria for seabed habitats have been defined.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Introduction Concerns regarding the effects of fishing on the marine environment have been restricted to the effects of harvesting on those species specifically sought by the fishermen and the by-catches of large fauna, such as cetaceans, birds, reptiles and non-target species of fish. However, in more recent times, there has been a growing appreciation of the wider and subtle effects that modern levels of fishing pressure exert on the marine ecosystem. These wider ecosystem effects of fishing activities on the marine environment have been projected to new prominence, through both scientific activity and media publicity, such that they are widely considered to be crucial considerations in any future management plans. The so-called ‘ecosystem approach’ to fisheries management was highlighted at a meeting held in Montpellier, France, in 1998 that focused exclusively on the ecosystem effects of fishing. This meeting, convened by the International Council for the Exploration of the Sea (ICES) and the Scientific Committee for Oceanographic Research (SCOR), attracted over 300 delegates and produced a symposium volume with 35 peerreviewed papers that provide an excellent reflection of the growing importance of this area of research (ICES Journal of Marine Science, Volume 57, no. 3). It is important to understand what we mean by the ‘ecosystem effects’ of fishing (Jennings and Kaiser, 1998). This can be summarized as follows. 1. The effects of fishing on predator–prey relationships, which can lead to shifts in community structure that do not revert to the original condition upon the cessation of fishing pressure (known as alternative stable states). 2. Fishing can alter the population size and body size composition of species, leading to a fauna composed of primarily small individual organisms (this can include the whole spectrum of organisms, from worms to whales). 3. Fishing can lead to genetic selection for different body and reproductive traits and can extirpate distinct local stocks. 4. Fishing can affect populations of nontarget species (e.g. cetaceans, birds, reptiles
and elasmobranch fishes) as a result of bycatches or ghost fishing. 5. Fishing can reduce habitat complexity and perturb seabed (benthic) communities. Points 1–4 are dealt with by other chapters in this volume; this chapter deals primarily with point 5, but should be considered in the context of the other effects of fishing activities. In recent years, the effects of towed bottom-fishing gear on benthic communities and habitats has received considerable media attention in both the trade and popular media press. This has been reflected by a rapid increase in the research effort that has addressed these issues (for reviews, see Dayton et al., 1995; Jennings and Kaiser, 1998; Kaiser, 1998; Watling and Norse, 1998; Auster and Langton, 1999; Kaiser and De Groot, 2000). In particular, Watling and Norse (1998) and Auster and Langton (1999) give lucid accounts of the ecological significance and ecosystem function of marine habitats and describe the processes that are likely to be affected by fishing activities.
Fishing as an Ecological Disturbance Physical disturbance The majority of seabed (demersal) fishing activity is undertaken in shallow seas on the continental shelf at depths of less than 100 m. However, deep-water fishing is an increasingly important sector of the industry and notably occurs around sea mounts at depths of more than 1000 m. Benthic communities within continental shelf environments experience continual disturbance at various scales (Hall, 1994). Large-scale natural disturbances, such as seasonal storms and regular (daily) scouring by tidal currents, form a background against which other smaller disturbances occur, such as those induced by predator feeding activities (Fig. 12.1). Even the small-scale disturbance effects of individual fauna may have a considerable additive effect on benthic communities, creating a long-term mosaic of patches in various states of climax or re-colonization (Grassle and Saunders, 1973; Connell, 1978). This may contribute to
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Fig. 12.1. The relative recovery rates of different scales of disturbance that occur in the marine environment. The figure shows various forms of fishing activity compared with naturally occurring disturbances such as predation effects and physical sources of disturbance (after Hall, 1994).
the inherent variability found within marine benthic systems, but can be obliterated by larger scale physical disturbances such as ice-scour or demersal fishing. It is important to consider the relative scale at which fishing disturbance occurs. Given a similar habitat, very intensive but highly localized fishing disturbance may have fewer ecological implications than less intense, but widespread, fishing disturbance. In any particular habitat, the associated fauna and flora presumably will have adaptations or life styles that enable them to persist in that environment. In other words, most communities have an in-built resilience to a certain level of physical disturbance. However, the scale and frequency of physical disturbance events may increase to a point where lasting ecological effects are observed, even against a background of natural disturbance. The additive effects of an entire fishing fleet may reach such a threshold. Shallow-water communities on exposed coastlines are likely to be the most resilient to physical disturbance from bottom fishing. For example, Posey et al. (1996) recently demonstrated that even largescale disturbances, such as hurricanes, have relatively short-term effects on shallow-water communities adapted to frequent physical disturbance (but see Rees et al., 1977). However, as habitat stability increases, the relative
effects of fishing will also increase, as will the longevity and severity of the ecological effects (Theil and Schriever, 1990; Kaiser and Spencer, 1996; Auster, 1998).
Species-dependent disturbance While it is possible to envisage how a bottomfishing gear that is towed over the seabed might lead to alterations in seabed habitat structure, it is perhaps less obvious how reductions in target and non-target fishes or other marine organisms might have consequences for seabed habitat. While the former is termed physical disturbance, the latter is an ecological disturbance that affects interactions that occur between different species. A classic example of species-dependent disturbance is the depletion of sea otters on the western seaboard of North America as a result of human harvesting. Sea otters predate sea urchins that eat kelp. The reduction in sea otter numbers led to an increase in sea urchin numbers, which increased the herbivore pressure on the kelp beds leading to their eventual decline (Estes and Duggins, 1995). Such a response is known as a trophic cascade and typically occurs in systems in which the linkages between different
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components of the system (otters, sea urchins and kelp) are very strong (Kaiser and Jennings, 2001). The interactions within the cascade are between organisms that are assigned to major trophic levels within a system, e.g. predator – herbivore – primary producer; very few cascades involve intermediate trophic levels. In such a case of a significant decrease in the population of a predator or herbivore, there are few others to take their place in the short term. This contrasts sharply with many of the more ‘open’ marine ecosystems, where there may be three or more predators that exert similar levels of predation on one or more species. For example, sharks, cetaceans and marine mammals all eat pilchard in the Benguela system off the eastern coast of Africa. Fishing reduces the numbers of large predatory fish, yet the loss of these high-level predators can have limited consequences for their prey species. In many of the species-rich marine ecosystems, there is a high degree of variance in size within the main phylogenetic groupings (containing genetically related species). This is particularly apparent on tropical reefs. Moreover, the phylogenetic groupings tend to contain a large number of species, with a wide range of life history traits, behavioural differences and feeding strategies (Jennings and Kaiser, 1998). As a result, predation is diffuse (Hixon, 1991) and, while the overall effect of all piscivores on their prey can be substantial, the impact of any one individual species, or small group of species, is minor.
Physical Disturbance of Habitats by Fishing Activities Towed bottom-fishing gears (trawls, dredges and drags, hydraulic devices) are used to catch those species that live in, on or in association with the seabed. Such gear is designed to catch bottom-dwelling species; hence, they are intended to remain in close contact in the seabed. The passage of this fishing gear over the seabed can be summarized as follows:
•
Disturbance of the upper layers of the seabed causing short-term re-suspension of sediments, re-mineralization of
• • •
nutrients and contaminants, and resorting of sediment particles. Direct removal, damage, displacement or death of a proportion of the animals and plants living in or on the seabed. A short-term attraction of carrion consumers into the path of the fishing gear. The alteration of habitat structure (e.g. flattening of wave forms, removal of rock, removal of structural organisms).
Effects of re-suspension The direct physical contact of fishing gear with the substratum can lead to the resuspension of sediments and the fragmentation of rock and biogenic substrata. The re-suspension, transport and subsequent deposition of sediment may affect the settlement and feeding of the biota in other areas. Sediment re-suspended as a result of bottom fishing will have a variety of effects including: releasing nutrients held in the sediment; exposure of anoxic layers; release of contaminants; increasing biological oxygen demand; and smothering of feeding and respiratory organs. The quantity of sediment re-suspended by trawling depends on sediment grain size and the degree of compaction, and is higher on mud and fine sand than on coarse sand. Transmissiometers, which measure background light levels in water, frequently record the highest levels of turbidity during periods of trawling activity (Churchill, 1989). In deeper water, where stormrelated bottom stresses have less influence, otter trawling activity contributed significantly to the re-suspension of fine material. Churchill (1989) calculated sediment budgets for areas of the mid-Atlantic Bight and concluded that trawling was the main factor initiating the offshore transport of sediment at depths of 100–140 m. However, the transport of sediment resulting from fishing activities would not produce significant large-scale erosion over a period of a few years. The effects of sediment re-suspension are clearer in deep-water environments that are relatively undisturbed. Thiel and Schriever (1990) experimentally harrowed an area of
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seabed at a depth of 4000 m. Their observations revealed that 80% of their study site was covered by fine material that had settled out from the resultant sediment plume. Although this study was designed to imitate the effects of deep-sea mining, the observations are also relevant for deep-sea trawling activities. The observations of sediment re-suspension in the deep sea may resemble the seasonal settlement of organic material that occurs in deep-sea regions (Angel and Rice, 1996). The sediment–water interface of marine sediments is an important site of benthic primary production. Brylinsky et al. (1994) found that benthic diatoms bloomed within otter door tracks 1 month after they had been created. They reasoned that the bloom was triggered by the release of nutrients from the sediment following trawling. The intensive trawling of Posidonia oceanica meadows in the Mediterranean Sea may lead to reductions in littoral primary productivity since large areas of P. oceanica are reported to have been killed by the mechanical action of fishing gears and the deposition of re-suspended sediment (Guillén et al., 1994). These meadows are known to be important sources of primary production, although the consequences of losses in production are not known. It is unlikely, using existing data, that large-scale changes in primary production could be correlated reliably with changes in fishing intensity.
Effects of fishing on the habitat
Effects of drive netting, poisons and explosives in the tropics Techniques such as drive netting, seining, poison and explosive fishing are used principally by small-scale and artisanal fishers fishing on tropical reefs. Although the effects attributable to the activities of individual fishers often are small in comparison with those attributable to commercial fishing boats using towed gears, the combined effects of their activities are considerable, given the large proportion of the coastal population involved in fishing (Pauly, 1988; Pauly et al.,
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1989; Dalzell et al., 1996). Many of the fishing techniques used to catch reef-associated fishes cause direct physical damage to the reef substratum. The most widely used destructive fishing techniques are drive netting (Carpenter and Alcala, 1977; Gomez et al., 1987), trapping (Munro et al., 1987) and fishing with explosives (Munro et al., 1987). In addition, those poisons widely used to catch fishes for the aquarium trade and consumption have the potential to cause chemical damage to corals and non-target fishes and invertebrates (Rubec, 1986; Eldredge, 1987; Pyle, 1993). Corals perform several important functions in tropical environments. They provide substrata for primary production, habitats for invertebrates and fishes, and often play a key role in protecting coasts from wave exposure and erosion. The rate at which reefs develop is determined by the balance between rates of accretion owing to the growth of corals, hydrocorals and coralline algae, and erosion owing to mechanical processes and bio-erosion. Fishing affects reefs directly when gears contact the reef substratum or indirectly by altering the relationships between those communities of plants, invertebrates and fishes that determine rates of reef accretion and bio-erosion. Coral accretion relies upon the successful settlement of young corals, and the maintenance of suitable conditions for their growth (Pearson, 1981). These processes may be affected by fishing activities. Drive netting techniques are used to catch a range of reef-associated fishes, which shelter within the reef matrix or shoal above the reef. These techniques are used extensively on coral reefs, and may range from small-scale village-based operations involving four or five fishers, to large commercial operations that target offshore reefs in the Philippines and South China Sea, and involve hundreds of divers (McManus, 1996). The process of drive netting requires that the fishers (who stand on the reef or dive) scare reef-associated fishes towards an encircling net or trap, using scaring devices such as weighted lines or poles. In shallow water, corals often are broken deliberately to scare closely reef-associated fishes such as groupers (Epinephelinae), snappers or emperors from their refuges. In deeper water,
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the Kayakas and muro-ami drive netting techniques involve teams of swimmers that repeatedly drop weighted scarelines on to the reef in order to drive fishes towards a bag net. Carpenter and Alcala (1977) calculated the damage to 1 ha of reef during a single muro-ami operation involving 50 fishers who each struck the bottom 50 times with a 4 kg weighted scareline: 6% of the total area of coral present was damaged. Blast fishing is practised on many reefs in the Atlantic, Pacific and Indian Oceans (Gomez et al., 1981; Polunin, 1983; Galvez and Sadorra, 1988; Ruddle, 1996). A variety of explosives are used, including those obtained from mines or removed from armaments. Pelagic fishes living above the reef are often targeted, rather than fishes living in direct association with the reef (Saila et al., 1993). Owing to the considerable variation in the types and sizes of charges used, and the depths at which they explode, it is difficult to make useful generalizations about the damage that they will cause. Alcala and Gomez (1987) report that a bottle bomb exploding at or near the bottom will shatter all corals within a radius of 1.15 m, and that a gallonsized drum will have the same effect over a radius of 5 m. A ‘typical’ charge will kill most marine organisms, including invertebrates, within a radius of 77 m. Such techniques are highly unselective, and Munro et al. (1987) report that post-larval and juvenile fishes are also killed. These young fishes would be about to recruit to the reef habitat, and the repeated effects of blast fishing on a large scale would reduce fish production from the reef. On those reefs from 15° to 30° either side of the equator, which are susceptible to hurricane damage, the effects of blast fishing often are localized and negligible in comparison with those of hurricanes (S. Jennings, personal observation). In other areas, especially in the Philippines, damage attributable to blast fishing is an increasing cause of concern. Stupefacients are used widely by reef fishers. Traditionally, poisons extracted from plants were used extensively for reef fishing but, in the last few decades, synthetic chemicals such as sodium cyanide and chlorine have been used more frequently (Rubec, 1986; Eldredge, 1987). McAllister (1988) estimated
that 150 t of sodium cyanide is used annually on Philippine reefs to catch aquarium fishes. There is little knowledge of the effects of these chemicals on the various life history stages of the reef biota (Rubec, 1986; Pyle, 1993), and while concentrations of stupefacients that have an acute effect are quickly dispersed, the chronic effects may be significant. The long-term direct effects of fishing on reefs are determined largely by the rate at which coral can accrete in relation to the rate at which it is damaged. The recovery and recolonization of coral communities following mechanical damage by fishing gears takes place when partially damaged colonies or coral fragments re-grow and when the substratum becomes suitable for coral settlement (Pearson, 1981). Saila et al. (1993) developed a model to examine the effects of blast fishing on reefs in the Philippines. At present-fishing intensities, the loss of diversity and coral cover would continue for approximately 25 years before recovery is expected. Coral growth rates are highly variable: 0.7–17.2 cm year−1 for branching species and 0.5–1.9 cm year−1 for massive species (Loya, 1976; Huston, 1985; Witman, 1988). Several studies of reef development following hurricanes and other natural events provide a useful guide to recovery rates. Published estimates of recovery time often vary widely because they reflect differences in the authors’ assumptions regarding the organization of coral communities and the meaning of ‘stability’ (Moran, 1986; Done, 1987, 1988; Done et al., 1988; Endean et al., 1988; Moran, 1990; Turner, 1994; McClanahan et al., 1996). However, a coral community dominated by fast growing branching species and which provides a suitable habitat for many reef fishes would develop within 5 years (Pearson, 1981).
Effects of towed bottom-fishing gear The short-term effects of fishing on seabed biota are well documented in recent studies (for reviews, see Jennings and Kaiser, 1998; Kaiser and De Groot, 2000). The results from short-term studies are informative and often have confirmed our expectations of the type of changes that might occur as a result of fishing activity. Nevertheless, the usefulness of
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each study on its own is limited by factors such as the specific location, type of gear used and season during which the study in question was undertaken. Viewed on their own, these individual studies can only be used to predict the outcome of fishing activities in a restricted number of situations. However, Collie et al. (2000) overcame this problem by extracting summary data from a population of fishing impact studies and undertook a meta-analysis (e.g. Gurevitch and Hedges, 1999) of the combined data set to ask the following questions:
• • •
Are there consistent patterns in the responses of benthic organisms to fishing disturbance? How does the magnitude of this response vary with habitat, depth, disturbance type and among taxa? How does the recovery rate of organisms vary with these same factors?
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Collie et al. (2000) found that the magnitude of the immediate response (i.e. change in abundance or biomass) of organisms to fishing disturbance varied significantly according to the type of fishing gear used in the study, the habitat in which the study was undertaken, and among different taxa. EFFECTS OF DIFFERENT GEARS The initial impacts of different fishing gears were mainly consistent with expectations. Intertidal dredging activities had a more marked effect than scallop dredging, which in turn had greater effects than otter trawling (Fig. 12.2). Although at first sight the apparent lack of effect from beam trawling is somewhat surprising, it is suspected that the relative paucity of data for this gear is almost certainly part of the explanation. It should also be borne in mind, however, that beam-trawling studies generally were conducted in relatively
Fig. 12.2. The predicted mean response derived from ANOVA of: (a) the response of different taxa to physical disturbance that occurred immediately after that disturbance had occurred; (b) the response of invertebrate abundance or biomass in different habitats; and (c) the response of invertebrate abundance or biomass to different gear types. Data are on a transformed scale where values correspond to percentage declines from controls as follows: –0.1 = 10%; –0.22 = 20%; –0.35 = 30%; –0.5 = 40%; –0.68 = 50%; –1.35 = 75%; –4.61 = 100%. In all cases, the initial response of the fauna was negative.
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dynamic sandy areas, where initial effects may be less apparent or are less easily detected. Fishing disturbance effects of intertidal dredging are likely to have the greatest initial effects on the biota because fishers are able to use the harvesting machinery accurately, working parallel lines along the shore. In contrast, fishers using towed nets in subtidal areas are unable actually to see precisely where their gear is fishing, although technological advances in positioning systems are making it increasingly easier to achieve very accurate positioning of fishing gear on the seabed. It is also important to note that it is easier for a scientist to collect samples accurately from intertidal compared with subtidal areas, where sampling error undoubtedly is introduced, with potential detrimental effects for the statistical power to detect change. Otter trawling appears to have the least significant impact on fauna compared with other gear, although it is necessary to flag a few warnings about this observation. First, it is the otter doors that hold the wings of the otter trawl open that have the greatest impact on the sediment habitat. However, the otter doors constitute a small proportion of the total width of the gear (~ 2 m against 40–60 m). Secondly, none of the studies published at that time considered the effects of rockhopper otter trawls on seabed communities. Recent studies have shown that rock hopper gears have considerable negative shortterm effects on emergent epifauna (Prena et al., 1999; McConnaughey et al., 2000; Pitcher et al., 2000). EFFECTS IN DIFFERENT HABITATS Several authors have suggested that the relative ecological importance of fishing disturbance will be related to the magnitude and frequency of background of natural disturbances that occur in a particular marine habitat (Kaiser, 1998; Auster and Langton, 1999). Certainly, it makes intuitive sense that organisms that inhabit unconsolidated sediments should be adapted to periodic sediment re-suspension and smothering. Similarly, it seems plausible that organisms living in sea-grass beds rarely experience repeated intense physical disturbances or elevated water turbidity as created
by bottom fishing gear (Fig. 12.1). Indeed, such intuition has been the cornerstone of hypotheses about impacts and recovery dynamics for benthos (e.g. Hall, 1994; Jennings and Kaiser, 1998). However, Collie et al. (2000) found that their initial impact results with respect to habitat were somewhat inconsistent among analyses. While the initial responses to fishing disturbance of taxa in sand habitats usually were less negative than in other habitats, a clear ranking for expected impacts did not emerge (Fig. 12.2). Such inconsistencies may reflect interactions among the factors arising from the unbalanced nature of the data, with many combinations of gear and habitat absent. For example, the relatively low initial impact on mud habitats may be explained by the fact that most studies were undertaken with otter trawls. If data were also available for the effect of dredgers on mud substrata, a more negative response for this habitat might have been observed. Nevertheless, it should be borne in mind that initial effects of disturbance might be hard to detect in mud communities, which often have low abundances of biota and which tend to be burrowed deep (10–200 cm) within the sediment. Presumably, the deeply burrowed fauna would be relatively well protected from the physical effects of disturbance, although the passage of the gear will cause their burrows to collapse. Whether these inconsistencies can be explained in this way can only await further study. It is important not to classify habitats by the particular nature of the sediment and to appreciate the generality of the categorization used by Collie et al. (2000). For example, intertidal sandflats inhabited by high densities of tubiculous worms such as spionids will be more stable (and hence more adversely affected by fishing) than sandflats with relatively little infauna (Thrush et al., 1996). IMMEDIATE EFFECTS ON BIOTA Collie et al. (2000) found that the most consistently interpretable result within their meta-analysis was the vulnerability of fauna, with a ranking of initial impacts that concurred broadly with expectations based on morphology and behaviour. The polychaete, Arenicola sp.,
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had the greatest initial response to disturbance, which is not surprising given that this was the target of a commercial fishery (Fig. 12.2). Collie et al. (2000) also undertook a regression tree analysis that perhaps provides the first quantitative basis for predicting the relative impacts of fishing under different situations. Following the tree from its root to the branches, we can make predictions, for example, about how a particular taxon would be affected initially by disturbance from a particular fishing gear in a particular habitat. Thus, trawling would reduce anthozoa (anemones, soft corals and sea ferns) by 68%, whereas asteroid starfishes would only be reduced by 21%. Similarly, repeated (chronic) dredging is predicted to lead to 93% reductions for anthozoa, malacostraca (shrimps and prawns), ophiuroidea (brittlestars) and polychaeta (bristle worms), whereas a single (acute) dredge event is predicted to lead to a 76% reduction. This approach ultimately might provide a useful quantitative framework for predicting fishing impacts that can be updated and refined as new data emerge.
Recovery Rates After Trawl Disturbance Soft sediments From a personal perspective, in an environment that is open to disturbance by fishing gear, the short-term effects of bottom-fishing disturbance on habitats and their biota are of interest, but of far less ecological importance than the issue of the potential for recovery or restoration. Unfortunately, relatively few studies of trawl disturbance have included a temporal component of sufficient duration to address longer term changes that occur as a result of bottom-fishing disturbance. This is almost certainly a result of the conflict between financial resources, project duration, statistical and analytical considerations. Nevertheless, Collie et al. (2000) were able to incorporate studies that included a recovery component into their analysis. This permitted
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them to speculate about the level at which physical disturbance becomes unsustainable in a particular habitat. For example, their study suggested that sandy sediment communities are able to recover within 100 days, which implies that they could perhaps withstand 2–3 incidents of physical disturbance per year without changing markedly in character (Fig. 12.3). This level of fishing disturbance is the average predicted rate of disturbance for the whole of the southern North Sea. However, when fishing effort data are collected at a fine spatial (9 km2) resolution (Rijnsdorp et al., 1998), it becomes clear that effort is distributed patchily and that some relatively small areas of the seabed are visited by more than 400 trawlers each year. This level of fishing equates to a total disturbance of approximately eight times per year (Rijnsdorp et al., 1998). If our recovery rate estimates for sandy habitats are realistic, this would suggest that these areas of the seabed are held in a permanently altered state by the physical disturbance associated with fishing activities. At this point, there are some important limitations within the data that should be considered. First, the small spatial scale (the maximum width of most of the disturbed areas examined was < 50 m) of most of the trawl impact studies make it likely that much of the re-colonization was via active immigration into disturbed patches rather than reproduction within patches. The authors found recovery to be slower if the spatial scale of impact was larger, as it would be on heavily fished grounds due to the additive effects of an entire fleet of trawlers. Secondly, it should be noted, that while one might predict accurately the recovery rate for small-bodied taxa such as polychaetes, which dominate the data set, sandy sediment communities often contain one or two long-lived and therefore vulnerable species. Note, for example, the occurrence of the large bivalve Mya arenaria in the intertidal zone of the Wadden Sea. While the majority of the benthos in this environment recovered within 6 months of lugworm dredging, the biomass of M. arenaria remained depleted for at least 2 years (Beukema, 1995). This delayed recovery of larger bodied
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organisms is no doubt even more important in habitats that are formed by living organisms (e.g. soft corals, sea fans and mussels) as the habitat recovery rate is linked directly to the re-colonization and growth rate of these organisms. By now, there is sufficient evidence in the literature to suggest that under conditions of repeated and intense bottomfishing disturbance, a shift from communities dominated by relatively high biomass species towards dominance by high abundances of small-sized organisms will occur.
Biogenic habitats
Fig. 12.3. Results from a meta-analysis of the effects of fishing disturbance on benthic communities. The scatter plots of the relative change of all species (each data point represents the relative abundance of a different species on each different sampling date) in different habitats at time intervals after the occurrence of a fishing disturbance. The fitted curves show the predicted time trajectory for recovery to occur. On the y-axis, 0 shows no relative change in abundance; negative values show a relative decrease in abundance.
It is clear that intensively fished areas are likely to be maintained in a permanently altered state, inhabited by fauna adapted to frequent physical disturbance. These effects will be most apparent for stable types of habitats that contain structural biogenic components. Presumably, such habitats will have the longest recovery time compared with less stable substrata. Yet it is for these habitats that the paucity of data is most apparent. While it would appear that none of the habitats included in Collie et al.’s (2000) study fall into this category, some new data are beginning to emerge. Hall-Spencer and Moore (2000) examined the effects of fishing disturbance on maerl beds. Maerl beds are composed of highly dichotomous calcareous algae. This forms a complex substratum with a high degree of three-dimensional complexity. Not surprisingly, the associated assemblages have high diversity and many of the associated species are large bodied and slow growing. Hall-Spencer and Moore (2000) showed that 4 years after the occurrence of an initial scallop-dredging disturbance had occurred, certain fauna, such as the nestbuilding bivalve Limaria hians, had still not re-colonized trawl tracks (Fig. 12.4). Similarly, work by Sainsbury (1987) and Sainsbury et al. (1997) suggests that recovery rates may exceed 15 years for sponge and coral habitats off the western coast of Australia. The presence of such habitats was important for fish species of commercial importance.
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Fig. 12.4. Numbers of byssus nests of the bivalve Limaria hians recorded within test and control plots prior to and over 4 years after experimental scallop dredging on the previously undredged ground. Error bars are ± SE (n = 20), significant differences (ANOVA on log transformed data) are indicated (*P < 0.05) (adapted from Hall-Spencer and Moore, 2000).
Deducing the Effects of Chronic Disturbance The perceived problems that might be associated with intense and prolonged bottomfishing disturbance have only been examined in the last 20 years. However, bottom-fishing fleets have been in operation much longer. For example, the beam trawl fleet in the southern North Sea expanded dramatically through the 1960s and 1970s. Consequently, many present-day studies have been undertaken in what is already a considerably altered environment. Despite our efforts to predict the outcome of fishing activities for existing benthic communities, we often are unable to deduce the original composition of the fauna because data gathered prior to the era of intensive bottom fishing are sparse. This is an important caveat, because recent analyses of the few existing historical data
sets suggest that larger bodied organisms (both fish and benthos) were more prevalent prior to intensive bottom trawling (Greenstreet and Hall, 1996; Frid et al., 2000; Rumohr and Kujawski, 2000). Moreover, in general, epifaunal organisms are less prevalent in areas subjected to intensive bottom fishing (Collie et al., 1997; Sainsbury et al., 1997; Kaiser et al., 2000a, b; McConnaughy et al., 2000; Rumohr and Kujawski, 2000). An important consequence of this effect is the reduction in habitat complexity (architecture) that accompanies the removal of sessile epifauna. Nevertheless, it has been hard to demonstrate convincingly that towed bottomfishing activity has been responsible for changes in bottom fauna and habitats. Often, effort data are lacking at a scale or over a time period that is relevant to ascertain the disturbance history of a particular area of seabed. In the few instances when such data
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have been available, observations have indicated consistently a shift from dominance by high-biomass organisms towards communities dominated by small-bodied opportunistic species (Collie et al., 1997; Engel and Kvitek, 1998; Bradshaw et al., 2000; Kaiser et al., 2000a,b; McConnaughey et al., 2000). It is becoming increasingly apparent that habitat modification appears to have important consequences for fish communities (Sainsbury, 1987; Auster and Langton, 1999; Kaiser et al., 1999).
Indirect Effects of Fishing Interactions that result from exploitation of target species on tropical reefs The overexploitation or effective conservation of fishes on tropical reefs can affect community structure and a range of ecosystem processes. Many reefs are fished intensively since they provide the main sources of protein and income for coastal people with few other opportunities for fishing, farming or hunting. Reef fishers target species from all trophic groups and, on many fished reefs, the abundance of herbivorous and invertebrate feeding fishes has been reduced by an order of magnitude or more (Russ, 1991). The main groups of algal consumers on reefs are herbivorous fishes and sea urchins. The abundance of sea urchins is regulated by recruitment success, food supply and natural mortality due to predation and disease. The main urchin predators are fishes such as emperors (family Lethrinidae) and triggerfishes (Balistidae) that are also targeted by fishers (McClanahan, 1995b). On some reefs in the Caribbean and East Africa, fish predation appears to play a key role in controlling the abundance of urchin populations and they have proliferated following the overfishing of their predators (McClanahan, 1992, 1995a). Once urchins have become abundant, they graze the majority of algal production. Urchins can tolerate low algal biomass because they have low consumption and respiration rates. This allows them to outcompete herbivorous fishes that have higher
consumption and respiration rates and reach maximum biomass levels an order of magnitude higher (McClanahan, 1992). Since the herbivorous fishes are poor competitors, they may not recover to former levels of abundance when fishing is stopped (McClanahan, 1995a). As they graze, urchins erode the reef matrix and prevent the settlement and growth of coral recruits. Unless recruitment failure or urchin disease leads to a collapse of urchin populations, other intervention is needed to promote recovery of the reef ecosystem. McClanahan et al. (1996) attempted such intervention on a small scale by deliberate removal of urchins. When they removed urchins from unfished experimental plots on Kenyan reefs, there were significant increases in algal cover and fish abundance within 1 year. However, on fished reefs, herbivorous fishes were less abundant, and the algae rapidly overgrew corals as they proliferated. The ecosystem shifts that McClanahan et al. (1996) induced in fished areas by urchin removal were remarkably similar to those observed in the heavily fished Caribbean when there was mass urchin mortality following disease. Here, the loss of urchins led to heavy growths of algae that soon dominated the reef community (Carpenter, 1985). The effects of predator removal on urchin populations contrast with the effects of piscivore removal on reef fish populations. While many studies have shown that the abundance of piscivorous reef fishes is dramatically reduced by fishing, there is little evidence for a corresponding increase in the abundance of their prey. We will give some examples of this effect, and consider why the response of prey fish communities is so weak. Several studies have documented significant decreases in the abundance of piscivorous target species following fishing, and yet there was no evidence for a corresponding increase in the abundance of their prey (Bohnsack, 1982; Russ, 1985; Jennings and Polunin, 1997). The reasons for this are likely to be linked to the structure of reef fish communities, where phylogenetic groupings contain many species, with a wide range of life history traits, behavioural differences and feeding strategies (Hiatt and Strasburg, 1960; Parrish et al., 1985, 1986; Hixon, 1991).
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Moreover, most fish species undergo marked ontogenetic changes in diet, and act as the prey and predators of other species in the course of their life history. As a result, while the collective impacts of predators are large, the impacts of individual predator species on the dynamics of their prey are minor. This effect was termed diffuse predation by Hixon (1991). It is worth noting that on much smaller scales (m2 rather than km2), there is some evidence for the role of predation as a structuring force, particularly when habitat or refuge space is directly limited. Thus, Caley (1993), Hixon and Beets (1993) and Carr and Hixon (1995) have conducted elegant studies that demonstrated that experimental reductions in piscivore abundance lead to detectable decreases in the abundance and diversity of their prey. However, even at these scales, it is widely accepted that recruitment variation has a more significant impact on population structure (Sale, 1980; Doherty, 1991; Doherty and Fowler, 1994).
Essential Fish Habitat Recent amendments to the USA Magnuson– Stevens Act require fisheries managers to define ‘essential fish habitat’ (EFH) and address the impact of fishing gear in their management plans (Benaka, 1999). This is probably one of the first legislative steps taken in fisheries management that will require the assimilation and application of the scientific knowledge outlined in the paragraphs above. In many ways, this legislation is one of the first measures to embrace an ecosystem perspective in fisheries management. In some instances, it is fairly simple to identify those habitats that might be considered essential to the life history of some species. Such habitats include spawning and nursery areas, many of which are protected from fishing activity in European waters. However, of equal relevance are the habitat quality issues that affect the acquisition of food and the avoidance of predators. Hence, there is an urgent need to identify those habitats that have an important or ‘essential’ functional
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role for particular species or types of fish (e.g. piscivores/herbivores/omnivores or flatfish/roundfish) at other stages of their life history. Previous studies of the relationship between fish and shellfish assemblages and their environment have focused on variables such as salinity, depth and substratum type (e.g. Overholtz and Tyler, 1985; Smale et al., 1993). Yet, while such environmental parameters are in some cases good correlates of certain fish assemblages, they do not necessarily define the essential features of a specific habitat, rather they constitute a component of that habitat. Habitat complexity and composition (e.g. grain size composition) appear to be important physical features for some fish species (e.g. Sainsbury, 1987; Gibson and Robb, 1992). Many studies have already demonstrated the relationship between flatfish species and the sediment particle composition of the seabed, which may be more important than the occurrence of associated epibenthic structures or fauna that occur in that habitat (e.g. Gibson and Robb, 1992; Rogers, 1992). Hence, a specific particle size composition may be essential for flatfish, whereas the presence of large sessile epifauna or rocky substrata might be considered non-essential. In contrast, there is good evidence to suggest that structural complexity can have important implications for the survival of juvenile roundfishes (e.g. Walters and Juanes, 1993). Habitat complexity is a product of the surface topography and internal structure of the substratum and the sessile epifauna that grow upon it. Reef-forming organisms can result in habitats of very high complexity, providing a multitude of refuges for a diverse range of species. More subtle features such as sand ridges and pits created by the feeding or burrowing action of benthic fauna may provide shelter for bottom-dwelling fish species (e.g. Auster et al., 1997). Bottom-fishing activities are capable of greatly reducing habitat complexity by either direct modification of the substratum or removal of the fauna that contribute to surface topography (Jennings and Kaiser, 1998; Auster and Langton, 1999). Hence, degradation of habitat complexity by fishing activities may lead to changes in the associated fish assemblages (e.g. Sainsbury
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et al., 1997). Alteration of habitat features has been shown to have important consequences for freshwater fishes, and this is the caveat that underpins much of the ecological restoration projects centred on salmonid habitats. An initial study of habitat–fish assemblage relationships indicated that even subtle alterations in habitat characteristics can be linked to a shift in the dominance of certain fish species within the assemblage (Kaiser et al., 1999). Presumably, a good understanding of the link between fish and their habitat would enable us to predict the consequences of habitat alteration. For example, for certain species such as sole (Solea solea) that preferentially live in relatively uniform sandy areas, the exclusion of towed bottom-fishing gear from an area of the seabed could permit the growth of emergent sessile fauna that make the environment better suited to predatory flatfishes such as plaice (Pleuronectes platessa) and dab (Limanda limanda) (Kaiser et al., 1999). Thus, in the case of the sole fishery, the fishing activity may maintain the seabed habitat in a condition that favours the target species. Quite clearly, the opposite would be true for any species of fish favouring more complex habitats.
Integrating Habitat Conservation Objectives into Fisheries Management It would appear that with sufficient scientific information, it should be possible to formulate a regime of fishing effort (= physical disturbance for towed bottom-fishing gear) that would be environmentally sustainable. Here we define environmentally sustainable as the process by which the habitat and its associated biological assemblage can recover before a subsequent disturbance event. For example, in shallow sandy areas of the seabed, two to three physical disturbances of the seabed every year may have little or no net effect on the habitat or resident assemblage. However, at present, the definition of sandy areas is too imprecise a habitat criterion on which to base such a management plan. We know, for example, that sand flats that are dominated
by tube-building spionid worms take much longer to recover if these worms are removed through physical disturbance, as the worms normally have a stabilizing effect on the habitat (Thrush et al., 1996). Nevertheless, the complete exclusion of bottom-fishing disturbance from sandy habitats that are fished at present may actually have a negative effect on the fishery, as suggested above. Physical disturbance will, to some extent, promote dominance by opportunistic species such as small polychaetes that form a major component of the ecosystems of many commercially important flatfish species (Rijnsdorp and van Leeuwen, 1996). What is clear from the studies undertaken to date is that there exist communities and habitats that are so sensitive to physical disturbance that all forms of bottom fishing with towed gear should be considered for exclusion from these areas forthwith. As a matter of urgency, there is need to identify other habitats that have long recovery times and that are exposed (or might in the future be exposed) to towed bottom-fishing gear – the most likely candidates are those that contain a high proportion of structural fauna. In European waters, examples of such habitats would include:
• • • • • •
Deep sea coral reefs of Lophelia pertusa Maerl beds Reefs of mussels (Modiolus modiolus) and Limaria hians Areas of the seabed with aggregations of sea fans Beds of fan mussels (e.g. Atrina fragilis) Sea-grass meadows.
It is important at this point to define what we mean by sensitive fauna or habitats. Sensitive fauna may be defined by their physical attributes (e.g. fragility of body structure), their reproductive strategies (e.g. infrequent recruitment or low reproductive output) or remaining population size (e.g. the lower the population size, the more vulnerable to extirpation that species will be). Sensitive, non-sensitive, structured and non-structured fauna and habitats – all will be affected to some degree by towed bottom-fishing gear. However, for effective management, we need
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to define and identify vulnerable species or habitats, and the extent (in area) of these resources, and the management approach(es) that should be used. This usually requires clearly defined management objectives (that can be measured and monitored) and data on fishing effort (level and spatial distribution), impact and recovery times. From this information, management strategies can be developed and tested against the objective(s). We can then start defining what we mean by a sustainable fishery.
Fishery Management Measures that Include Habitat Management Fisheries management that includes habitat management as an objective could be achieved through a number of different mechanisms. Total exclusion of all fishing effort will achieve habitat conservation, provided that there are no other extrinsic factors that negatively affect that habitat (e.g. agricultural runoff from adjacent land masses). Other measures include networks of area closures at different spatial scales that are determined by the demography of the species to be protected. Temporal closures may achieve habitat preservation if the habitat or species in question have generation times that fall within the time scale of the temporal closure. The total exclusion of certain forms of fishing activities from areas of the seabed will inevitably lead to opposition from the fishing industry and, by its nature, such measure are extreme. Nevertheless, a recent large-scale study on the NW Atlantic coast of North American has demonstrated elegantly the effectiveness of such large-scale closures. Alternatively, inshore fisheries lend themselves to the partitioning of seabed resources such that certain areas of the seabed can continue to be exploited using gear that cause minimal environmental damage. The two following examples illustrate the potential of such approaches. In New England, USA, seasonal closed areas have been an important component of fisheries management since the early 1970s,
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but had little impact on the groundfish stocks that they were designed to protect. In 1994, three large areas that in total covered 17,000 km2 of the seabed were closed all year to all fishing gears that might retain groundfish (trawls, scallop dredges, hooks, etc.). These closed areas were maintained for 5 years and were found effectively to protect the more sedentary components of the assemblage, such as flatfishes, skates and scallops. Although less protection was afforded to cod (Gadus morhua) and haddock (Melanogrammus aeglefinus), additional legislation to protect specific important juvenile habitat lowered stock-wide mortality rates. Scallop dredgers were excluded because they took a by-catch of groundfish species. The relaxation of fishing effort on scallops had dramatic effects and led to a 14-fold increase in scallop biomass within the closed areas during 1994–1998 (Fig. 12.5). A portion of the closed areas was re-opened to scallop dredging in 1999 (Murawski et al., 2000). The returns of scallops during this period were so encouraging that managers are now contemplating a formal ‘area rotation’ scheme for this fishery, presumably on a time scale of 4–5 years. The second example comes from an inshore fishery off the south coast of England. When two commercially important species co-exist in the same habitat, conflict may arise between different sectors of the fishing industry. A good example of this situation is when fishers using towed bottom-fishing gears (scallop dredges, beam trawls and otter trawls) operate in the same areas in which fixed bottom gear (crab pots) are deployed. Kaiser et al. (2000b) examined an area subject to a voluntary agreement between these two sectors of the fishing industry such that some areas are used exclusively by fixed-gear fishers, some are shared seasonally by both sectors and others are open to all methods of fishing year-round. This agreement was enacted to resolve conflict between the two sectors of the industry. An additional perceived benefit of this agreement was the possible protection of the seabed from towed bottom-fishing gear. Kaiser et al. (2000b) undertook comparative surveys of the benthic habitat and communities within the area covered by the agreement
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Fig. 12.5. Standardized abundance of sea scallops (numbers per dredge tow) by shell height, taken in the July National Marine Fisheries Service dredge survey on Georges Bank. Data are presented separately for the areas closed and those open to scallop dredging. Harvestable animals are indicated by the 50% selection line (from Murawski et al., 2000).
Fig. 12.6. Abundance/biomass curves of samples collected from areas protected from towed bottomfishing gear (low disturbance), areas open seasonally to towed bottom-fishing gear and those areas that are fished all year with towed bottom-fishing gear (high disturbance). As the level of bottom fishing disturbance increases, the biomass curve (B) converges with the abundance curve (A), which is a typical response in stressed communities (adapted from Kaiser et al., 2000b).
and compared different areas subjected to a range of fishing disturbance regimes. Communities found within the areas closed to towed fishing gears were significantly different from those open to fishing either permanently or seasonally. Abundance/biomass curves plotted for the benthic fauna demonstrated that the communities within the closed areas were dominated by higher biomass and emergent fauna that increase habitat
complexity (Fig. 12.6). Areas fished by towed gear were dominated by smaller bodied fauna and scavenging taxa. While it would appear that gear restriction management regimes have the added benefit of conserving habitats, target species and benthic fauna within the management area, it is at present not possible to determine whether there are any wider benefits for the fishery that exploits the target species outside the management area.
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Future Research Priorities With respect to the design of future studies, we feel that experimentalists wishing to address the fishing impacts issue would be best served by abandoning short-term, small-scale pulse experiments. Instead, the scientific community should be arguing for support to undertake much larger scale press and relaxation experiments. One half of the experiment has already been done – since fishing activity has been providing the press for many years; what we now require are more carefully designed closed-area contrasts. There are clear advantages to this approach. First, the results obtained are clearly interpretable in terms of real-world intensities of fishing disturbance. Secondly, the spatial scale of the protected areas could be relatively small (and hence replicated, to fulfil the requirements for sound experimental design) without compromising unduly the interpretation of recovery dynamics: estimates of recovery in small protected areas in a sea of disturbance are likely to be conservative, while recovery in small, deliberately disturbed patches are not. Thirdly, the experiments would be conducted in the very habitats (i.e. real fishing grounds) concerning which the question of recovery is actually being posed. Our current understanding of the functional role of many of the larger bodied long-lived species is limited and should be addressed to predict the outcome of permitting chronic fishing disturbance in areas where these animals occur. In addition, our understanding of the ecosystem services that many of these species provide is limited by a paucity of scientific understanding (e.g. Hall-Spencer and Moore, 2000). To date, the majority of studies that have addressed fishing impacts on the seabed have concentrated on the biota, with little reference to gear–sediment interactions (Pilskaln et al., 1998). As fishing gears disturb soft sediment, they produce sediment plumes and re-mobilize previously buried organic and inorganic matter. Presumably this increases the release of nutrients into the water column and has important consequences for rates of
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biogeochemical cycling. To date, this issue has received little attention, with the exception of one or two as yet unpublished studies.
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of massive corals of the genus Porites: evidence of population resilience. Coral Reefs 6, 75–90. Done, T.J. (1988) Simulation of recovery of predisturbance size structure in populations of Porites corals damaged by crown of thorns starfish Acanthaster planci (L.) Marine Biology 100, 51–61. Done, T.J., Osborne, K. and Navin, K.F. (1988) Recovery of corals post-Acanthaster: progress and prospects. Proceedings of the Sixth International Coral Reef Symposium 2, 137–142. Eldredge, L. (1987) Poisons for fishing on coral reefs. In: Salvat, B. (ed.) Human Impacts on Coral Reefs: Facts and Recommendations. Antenne Museum, French Polynesia, pp. 61–66. Endean, R., Cameron, A.M. and de Vantier, L.M. (1988) Acanthaster planci predation on massive corals: the myth of rapid recovery of devastated reefs. Proceedings of the Sixth International Coral Reef Symposium 2, 143–155. Engel, J. and Kvitek, R. (1998) Effects of otter trawling on a benthic community in Monterey Bay National Marine Sanctuary. Conservation Biology 12, 1204–1214. Estes, J.A. and Duggins, D.O. (1995) Sea otters and kelp forests in Alaska: generality and variation in a community ecological paradigm. Ecological Monographs 65, 75–100. Frid, C.L.J., Harwood, K.G., Hall, S.J. and Hall, J.A. (2000) Long-term changes in the benthic communities on North Sea fishing grounds. ICES Journal of Marine Science 57, 1303–1309. Galvez, R. and Sadorra, M.S.M. (1988) Blast fishing: a Philippine case study. Tropical Coastal Area Management 3, 9–10. Gibson, R.N. and Robb, L. (1992) The relationship between body size, sediment grain size and the burying ability of juvenile plaice Pleuronectes platessa (L.). Journal of Fish Biology 40, 771–778. Gomez, E.D., Alcala, A.C. and San Diego, A.C. (1981) Status of Philippine coral reefs. Proceedings of the Fourth International Coral Reef Symposium 1, 275–282. Gomez, E., Alcala, A. and Yap, H. (1987) Other fishing methods destructive to coral. In: Salvat, B. (ed.) Human Impacts on Coral Reefs: Facts and Recommendations. Antenne Museum, French Polynesia, pp. 67–75. Grassle, J.F. and Saunders, H.L. (1973) Life histories and the role of disturbance. Deep Sea Research 20, 643–659. Greenstreet, S.P.R. and Hall, S.J. (1996) Fishing and groundfish assemblage structure in the northwestern North Sea: an analysis of long-term and spatial trends. Journal of Animal Ecology 65, 577–598.
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Guillén, J., Ramos, A., Martinéz, L. and Sánchez Lizaso, J. (1994) Antitrawling reefs and the protection of Posidonia oceanica (L.) delile meadows in the western Mediterranean Sea: demands and aims. Bulletin of Marine Science 55, 645–650. Gurevitch, J. and Hedges, L.V. (1999) Statistical issues in ecological meta-analysis. Ecology 80, 1142–1149. Hall, S.J. (1994) Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology Annual Review 32, 179–239. Hall-Spencer, J.M. and Moore, P.G. (2000) Impact of scallop dredging on maerl grounds. In: Kaiser, M.J. and De Groot, S.J. (eds) Effects of Fishing on Non-target Species and Habitats: Biological, Conservation and Socio-economic Issues. Blackwell Science, Oxford, pp. 105–118. Hiatt, R.W. and Strasburg, D.W. (1960) Ecological relationships of the fish fauna on coral reefs of the Marshall Islands. Ecological Monographs 30, 65–127. Hixon, M. (1991) Predation as a process structuring coral reef fish communities. In: Sale, P. (ed.) The Ecology of Fishes on Coral Reefs. Academic Press, San Diego, pp. 475–508. Hixon, M.A. and Beets, J.P. (1993) Predation, prey refuges and the structure of coral reef fish assemblages. Ecological Monographs 63, 77–101. Huston, M. (1985) Variation in coral growth rates with depth at Discovery Bay, Jamaica. Coral Reefs 4, 19–25. Jennings, S. and Kaiser, M. (1998) The effects of fishing on marine ecosystems. Advances in Marine Biology 34, 201–352. Jennings, S. and Polunin, N.V.C. (1997) Impacts of predator depletion by fishing on the biomass and diversity of non-target reef fish communities. Coral Reefs 16, 71–82. Kaiser, M.J. (1998) Significance of bottomfishing disturbance. Conservation Biology 12, 1230–1235. Kaiser, M.J. and De Groot, S.J. (2000) Effects of Fishing on Non-target Species and Habitats: Biological, Conservation and Socio-economic Issues. Blackwell Science, Oxford. Kaiser, M.J. and Jennings, S. (2001) An ecosystem perspective on conserving targeted and nontargeted species. In: Reynolds, J.D., Mace, G.M., Redford, K.H. and Robinson, J.G. (eds) Conservation of Exploited Species 343–369. Cambridge University Press, Cambridge. Kaiser, M.J. and Spencer, B.E. (1996) The effects of beam-trawl disturbance on infaunal communities in different habitats. Journal of Animal Ecology 65, 348–358.
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Kaiser, M.J., Rogers, S.I. and Ellis, J.R. (1999) Importance of benthic habitat complexity for demersal fish assemblages. In: Benaka, L. (ed.) Fish Habitat: Essential Fish Habitat and Restoration. American Fisheries Society, 22, Bethesda, Maryland, Symposium pp. 212–223. Kaiser, M.J., Ramsay, K., Richardson, C.A., Spence, F.E. and Brand, A.R. (2000a) Chronic fishing disturbance has changed shelf sea benthic community structure. Journal of Animal Ecology 69, 494–503. Kaiser, M.J., Spence, F.E. and Hart, P.J.B. (2000b) Fishing gear restrictions and conservation of benthic habitat complexity. Conservation Biology 14, 1512–1525. Loya, Y. (1976) Recolonization of Red Sea corals affected by natural catastrophes and manmade perturbations. Ecology 57, 278–289. McAllister, D.E. (1988) Environmental, economic and social costs of coral reef destruction in the Philippines. Galaxea 7, 161–178. McClanahan, T.R. (1992) Resource utilization, competition and predation: a model and example from coral reef grazers. Ecological Modelling 61, 195–215. McClanahan, T. (1995a) A coral-reef ecosystemfisheries model – impacts of fishing intensity and catch selection on reef structure and processes. Ecological Modelling 80, 1–19. McClanahan, T. (1995b) Fish predators and scavengers of the sea urchin Echinometra mathaei in Kenyan coral-reef marine parks. Environmental Biology of Fishes 43, 187–193. McClanahan, T., Kakamura, A., Muthiga, N., Yebio, M. and Obura, D. (1996) Effects of sea-urchin reductions on algae, coral and fish populations. Conservation Biology 10, 136–154. McConnaughey, R.A., Mier, K.L. and Dew, C.B. (2000) An examination of chronic trawling effects on soft-bottom benthos of the eastern Bering Sea. ICES Journal of Marine Science 57, 1377–1388. McManus, J. (1996) Social and economic aspects of reef fisheries and their management. In: Polunin, N. and Roberts, C. (eds) Reef Fisheries. Chapman and Hall, London, pp. 249–281. Moran, P.J. (1986) The Acanthaster phenomenon. Oceanography and Marine Biology Annual Review 24, 379–348. Moran, P.J. (1990) Acanthaster planci (L.): biographical data. Coral Reefs 9, 95–96. Munro, J.L., Parrish, J.D. and Talbot, F.H. (1987) The biological effects of intensive fishing upon reef fish communities. In: Salvat, B. (ed.) Human Impacts on Coral Reefs: Facts and Recommendations. Antenne Museum, French Polynesia, pp. 41–49.
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Murawski, S.A., Brown, R., Lai, H.-L., Rago, P.J. and Hendrickson, L. (2000) Large-scale closed areas as a fishery-management tool in temperate marine systems: the Georges Bank experiment. Bulletin of Marine Science 66, 775–798. Overholtz, W. and Tyler, A. (1985) Longterm responses of demersal fish assemblages of Georges Bank. US Fisheries Bulletin 83, 507–520. Parrish, J.D., Callahan, M.W. and Norris, J.E. (1985) Fish trophic relationships that structure reef communities. Proceedings of the Fifth International Coral Reef Symposium 4, 73–78. Parrish, J., Norris, J., Callahan, M., Magarifugi, E. and Schroeder, R. (1986) Piscivory in a coral reef community. In: Caillet, G.M. and Simenstad, C.A. (eds) Gutshop ‘81: Fish Food Habits and Studies. University of Washington, Seattle, pp. 73–78. Pauly, D. (1988) Some definitions of overfishing relevant to coastal zone management in Southeast Asia. Tropical Coastal Area Management 3, 14–15. Pauly, D., Silvestre, G. and Smith, I.R. (1989) On development, fisheries and dynamite: a brief review of tropical fisheries management. Natural Resource Modelling 3, 307–329. Pearson, R. (1981) Recovery and recolonisation of coral reefs. Marine Ecology Progress Series 4, 105–122. Pilskaln, C.H., Churchill, J.H. and Mayer, L.M. (1998) Resuspension of sediment by bottom trawling in the Gulf of Maine and potential geochemical consequences. Conservation Biology 12, 1223–1229. Pitcher, C.R., Poiner, I.R., Hill, B.J. and Burridge, C.Y. (2000) Implications of the effects of trawling on sessile megazoobenthos on a tropical shelf in northeastern Australia. ICES Journal of Marine Science, 57, 1359–1368. Polunin, N.V.C. (1983) The marine resources of Indonesia. Oceanography and Marine Biology Annual Review 21, 455–531. Posey, M., Lindberg, W., Alphin, T. and Vose, F. (1996) Influence of storm disturbance on an offshore benthic community. Bulletin of Marine Science 59, 523–529. Prena, J., Schwinghamer, P., Rowell, T.W., Gordon, D.C.J., Gilkinson, K.D., Vass, W.P. and McKeown, D.L. (1999) Experimental otter trawling on a sandy bottom ecosystem of the Grand Banks of Newfoundland: analysis of trawl bycatch and effects on epifauna. Marine Ecology Progress Series No. 181. Pyle, R. (1993) Marine aquarium fish. In: Wright, A. and Hill, L. (eds) Nearshore Marine Resources of
the South Pacific. Institute of Pacific Studies, Suva, pp. 135–176. Rees, E.I.S., Nicholaidou, A. and Laskaridou, P. (1977) The effects of storms on the dynamics of shallow water benthic associations. In: Keegan, B.F., Ceidigh, P.O. and Boaden, P.J.S. Biology of Benthic Organisms. Pergamon, Oxford, pp. 465–474. Rijnsdorp, A.D. and Leeuwen, P.I.v. (1996) Changes in growth of North Sea plaice since 1950 in relation to density, eutrophication, beam-trawl effort, and temperature. ICES Journal of Marine Science 53, 1199–1213. Rijnsdorp, A.D., Buijs, A.M., Storbeck, F. and Visser, E. (1998) Micro-scale distribution of beam trawl effort in the southern North Sea between 1993 and 1996 in relation to the trawling frequency of the sea bed and the impact on benthic organisms. ICES Journal of Marine Science 55, 403–419. Rogers, S.I. (1992) Environmental factors affecting the distribution of Dover sole (Solea solea L.) within a nursery area. Netherlands Journal of Sea Research 29, 151–159. Rubec, P. (1986) The effects of sodium cyanide on coral reefs and marine fish in the Philippines. In: McClean, J., Dizon, L. and Hosillos, L. (eds) The First Asian Fisheries Forum. Asian Fisheries Society, Manila, pp. 297–302. Ruddle, K. (1996) Traditional management of reef fishing. In: Polunin, N. and Roberts, C. (eds) Reef Fisheries. Chapman and Hall, London, pp. 315–335. Rumohr, H. and Kujawski, T. (2000) The impact of trawl fishery on the epifauna of the southern North Sea. ICES Journal of Marine Science 57, 1389–1394. Russ, G.R. (1985) Effects of protective management on coral reef fishes in the central Philippines. Proceedings of the Fifth International Coral Reef Symposium 4, 219–224. Russ, G.R. (1991) Coral reef fisheries: effects and yields. In: Sale, P. (ed.) The Ecology of Fishes on Coral Reefs. Academic Press, San Diego, pp. 601–635. Saila, S., Kocic, V. and McManus, J. (1993) Modelling the effects of destructive fishing practices on tropical coral reefs. Marine Ecology Progress Series 94, 51–60. Sainsbury, K.J. (1987) Assessment and management of the demersal fishery on the continental shelf of northwestern Australia. In: Polovina, J.J. and Ralston, S. (eds) Tropical Snappers and Groupers – Biology and Fisheries Management. Westview Press, Boulder, Colorado, pp. 465–503. Sainsbury, K.J., Campbell, R.A., Lindholm, R. and Whitelaw, A.W. (1997) Experimental
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management of an Australian multispecies fishery: examining the possibility of trawl-induced habitat modification. In: Pikitch, K., Huppert, D.D. and Sissenwine, M.P. (eds) Global Trends: Fisheries Management. Symposium 20. American Fisheries Society, Bethesda, Maryland, pp. 107–112. Sale, P.F. (1980) The ecology of fishes on coral reefs. Oceanography and Marine Biology Annual Review 18, 367–421. Smale, M., Roel, B., Badenhorst, A. and Field, J. (1993) Analysis of the demersal community of fish and cephalopods on the Agulhas Bank, South Africa. Journal of Fish Biology 43A, 169–191. Theil, H. and Schriever, G. (1990) Deep-sea mining, environmental impact and the DISCOL project. Ambio 19, 245–250. Thrush, S.F., Whitlatch, R.B., Pridmore, R.D., Hewitt, J.E., Cummings, V.J. and Wilkinson,
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M.R. (1996) Scale-dependent recolonization: the role of sediment stability in a dynamic sandflat habitat. Ecology 77, 2472–2487. Turner, S. (1994) The biology and population outbreaks of the corallivorous gastropod Drupella on Indo-Pacific reefs. Oceanography and Marine Biology Annual Review 32, 461–530. Walters, C.J. and Juanes, F. (1993) Recruitment limitation as a consequence of natural selection for use of restricted feeding habitats and predation risk taking by juvenile fishes. Canadian Journal of Fisheries and Aquatic Science 50, 2058–2070. Watling, L. and Norse, E.A. (1998) Disturbance of the seabed by mobile fishing gear: a comparison to forest clearcutting. Conservation Biology 12, 1180–1197. Witman, J. (1988) Effects of predation by the fireworm Hermodice carunculata on milleporid corals. Bulletin of Marine Science 42, 446–458.
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13
The Magnitude and Impact of By-catch Mortality by Fishing Gear Robin Cook FRS Marine Laboratory, Aberdeen, UK
Abstract Most fishing operations trap organisms that are not the primary fishing target, and are commonly referred to as the by-catch. It may include small individuals of the target species, or other species with little or no commercial value. The problem is widespread, with a global estimate of approximately 20 million t, equivalent to about a quarter of the total world landings. Shrimp fisheries tend to generate the largest quantities of by-catch, and fisheries for small pelagics the least. By-catch rates in mixed demersal and large pelagic fisheries are intermediate. By-catch arises because fishing gears have imperfect selection properties, but the problem is made worse by economic pressures resulting from overexploitation. This leads to inefficient use of resources and changes in the abundance of both target and non-target species. Some by-catch species, including certain fish, reptiles, birds and mammals, may be threatened with extinction. Raised public awareness means that these conservation issues increasingly influence fishery management. Much of the by-catch is simply discarded at sea. While not intended, the imposition of regulations such as minimum landing sizes and catch restrictions may encourage discarding. Most discards do not survive, but the material provides food for other organisms, especially scavengers, whose abundance may increase. Technical conservation measures, which involve modifications to fishing gear or practices, offer an effective means of reducing by-catch. For trawls, these include grids and square mesh panels that sort animals by size, allowing a part of the catch to escape. For fixed gear, methods can be used to prevent the capture of large animals such as birds and mammals. The successful use of these devices, however, depends on overcoming gear handling constraints and the short-term economic losses often associated with their use. By-catch is just one component of the total mortality of species affected by fishing. Hence by-catch is not an isolated issue. Addressing the problem requires consideration of the broader question of resource management, including the target species. Success in reducing by-catch requires that chronic problems of excessive exploitation must be tackled, and this remains a major challenge worldwide.
When sea-gulls follow trawlers, it is because they know sardines will be thrown into the sea.
Introduction Eric Cantona, an international footballer who played for France and Manchester United, was once asked if he regretted assaulting a member of the crowd who had made racist remarks. He made the enigmatic reply:
The observation puzzled the international press corps, but struck an immediate chord with those who study fisheries. His observation, while curious within the context
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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it was made, neatly summarizes important aspects of the fishing process that have become of considerable topical interest. First, it draws attention to the fact that fishing operations result in a by-product that is usually unwanted and hence discarded. Secondly, the discarded material is of interest to other organisms, birds in Cantona’s example, and hence there is an effect on the ecosystem beyond the direct removal of fish from the sea. This chapter focuses primarily on the first of these effects, though the implications of the second may be at least as important in the long-term exploitation of living marine resources. Unfortunately, the term ‘by-catch’ is not a uniquely defined concept, and experts around the world use it to mean slightly different things. While we should not agonize at length about what has an essentially intuitive meaning, it is necessary to clarify terminology for the purpose of this discussion. Most fishing operations, whether they employ towed or fixed gears, trap organisms that are not the primary target of the fisher. These organisms are commonly referred to as the by-catch. It may include small individuals of a target species, or other species that have little or no commercial value. Frequently, a large fraction of the by-catch is discarded for economic or legal reasons, and this portion is commonly referred to as ‘discards’. However a portion of the by-catch may have some value, and is retained. This portion is often called the ‘incidental catch’. It is not always clear precisely where the distinction lies between incidental and target catches, since the difference is related to the value of the species, and this may change. Notwithstanding this problem, we may summarize the total catch retained by the gear by the simple formulae: Total catch = target catch + by-catch By-catch = incidental catch + discards This chapter focuses more on the discards, since this is most clearly the unwanted part of the catch, and the part that everyone would agree should be minimized. It is also, arguably, the larger of the two elements, but
this clearly will depend very much on the fishery concerned.
Reasons for By-catch and Discarding It is easiest to exemplify the problem of unwanted catch by considering a typical trawl. While other gears may have very different fishing properties, the principles illustrated by trawls are almost universally applicable in terms of discards. A trawl can be thought of as a filtering device, where all the organisms that enter the gear are sorted by the meshes, and some, by virtue of their size or shape, are retained in the net. The fisher can therefore control the characteristics of the catch by adapting the retention properties of the gear and the range of animals placed at risk of capture. The degree of control inevitably is incomplete and is the underlying cause of discarding.
Species of no commercial value While some large, densely schooling species, such as herring, capelin and mackerel, allow a fisher to target a single species, most fish occur in mixed assemblages, many members of which are at risk of capture during the passage of the gear. It is typical in demersal fisheries that a large number of species are taken together in a single haul or set. In temperate waters, this may be only a matter of dozens, while in tropical waters it may amount to hundreds. In these, all too common, circumstances, many species are caught that have no commercial value and are simply discarded. Shrimp fisheries, which take a large by-catch of fish, are one of the largest producers of discards in this category. Generally, the by-catch comprises mostly fish, but there may also be other animals, including reptiles (turtles and snakes), birds and marine mammals. The last-named grouping of higher vertebrates tends to attract special attention because of public interest, though scientifically the issues surrounding these animals are no different from other species.
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Target species: individuals of low economic value Because most fishing gears have imperfect size selection characteristics, they may retain fish that are too small (or perhaps too large) to have any commercial value. Trawls, for example, usually retain a range of sizes below the minimum target size, and these typically are discarded. This category can account for large quantities of discards. A further complication of this effect occurs with ‘high grading’. The problem usually arises when large numbers of the target species are being caught, and limitations of space, processing or quota result in only the higher value individuals being retained. Some fish, regardless of size, may be uneconomic to land. In the case of some fixed gears, such as gill nets, fish may be retained for a long period in the net before the gear is retrieved, during which time the quality of the flesh deteriorates and becomes unfit for sale. In addition, some fish may be damaged or diseased. However, this class of discard is usually of minor importance.
Target species: legal restrictions Management measures can often have a direct effect on discarding behaviour. It is common practice to set a minimum landing size (MLS) for commercial fish, with the intention of discouraging the capture of small, juvenile fish. Where gears retain fish below the MLS, these usually must be discarded, but, even if retained, they are an unwanted part of the catch, generating undesirable additional mortality. Catch quota restrictions can lead to fish being discarded. The most commonly cited reason for this is in mixed fisheries, where a quota for one species is exhausted while fishing continues for other species. When the species are taken together, the stock with no remaining quota is discarded. The extent to which discarding of this type occurs depends very much on the effectiveness of enforcement, since over-quota fish are often landed covertly. Whether or not the over-quota fish
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are discarded, this portion of the catch constitutes undesirable by-catch.
Overview of Global By-catch and Discards An assessment of global fisheries by-catch and discards was prepared by Alverson et al. for the FAO in 1994. This work provides probably the most comprehensive evaluation of the problem and forms the basis of this section of the chapter. Inevitably, data pertaining to fisheries worldwide are incomplete. Furthermore, the nature of discarding means that data collection must be done at sea, which is expensive and usually imprecise. These problems notwithstanding, the mean total weight of discards estimated by Alverson et al. was in the range of 17.9–39.5 million t annually. The most recent estimates of total discards suggest they are in the lower end of this range, perhaps 20 million t (FAO, 1999a). This means that about one-fifth of the total world catch of fish is discarded. Even at face value, this is a large biomass, but, set in the context that a large number of the world’s fish stocks are already overexploited, it is an indication that by-catch is likely to be highly undesirable. Examining the distribution of discards by region (Fig. 13.1) indicates that over half the total world discards occur in only three regions, the Northwest Pacific, Northeast Atlantic and the West Central Pacific. In the Northwest Pacific, discards arise from a variety of fisheries, including crab, shrimp, mackerel, jack mackerel, cod and pollock fisheries. In the Northeast Atlantic, most discards occur in traditional roundfish and flatfish fisheries, while in the West Central Pacific, shrimp fisheries generate most discards. While these regions appear to produce the largest quantity of discards, it should be remembered that they also account for the largest landings and hence do not necessarily represent the highest rate of discarding. Figure 13.2 shows an estimate of the discard rate by region. Clearly, the differences between the discard rates are much smaller than the absolute quantity discarded, with
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Fig. 13.1. The discard weight by world region as a percentage of total discards. Data from Alverson et al. (FAO 1994).
Fig. 13.2. The estimated discard rate, i.e. the weight discarded per unit weight landed, classified by region. Data from Alverson et al. (FAO, 1994).
typical rates between 15 and 35%. It should also be borne in mind that discards have been studied more intensively in some areas, such as the Northeast Atlantic, than in others. Lower discard rates in some areas may simply be a reflection of lack of data. Figure 13.3 shows the discard rate, expressed as the unit weight discarded per unit weight landed, for the 20 fisheries with the highest recorded rates. It is noticeable that 13 of these fisheries are for shrimp, while the remainder are mainly fisheries for demersal
flatfish or roundfish. In fact, about 33% of all discards are produced in shrimp fisheries, and many of these take place in tropical regions. Fishing gear perhaps provides the simplest way of comparing the importance of different fisheries in relation to discards (Fig. 13.4). While gear does not define fisheries, it is a very good indicator of the type of target species. Clearly, shrimp trawls are by far the most important gear in the generation of discards. This arises because most shrimps are small and hence small-mesh nets are used,
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Fig. 13.3.
The 20 fisheries with the highest rate of discarding. Data from Alverson et al. (FAO, 1994).
Fig. 13.4.
The mean discard rate by principal gear type. Data from Alverson et al. (FAO, 1994).
which also retain a large variety of fish that are found in the same habitat. At the other end of the range, pelagic trawls have very low discard rates. These gears generally are used to catch small pelagic fish that are densely schooling and can be caught with very little by-catch. When discarding does occur in these fisheries, it is often the result of so-called
‘slippage’ which is a variant on high grading. When the gear is brought to the surface, the fisher may decide that the size of the fish in the catch is unsatisfactory or the wrong species and simply empty the net in the water without bringing the fish aboard. Herring and mackerel fisheries in the Northeast Atlantic can suffer from this problem.
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Figure 13.5 shows the discard rate associated with different target species groups. It confirms the emerging picture associated with gear, where crustaceans (shrimps, prawns, crabs, etc.) and flatfish are related to very high discard rates, and small pelagics tend to be associated with low rates. Large pelagics and roundfish are intermediate.
The Biological Impact of By-catch Figure 13.6 illustrates the fate of animals encountering a fishing gear, in this case a trawl, but the process would be similar for most kinds of gear. The diagram shows that on encountering a gear, some fish are retained and are taken aboard the vessel. Of these, some are discarded and are returned to the sea. Not all discarded fish die. Some species, notably certain flatfish, can survive the fish capture process quite well (Berghahn et al., 1992) but this is not the norm. For example, Hill and Wassenberg (1990) found that only 1–2% of fish survived, though as much as 50% of crustaceans survived. Some animals that interact with the gear escape, but may nevertheless die, contributing to the overall mortality caused by fishing.
Fig. 13.5. 1994).
Such mortality is sometimes referred to as ‘unaccounted mortality’ (ICES, 1997) because it is rarely considered in the assessment and evaluation of fish stocks. Hence the overall direct by-catch impact of fishing will be due to the deaths of the discards and the deaths related to contact with the gear. However, the dead organisms are available to scavengers, and this is an additional route by which the ecosystem may be affected. The unaccounted mortality due to contact with the gear may be large (Cook, 1998), though few studies have quantified it. Species that are especially fragile are most at risk, and this includes many deep-water fish. As well as the biological impact of this largely unknown mortality, failure to quantify it can result in bias in stock assessments, particularly where evaluations of mesh size changes are concerned. While the total quantity of by-catch or discards may give an impression of the size of the problem, it does not characterize adequately the potential effects on populations or ecosystems. Even if the quantity discarded is large in terms of weight or numbers of individual animals, it still may only represent a small portion of the total population affected. It can therefore be more useful to consider the mortality rate resulting from fishing, rather
The discard rate associated with different target species. Data from Alverson et al. (FAO,
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Fig. 13.6.
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Schematic representation of the fate of fish that enter a fishing net.
than simple quantities alone. Mortality rate is a measure of the proportion of a population killed due to a particular cause. A sum of all the mortalities, such as those caused by fishing, predation and disease, is the total mortality. At a simple level, if the total mortality affecting a population exceeds the rate at which new individuals enter it, then the population will decline. Mortality due to by-catch is just one component of the total mortality and cannot, therefore, be considered in isolation. It may be especially important for a target species that also suffers a by-catch mortality since the marginal additional mortality due to by-catch might be sufficient to render exploitation non-sustainable. It can be useful to distinguish between the by-catch mortalities affecting target and non-target species because the implications for managers may be different. Where a species is taken only as by-catch, then concerns will centre mainly on the conservation and ecosystem consequences of fishing. If, however, a species is also a target species, then by-catch will have additional implications for the management-directed fishery on it. It is also convenient to distinguish between the by-catch of fish and that of higher vertebrates, such as reptiles, birds and mammals, primarily because of the different public perception of these species. Each of these categories is dealt with below.
By-catch mortality: non-target species The fisheries for small crustacea have already been noted for their large production of fish by-catch. Alverson et al. (FAO, 1994) note that the Australian northern prawn fishery discards more than 75 families of fish. Shrimp fisheries are implicated in high mortalities of non-target fish species (Blaber et al., 1990) and are cited as the cause of the reduction of the croakers to very low levels (Cittendon and McEachran, 1975; Tillman, 1992). Where the distribution and vulnerability to fishing of both target and non-target species are similar, it might reasonably be expected that the two categories will suffer comparable fishing mortalities (Pope et al., 2000). Stratoudakis (1997) found that the mortalities of common dab (Limanda limanda) and grey gurnard (Eutrigla gurnardus) were in the same range as the main target species (gadoids). These mortalities are large enough to result in the equivalent of growth overfishing, but without a high risk of extinction. However, some fish species that have low reproductive rates may suffer unsustainable mortalities, and this appears to be the case for the common skate (Raja batis) in the Irish Sea (Brander, 1981). These large skates have low reproductive rates and, like sharks, are particularly vulnerable to even moderate mortalities. The problem is likely to be widespread and is one
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of the principal concerns associated with the mortalities inflicted through by-catch.
By-catch mortality: target species The biological effects of excessive total mortality on target species may be no different from that on non-target species. The significance of the by-catch mortality has more to do with the utilization efficiency of the resource and the extent to which lack of knowledge of such mortalities can undermine the assessment and management of stocks. The by-catch of target species can occur in broadly two ways. First, it may arise when a fishery for one target species takes a by-catch of a species that is the target of another fishery. In a study of Singapore shrimp fisheries, for example, 32% of the by-catch was juveniles of commercially important species (Abdullah et al., 1983). This means that not only does the shrimp fishery remove potential yield from the directed fishery, but also, unless it is taken into account, management of the commercial fish stocks is likely to be misled by bias in scientific assessments. The second category of target species by-catch arises when discards of the target species occur within the target fishery itself. It most commonly affects small individuals of the target species. Once again, killing
juveniles removes potential yield, and, unless taken into account in assessment, will detract from assessments and good management decisions. An example of the magnitude of the lost yield can be seen in Fig. 13.7, which shows how large the gain could be for North Sea haddock if all discards could be avoided. Not only is the yield substantially improved, but the spawning stock biomass also nearly doubles (Shepherd, 1990).
By-catch mortality: reptiles, birds and marine mammals Reptiles, birds and marine mammals are no more a distinct group than fish or marine invertebrates that are affected by fishing. They do differ, however, in the public perception of their importance, especially in the western hemisphere, where conservation issues have tended to overtake resource exploitation considerations for these species. This separates them culturally rather than scientifically, and is the main reason for affording them particular mention. If there is a scientific characteristic that sets them apart, it is that they have low reproductive rates and a late age of maturity, which makes their populations more vulnerable to additional mortalities arising from fishing. Turtles may be taken in almost all types of fishing gear. In many instances, they may be
Fig. 13.7. An example of the effect of discarding on the potential yield of a stock. The figure shows the expected yield per recruit for North Sea haddock if all discarding were avoided.
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released unharmed or taken as food. Shrimp trawls have been identified as a major cause of turtle mortality in the USA (Magnuson et al., 1990) and their threatened or endangered status precipitated the use of turtle-excluding devices (TEDs) in fishing gear. In at least one case, loggerhead turtles, the mortality due to fishing was sufficient to prevent stock recovery. However, the use of excluding devices is increasing and should mitigate the problem. Tasker et al. (2000) review the impact of fishing on marine birds. Both drift nets and longlines have caused mortalities among albatrosses and petrels, particularly in the Southern Ocean and North Pacific, though the use of drift nets has declined. Some albatrosses have very low reproductive rates, and even very small incidental mortalities are enough to threaten some species with extinction. In contrast, another important effect of fishing is the increased availability of food to scavenging species, notably gulls. The production of discards and offal by fishing vessels makes hitherto inaccessible food obtainable to these species, and is linked to population increases in a number of seabirds. Seals and small cetaceans are the main groups involved in mammal incidental catches. For seals, it is catches in drift net, gill net and trawl gear that appear to be involved primarily. Incidental catches of seals have been implicated as the cause of decline in a number of species around the world, including fur seals and sea lions (Woodley and Lavigne, 1991). Despite incidental catches, certain seal populations are increasing, such as grey seals around the British Isles (Hiby et al., 1996), at least partly attributable to a cessation of hunting. Of the marine mammals taken as bycatch, dolphins and porpoises tend to attract the greatest public attention. Most incidental catch is associated with drift nets and gill nets, but tuna purse seines are also involved. Incidental mortalities in excess of about 2% per year are often quoted as not sustainable for small cetacean populations (Perrin et al., 1994). In contrast, typical fishing mortality rates for target fish species are well in excess of 10%, and it is not unusual to observe rates above 50%. This indicates that a very small additional mortality for cetaceans due to
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fishing is enough to endanger their populations. Inevitably, it is very difficult to obtain precise estimates of such mortalities, but those that have been made often indicate rates that are close to unsustainable. Uncertainty in the estimates leads to debate about the detrimental effects of fishing, but it is clear that even a small by-catch of these animals is a potential threat to their long-term persistence. The International Whaling Commission estimated that 13% of 54 populations of small cetaceans studied suffered unsustainable losses due to mortality from passive gears (Perrin et al., 1994). However, in recent years, the use of large pelagic drift nets increasingly has been restricted, and they were banned by the UN in 1991 (resolution 46/215), so some improvement would be expected.
Ecosystem effects The immediate effect of removing individuals from a population is to reduce its size. However, the biology of these populations and the way they interact with one another is complicated. In the long term in an ecosystem, there are likely to be changes contingent on fishing that are hard to predict. Other chapters in this volume discuss the ecosystem effects of fishing in more detail (Chapters 12 and 15), but it is important that the ramifications of the direct effects of fishing are not forgotten. In general, the effects of by-catch, in conjunction with the direct effect on target species, will tend to:
• • •
reduce the abundance of large individuals, particularly predators; increase the relative abundance of smaller, early maturing species with high reproductive rates (Jennings et al., 1999); and favour the increased abundance of scavengers.
It should also be remembered that by-catch mortality alone is not necessarily the main cause of changes to an ecosystem that can be caused by fishing. The effect on target species may be as large or larger. The particular significance of by-catch mortality is that it
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results in additional mortality for species or size classes that are not the intended target, and hence the ecosystem impact of fishing may be more widespread than would result from exploitation of the target species alone.
Drivers and Constraints Affecting By-catch The reasons for discarding have already been discussed, but these reasons do not really explain why by-catch and discarding appear to be such a large element of fishing. It is not immediately apparent, for example, why fishers catch and discard so many juveniles of the target species when this directly reduces long-term yield. In practice, by-catch resulting from the technological constraints imposed by the gear is made worse by economic forces, which drive the process. A substantial by-catch problem in a fishery is usually a symptom of resource overexploitation. This is best illustrated by considering how a fishery on a hitherto lightly exploited stock might develop. In the early stages of the fishery, the target population will comprise large individuals, which tend to realize high market value. The exploitation rate on these fish increases because fishing is profitable, but the survival rate of the fish decreases and, as a result, the number of large fish available to be caught declines. Soon fishers have to target smaller fish so that their profitability can be maintained. The pressure on size continues, resulting in smaller and smaller fish being targeted, until most of the stock comprises fish that are close to the minimum marketable size. In a trawl fishery, gear constraints impose limits on trying to catch all fish of a marketable size without also catching a substantial by-catch of undersized fish. This is illustrated in Fig. 13.8, which shows the proportion of fish retained by the gear at each size (indicated by the sigmoid curve). Some fish of all sizes are caught, but the proportion held by the gear declines with size. In the example illustrated, the minimum marketable size is 40, and all fish below this size are discarded (hatched area).
The upper panel in Fig. 13.8 shows the situation when a fishery is only lightly exploited and larger fish are relatively abundant. In these circumstances, a fishing vessel can obtain adequate reward with a selective net that allows small fish to escape. Some fish above the minimum marketable size escape (horizontal shading), but the losses will be small in relation to the total catch. Furthermore, the fisher avoids the handling nuisance of catching large numbers of small fish of no value. In contrast, the lower panel in Fig. 13.8 illustrates the situation when a stock becomes overexploited. Because so few large fish are in the sea, the fisher adjusts the gear so that more marketable fish are retained. In effect, the selection curve moves to the left. This eliminates the losses of marketable fish seen in the upper panel, but, in doing this, the quantity of fish below the minimum marketable size increases, resulting in more discarding. While the example described refers to a single target stock, similar processes affect the development of by-catch problems with multiple species. The simple process of reducing selectivity to catch smaller fish will also tend to result in greater by-catch of small species that previously would have passed through the gear. In addition, as the target species becomes scarce, the fisher needs to maintain catch rates by extending the range of species taken, and will deploy the gear in areas where more species are available. In doing so, low-value species will tend to be placed at risk, resulting in increased by-catch.
Mitigating the By-catch Problem Reducing overexploitation The previous section emphasizes the economic pressures that magnify the by-catch problem. It is important to recognize that these forces play a major role in attempts to reduce discards and by-catch. While limitations on gear design and operation set constraints on what can be achieved, success
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Fig. 13.8. A diagram illustrating the effect of overexploitation on the capture of small fish. This theoretical example assumes the minimum marketable size is 40. In the upper diagram, because large fish are abundant when exploitation is low, the selectivity curve adopted by fishers tends to capture few small fish, and some marketable fish escape. In the lower diagram, fishers are forced to try to catch all marketable fish and move the selectivity curve to the left. In so doing, large numbers of small fish are caught as by-catch and are discarded.
in trying to operate more selective fishing gears and practices ultimately will depend heavily on the economic forces to which fishers are subject. Nested within this is the degree to which management measures can be enforced. In the simple example described in Fig. 13.8, it emerges that, as exploitation increases, the pressure is to reduce the selectivity of gear
(moving the selection curve to the left), causing an increase in by-catch. In contrast, the traditional management response, in a trawl fishery, for example, is to legislate for improved selectivity, usually by specifying a larger mesh size. Using the example in Fig. 13.8, this would mean moving the selection curve to the right. Thus the fisher faces the opposing forces of short-term economic
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pressure and long-term management pressure. Where enforcement is weak – and this is the more common situation – the short-term economic forces will prevail. If conservation measures are to be introduced to reduce by-catch, it is essential that this be done at the same time as addressing the overexploitation problem. Reducing exploitation improves the survival of fish, leading to a greater abundance of large fish, and hence provides an incentive to adopt more selective gear.
Technical measures There are a wide variety of modifications to fishing gear that can be made to improve selectivity and hence mitigate the problem of by-catch. These are often referred to as technical measures. For trawls, which are one of the least selective gears, the principle essentially is to provide larger holes for the unwanted element of the catch to escape. The most obvious way to do this is to increase the mesh size, but the main drawback is that the conventional diamond mesh of nets may close under tension. Alternatives to mesh size increases are the insertion of panels made with square mesh (Fig. 13.9). Such panels are less susceptible to mesh closure and may be effective for roundfish if located appropriately in the net. They are less
effective for flatfish due to the shape of the mesh opening. A device that has attracted increasing attention is the rigid grid, placed somewhere in the cod end of a trawl. The grid acts as a sorting device (see Fig. 13.10), filtering larger organisms and diverting them to another part of the gear. Such devices either allow the small organisms to be retained while the larger ones escape, or vice versa. These devices are used in some shrimp fisheries in arctic waters to allow the fish component of the catch to escape. The same principle is applied in turtle-excluding devices, as used in shrimp fisheries in the USA. Grids offer a partial means of separating species, but this is based primarily on size. It is possible to sort species by exploiting their particular behaviour. This is done in separator trawls, where a horizontal panel in the net divides those species that try to escape by swimming upwards from those that try to escape downwards. In the Northeast Atlantic, this device can be used to separate haddock from cod (Fig. 13.11). By having separate cod ends for each part of the catch, it is possible to use the different mesh sizes that best suit each species. Static gears, such as gillnets and lines, offer different challenges. Gillnets and drift nets may entrap mammals, such as dolphins, because they cannot be seen either visually or by the animal’s echolocation system. Devices
Fig. 13.9. An example of fitting square mesh panels into a conventional trawl. Because square mesh does not close under tension, the meshes remain open, allowing small, unwanted fish to escape. The effectiveness of the panel may depend on its position. The diagram shows a number of possible positions for the panel.
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Fig. 13.10. Rigid grids can be used to sort fish. This diagram shows one method of selecting out different sizes of fish. A guide panel feeds fish towards the grid. Small fish or shrimps pass through into the upper cod end. Large fish and animals enter the lower cod end. Choosing appropriate mesh sizes for each cod end can be used to minimize by-catch. Similar arrangements can be made to allow turtles to escape.
Fig. 13.11. The principle of a separator trawl. In this example, two species of fish have differing escape responses. The small species tries to escape over the top of the net and is retained in the upper cod end. The large species tries to escape under the net and is retained in the lower cod end. Selection of the appropriate mesh size for each cod end can be used to optimize the catch composition. Such devices have been used to separate cod and haddock in the Northeast Atlantic.
can be attached to these to make the gear acoustically visible, and hence warn the animal of its presence so that avoiding action can be taken. For lines, selectivity can be achieved through choice of hook size and bait. Perhaps the highest profile concern is the ensnaring of seabirds such as albatrosses. Brothers et al. (FAO, 1999b) describe various devices, including setting lines at night when birds are absent, causing the lines to sink more quickly, or trailing streamers that discourage birds from attacking the bait.
Closed areas and seasons It is sometimes possible to identify fishing areas or seasons where by-catch problems are particularly severe. Juvenile fish, for example, may concentrate in nursery areas where shrimp are fished. Closing these areas at certain times to fishing may help to reduce by-catch problems. This is exemplified in Argentina, where a closed box is used to protect juvenile hake that are taken as by-catch in a shrimp fishery. Box closures can
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be simple and effective devices, but their principal drawback is that the closure is rarely complete. Other interests usually dictate that certain fisheries are allowed in the restricted area, and this often provides a loophole that significantly weakens the beneficial effects of the box.
Conclusions Examining estimates of worldwide by-catch reveals no region unaffected by the problem. While there may be differences in total quantities, by-catch rates in each region are comparable. Typically, a fifth of the catch is discarded. The largest single problem appears to be related to shrimp fisheries, though mixed demersal and large pelagic fisheries are also associated with high by-catch levels. There are perhaps two broad management issues related to by-catch. In traditional fishery management, by-catch of target species often represents an inefficient use of the resource and has exercised managers for many years. There is an established theory for quantifying the problem and it is of particular relevance to the fishing industry itself. In recent years, however, questions have been raised not only about the direct effects of fishing on target species, but also about the effects on other species, including those of no direct commercial value. These wider questions, often loosely referred to as ‘ecosystem concerns’, might be considered conservation issues. They appeal to a much wider public than the fishing industry alone. The mobilization of environmental opinion on marine issues means that fishery managers increasingly are obliged to consider conservation issues that may directly affect the economic exploitation of resources. This process of incorporating ecosystem considerations into fishery management is likely to continue, and, in certain areas, may predominate in management policy. The question of by-catch, especially of non-target species, is inextricably linked to these issues. Technical conservation measures, involving modifications to fishing gear or practices, do offer a means of reducing by-catch. These
measures can be shown to be effective, both under controlled conditions and in practical application. Their effectiveness is very much contingent on the penalty a fisher has to pay in operating them. This may be due to difficulty in handling a modified gear, but more commonly there is a short-term economic loss associated with implementation. Successful use of these devices, therefore, depends on overcoming the short-term losses. The disturbance caused by fisheries to populations and ecosystems ultimately is related to the total mortality of organisms affected by fishing. By-catch is just one component of this total mortality and therefore should not be seen as a separate issue in its own right. There are good reasons for supposing that by-catch is exacerbated by heavy exploitation and can be an indication of overfishing. These two observations indicate that addressing the problem of by-catch requires consideration of the broader question of resource management, which includes the target species mortalities. It is unlikely that success in reducing by-catch can be achieved without addressing chronic problems of excessive exploitation, which remains a major challenge worldwide. Indeed, simply achieving more rational exploitation of many target stocks would be a major step toward addressing the wider question of by-catch.
References Abdullah, J., Ismail, W. and Rhimah, W. (1983) Malaysia measures her by-catch problem. Infofish Marketing Digest No. 6/83, 11. Berghahn, R.M., Waltemath, M. and Rijnsdorp, A.D. (1992) Mortality of fish from the bycatch of shrimp vessels in the North Sea. Journal of Applied Icthyology 8, 293–306. Blaber, S.J.M., Brewer, D.T., Kerr, J. and Salini, J.P. (1990) Biomass, catch rates and abundance of demersal fishes, particularly predators of prawns, in a tropical bay in the Gulf of Carpentaria, Australia. Marine Biology 107, 397–408. Brander, K.M. (1981) Disappearance of common skate Raja batis from the Irish Sea. Nature 290, 48–49. Cittendon, M.E., Jr and McEachran, J.D. (1975) Fisheries on the white and brown shrimp
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grounds in the Northwestern Gulf of Mexico. In: Proceedings of the American Fisheries Society, 105th Annual Meeting, Las Vegas, Nevada, 13 September 1975. Cook, R.M. (1998) The Estimation of Mortality due to Passage through Fishing Gears. Annex 3 in the Report of the study group on the use of selectivity and effort measurements in stock assessment. ICES CM 1998/B:6. FAO (1994) A global assessment of fisheries by-catch and discards. Prepared by Alverson, D.L., Freeberg, M.H., Murawski S.A. and Pope, J.G. FAO Fisheries Technical Paper No. 339. FAO (1999a) The State of World Fisheries and Aquaculture 1998. FAO, Rome. FAO (1999b) The incidental catches of seabirds by long-line fisheries: a worldwide review and guidelines for mitigation. Prepared by Brothers, N.P., Cooper, J. and Lokkeborg, S. FAO Fisheries Circular No. 937. Hiby, A.R., Duck, C.D., Thompson, D., Hall, A.J. and Harwood, J. (1996) Seal stocks in Great Britain. NERC News (January), 20–22. Hill, B.J. and Wassenberg, T.J. (1990) Fate of discards from prawn trawlers in Torres Strait. Australian Journal of Marine Science 41, 53–64. ICES (1997) Report of the Study Group on Unaccounted Mortality in Fisheries. ICES CM 1997/B:1. Jennings, S., Greenstreet, S.P.R. and Reynolds, J.D. (1999) Structural changes in an exploited fish community: a consequence of differential fishing effects on species with contrasting life histories. Journal of Animal Ecology 68, 617–627. Magnuson, J.J., Bjorndal, K.A., DuPaul, W.D., Graham, G.L., Owens, D.W., Peterson, C.H.,
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Pritchard, P.C.H., Richardson, J.I., Saul, G.E. and West, C.W. (1990) Decline of Sea Turtles: Causes and Prevention. National Research Council, National Academy of Sciences, Washington, DC. Perrin, W.F., Donovan, G.P. and Barlow, J. (eds) (1994) Gillnets and cetaceans. International Whaling Commission Special Issue No. 15. Pope, J.G., MacDonald, D.S., Daan, N., Reynolds, J.D. and Jennings, S. (2000) Gauging the impact of fishing mortality on non-target species. ICES Journal of Marine Science 57, 689–696. Shepherd, J.G. (1990) Stability and the objectives of fisheries management: the scientific background. (MAFF Directorate of Fisheries Research, Lowestoft.) Laboratory Leaflet No. 64. Stratoudakis, Y. (1997) A study of fish discarded by Scottish demersal fishing vessels. PhD thesis, University of Aberdeen, UK. Tasker, M.L., Camphuysen, C.J., Cooper, J., Garthe, S., Montevecchi, W.A. and Blaber, S.J.M. (2000) The impacts of fishing on marine birds. ICES Journal of Marine Science 57, 531–547. Tillman, M.F. (1992) Bycatch: the issue of the 90’s. In: Proceedings of the International Conference on Shrimp Bycatch Lake Buena Vista, Florida, 24–27 May 1992. Sponsored by the Southeastern Fisheries Association, Tallahassee, Florida. NOAA/NMFS, Tallahassee, Florida, pp. 13–18. Woodley, T.H. and Lavigne, D.M. (1991) Incidental capture of pinnipeds in commercial fishing gear. International Marine Mammal Association Technical Report 91–01.
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The Effects of Fishing on Species and Genetic Diversity Ellen L. Kenchington
Centre for Marine Biodiversity, Bedford Institute of Oceanography, Dartmouth, Nova Scotia, Canada
There are genes that have not changed since the very first single-celled creatures populated the primeval ooze. There are genes that were developed when our ancestors were worm-like. There are genes that must have first appeared when our ancestors were fish. There are genes that exist in their present form only because of recent epidemics of disease. And there are genes that can be used to write the history of human migrations in the last few thousand years. (M. Ridley (1999) reflecting on the human genome)
Abstract The preservation of genetic resources has become an important element of conservation. This overview is meant to provide an understanding of the importance of conserving genetic variation at the level of both species and populations within species. The loss of species in the marine environment is not as extensive as in freshwater or terrestrial systems. However, we have an imperfect knowledge both of the numbers of marine species and of extinction events. New species are still being discovered, even in well-studied areas, while proving that something is no longer there has produced conservative estimates of losses. Extinction of marine mammals and gastropod molluscs has been documented. Of these, overfishing has caused the extinction of the Stellers sea cow and was instrumental in the loss of the Caribbean monk seal. Within species, genetic diversity is partitioned among and within populations. Overfishing is seen as the major threat to the loss of marine populations, while habitat degradation is threatening anadromous, estuarine and freshwater species, and population extinction has been documented. The number of spawning components is a guide to assisting managers in preserving this aspect of within-species diversity, as they often are identifiable in space and time. Certain species, such as herring, have a large number of populations, while others, such as mackerel, have fewer. Fishing can also alter genetic diversity within populations, even when numbers are high. When fishing is highly selective, it has the potential to permanently change the characteristics within a population, usually in directions of less economic value. Removing large fish generally appears to favour slow-growing, early maturing fish. At all three levels of organization, previous paradigms have not stood the test of time. Marine species can become extinct; marine fish have much more genetic structure than previously supposed; and selective fishing can cause heritable differences in yield and life history traits.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Introduction The cells of all living organisms contain a genetic code that directs biochemical processes and development. This code is composed of only four letters, A, G, C and T (or U), which transcribe, in groups of three, into one of 20 different types of amino acids, the so-called ‘building blocks’ of proteins. The nature of a protein can influence various characteristics of an individual, from temperature tolerance to coloration or body size. This translation of the code is the same in all organisms from dandelions to whales and, although established billions of years ago, was only revealed to us in the 1950s. From this basis, the great diversity of living organisms has grown. The genetic code has another distinguishing property, and that is replication. The code is responsible for living organisms producing similar copies of themselves. The code is the reason that children look like their parents or grandparents, and why cod produce other cod. However, most advanced organisms do not produce exact copies of themselves (clones). Instead, the codes provided by the mother and by the father are recombined to produce a unique individual in their offspring. The manner in which this occurs follows the rules of heredity, giving genetic data a unique theoretical base and genealogy, which is very different from all other types of data we collect. Any group within which genetic material can be exchanged between individuals, across generations (a species), represents a unique evolutionary lineage that cannot be replaced once lost. Within it is the story of the generations that have occurred from the pre-Cambrian seas through to historical times. While some would argue that species that have no closely related living relatives are more deserving of conservation than those with a number of close relatives (cf. Vecchione et al., 2000), the fact remains that the combination of genes in any species is unique. The maintenance of genetic diversity in marine species has become an important element of conservation. This overview is meant to provide an understanding of the
importance of conserving genetic variation at the level both of species and also of populations within species. Specifically, the effects that fishing practices have on genetic diversity at the species, population and genetic levels of organization are briefly reviewed and, wherever possible, examples are provided.
‘One Fish, Two Fish, Red Fish, Blue Fish’ – the Loss of Species in the Marine Environment The irrevocability of extinction is a prime motivator for conservationists. Most would agree that the loss of a species is something to avoid, and that humans have a stewardship responsibility in this area. However, species extinction also poses a major threat to biological diversity that reverberates through the levels of species assemblages and ecosystems. Further, when species of commercial interest become extinct, or are reduced to low levels, harvest pressure often is transferred to others with similar traits, magnifying the impact of loss on the system. Increasingly, scientists and economists are aware of the implication of species loss and its effect on biological processes and ecosystem function as well as on human society (Ehrlich and Ehrlich, 1994). Extinction has always been a part of the earth’s history, and fossil records tell us that the average longevity of a species is of the order of 1–10 million years, at least for marine invertebrates (May et al., 1995). Currently, there is an accelerated loss of species due to human activities. It has been suggested that recent extinction rates are 100–1000 times higher than the pre-human levels, with these rates potentially increasing by a factor of 10 in the near future (Pimm et al., 1995). Is species extinction a problem in the sea? One of the difficulties in addressing the issue of species loss in the marine environment is that we do not have an accurate list of the organisms that live there, or indeed, anywhere on earth. To date, about 300,000 marine species have been described globally, although precise estimates are hindered by
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deficiencies in our ability to recognize and name them. This has resulted in misidentification and changes to the numbers as mistakes are corrected (Vecchione et al., 2000). Scientists differ in their estimates of the total number of marine species, including as yet undiscovered organisms; some think that there are only about 800,000 (GESAMP, 1997) while others suspect that there might be several million or more (Malakoff, 1997). This discrepancy is seen to differing degrees in different groups, depending on how well known they are. For example, Grassle and Maciolek (1992) have suggested that the global species list just for the deep-sea invertebrates living in marine sediments might be as high as 10 million (only 100,000 have been described). This has been contested logically by Poore and Wilson (1993) with support from May (1993), who predicted that the numbers will be much lower (5 million and 500,000, respectively). The inconsistency is due in part to the methods of extrapolating discovery rates from newly explored habitats to broader oceanic areas. Overall, our sampling of deep-sea environments has been very sketchy and, when new habitats are first explored, it is expected that new species will be discovered. A recent research programme on Tasmanian seamounts identified about 300 species of fish and invertebrates, with approximately one-third new to science (Koslow et al., 2000). With vast areas, such as the West Indian Ocean, requiring scientific study, it may be some time before we have more precise estimates of the number of species on this planet. However, some parts of the world’s oceans have been studied intensively, and the discovery rate of new species in those areas is lower. For example, the marine waters of Britain and Ireland have been well studied since the 18th century. Of the 331 species of fish in the region listed by 1992, only 17 were described after 1900 (Costello et al., 1996). However, among the smaller forms of life, new discoveries continue even in these relatively well-studied environments (Costello et al., 1996). Those have included the exciting discovery of two new phyla (major groupings of species) of small marine organisms in recent decades. The discovery of phyla is truly
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extraordinary, as the 1.5 million or so species described on the planet can all be classified into only about 43 phyla. In 1983, the Danish zoologist R.M. Kristensen discovered the phylum Loricifera living in the tiny spaces between marine gravel in sediments off the French coast (Kristensen, 1983). Since then, at least ten more species have been described (with rumours of hundreds under investigation!), including inhabitants of coastal areas of North America. Also in 1995, in the Kattegat between Denmark and Sweden, another new phylum (Cycliophora) was discovered living on the mouth bristles of the Norwegian lobster and feeding on scraps of the lobster’s food (Conway Morris, 1995; Funch and Kristensen, 1995). This is not to say that only the smaller forms of life have remained a mystery. As recently as 1991, a new species of beaked whale (Mesoplodon peruvianus) was described from Peru (Reyes et al., 1991), and it is likely that yet others remain to be found. The most famous example of a new discovery is the coelacanth (Latimeria chalumnae), often referred to as the ‘living fossil’ (Fig. 14.1). The coelacanth lineage was thought to have been extinct for 80 million years when a live specimen was brought ashore by a fisherman off the eastern coast of South Africa in 1938. Miraculously, this specimen made it into the hands of people who could recognize it and alert the scientific community. Efforts to discover where the fish came from eventually identified a small population living off the Comoros Islands between Mozambique on the African continent and Madagascar, where it had been caught with enough frequency for it to have been given a name in the native language. Just recently, a second population was discovered from North Sulawesi (Erdmann et al., 1998; Fricke et al., 2000). This second find has since been described as a new species, Latimeria menadoensis, using genetic evidence. In 2000, a third population (perhaps another species?) was verified in the St Lucia Marine Protected Area on the northeast coast of South Africa. The total number of coelacanths may be greater than 100,000 (Hissmann et al., 1998). Clearly there remains much to learn about life in the oceans.
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Fig. 14.1. Maputo Museum photo of a Mozambique female coelacanth before dissection. The giant fish was 178 cm in length and weighed 98 kg. Only one Comoros Island coelacanth, caught in 1960, has rivalled it, at 180 cm and 95 kg. (Photo provided by Robin Stobbs and legend details provided by Jerome Hamlin, and downloaded from www.dinofish.com)
Just as we do not know precisely the number of species in the world’s oceans, neither do we know how many have become extinct due to human activities. Carlton (1993) has claimed that hundreds of marine invertebrates have not been seen since the 18th and 19th centuries (while recognizing that a number of factors other than extinction may account for some of these). There have also been cases where a species has been declared extinct, as with the hydrocoral Millepora boschmai, only to be found at a later date (Carlton, 1993). It is much harder to prove that something is no longer there than to prove that it is, especially when much of the marine environment is not readily accessible. This reality has sparked a debate on ‘burden of proof’ that remains to be resolved. Comparatively, there are many more documented cases of extinction in freshwater and on land than there are in marine environments (cf. Culotta, 1994; Huntsman, 1994; Ryman et al., 1995; Malakoff, 1997; Powles et al., 2000). This has been interpreted by some to reflect a resiliency of marine species to human impacts (cf. Huntsman, 1994; Powles et al., 2000; Hutchings, 2001). However, particular groups of marine organisms do appear to be more vulnerable. Extinction of marine mammals (e.g. Stellers sea cow, Caribbean monk
seal), gastropod molluscs (e.g. eelgrass limpet, Asian periwinkle), anadromous fish (New Zealand grayling) and an Australian red alga (Vanvoorstia bennettiana) have been documented, while several species of skates, sharks, sturgeons, pipefishes, seahorses and groupers, amongst others, are believed to be at risk (Carlton, 1993; IUCN, 1996; Morris et al., 2000; Musick et al., 2000; Powles et al., 2000). Additionally, species such as the Great auk and the Sea mink, which depended upon the marine environment, have also been lost.
The role of fishing in extinction events Of the documented marine species extinctions, fishing is directly accountable only for the loss of the Stellers sea cow, a large, slow moving marine mammal native to the Bering Sea. Within 27 years of discovery in 1741, it was driven to extinction by visiting sea-otter hunters, who used it for food. In the case of the Caribbean monk seal, an aggressive fishery for the skins and oil decimated numbers quickly once Europeans populated the West Indies in the 1600s (Debrot, 2000), though a few individuals survived the exploitation only to be persecuted by
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fishermen. The last sighting of this shy, formerly wide-ranging tropical species, was a small group of seals reported in 1952 (Wing, 1992, as cited in Debrot, 2000). Is it coincidental that marine mammals were the first to be exploited to extinction? Seals, sea cows and whales have particular characteristics that contribute to their vulnerability: (i) they take several years or more to mature; (ii) when they do have offspring, they have only one or two annually; (iii) they are large; and (iv) they spend part of their lives in confined habitats (beaches, ice flows, coastal waters); the last characteristic making them easy to hunt, even when scarce. Typically, for anadromous fish (fish that spend part of their time in freshwater and part in the open ocean) and fish living in estuaries and lagoons, habitat degradation is the most common factor leading to extinction from rivers and watersheds. Habitat changes can occur through the construction of bridges and dams, and through pollution. Because anadromous fish return to confined freshwaters to spawn, they have also been heavily fished. Overfishing in combination with habitat loss and other factors increases the vulnerability of anadromous fish to extinction (Powles et al., 2000). The New Zealand grayling is the only extinct anadromous fish documented to date. It was endemic to New Zealand, where it was abundant when Europeans first colonized in the 1860s. Habitat destruction, overfishing and the introduction of trout led to its extinction within 50 years. Currently, the sturgeons (Acipenser spp.) are the most vulnerable group of anadromous fish in North America, with the shortnose and white sturgeon listed as endangered by the American Fisheries Society (Musick et al., 2000). There are 27 species of sturgeon (some are entirely freshwater) and all are listed as either endangered or threatened by CITES. Sturgeons also have special characteristics which render them vulnerable to extinction: (i) very low productivity; (ii) specialized habitat requirements; and (iii) large size – at approximately 6 m in length, the beluga sturgeon is the largest freshwater fish. Amongst the purely marine fish, sharks, rays, jewfish, groupers, Pacific rockfish,
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swordfish and marlins share one or more of the following characteristics: (i) large body size; (ii) naturally lower numbers; (iii) relatively long life spans; (iv) late maturation and reproduction; and (v) occupation of coastal waters with humans. Many species of shark, including the great white shark, are listed by the IUCN as vulnerable to extinction, while 37 species of groupers are considered threatened (Morris et al., 2000). Other marine species are both rare and have localized distributions, rendering them vulnerable to habitat deterioration and overfishing (incidental fishing in the case of non-commercial species). Swaby and Potts (1990) classified the distribution of 165 rare marine fish in Great Britain and found that more than one-third had restricted distributions, which often were clearly defined by habitat boundaries. In contrast, other species have very large population sizes and occupy large areas of ocean, lay large numbers of small eggs, and have eggs that may drift long distances before hatching. Cod, haddock, small tunas, anchovy and mackerel fit this description, as do many invertebrates. These species are less likely to become extinct as the greatest threat facing them, overharvesting, is likely to be abetted for economic reasons at relatively high abundance. To date, there are no documented extinctions of marine fish or invertebrates due to fishing (McKinney, 1997). However, Powles et al. (2000) recently concluded, ‘there is no reason to assume that extinction is not a potential problem in the sea’.
The Loss of Populations and the Impact on Genetic Diversity Long before a species declines into extinction, it will suffer a reduction in the level of genetic diversity within and among its populations. Often range contraction and fragmentation of former distributions occur (e.g. Musick et al., 2000). Fragmentation causes the formation of small isolated populations, which are more vulnerable to genetic degradation. While documented extinctions of marine species are rare, the extinction of populations (extirpation) is more common (Musick, 1998;
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Musick et al., 2000). The salmon of the Pacific (chinook, sockeye, coho, pink and chum) have numerous populations, and 106 have become extinct on the west coast of the USA (Nehlsen et al., 1991), with another 142 extirpated in Canada (Slaney et al., 1996). Levin and Schiewe (2001) estimate that, over the last 30 years, salmon have been extirpated from 40% of their former range in the Pacific Northwest. The Adriatic Sea stock of the beluga sturgeon is now extinct due to exploitation. Amongst mammals, the Atlantic Ocean population of the gray whale has been hunted to extinction, and the Gulf of St Lawrence walrus population is extirpated (cf. Powles et al., 2000). As many as 82 species of marine fish in North America alone have populations that are believed to be under some level of extirpation risk (Musick et al., 2000).
Principles of genetic variation A gene is part of the genetic code responsible for producing a specific trait. The four letters of the genetic code are found in 64 combinations, although there are only 20 different types of amino acids. Different combinations of the code can produce the same amino acid, and when this happens the same protein is produced despite the differences in the code. Conversely, changes to the gene code may result in different proteins being produced, potentially creating differences in performance or physical traits. Such alternative variants of a specific gene are called alleles. The number of different alleles is a measure of genetic variation or diversity. The average effect of substituting one allele for another is referred to as ‘additive genetic variance.’ The genetic variation of a species is distributed both within populations, expressed as differences between individuals, and between populations, expressed as differences in the presence and frequency of alleles. When different populations arise, with little or no connection between them, they become genetically different from one another. Loss of such populations thus results in loss of genetic diversity within the species. The
existence of multiple spawning units is an indicator that populations may be isolated reproductively, particularly if the species shows spawning site fidelity or homing abilities. The organization of these populations in time and space, along with the ratio of within- and among-population variation are important to maintain in order to avoid negative genetic effects (Altukhov and Salmnekova, 1994). Among populations, genetic diversity can also be lost when populations that are not normally in contact with one another hybridize. This can occur when physical barriers are removed, when fish are introduced to an area or escape, or when migration patterns change due to environmental conditions. When populations are isolated for long periods, closely linked genes may evolve to work well with one another. Subsequent hybridization of such populations may result in these gene complexes breaking down, resulting in a weakening of the population over subsequent generations (Hindar et al., 1991). This phenomenon has been documented in Atlantic salmon (Ståhl, 1981; Emlen, 1991). There are also numerous cases of hybridization between closely related species occurring through introductions, resulting in a loss of diversity, particularly in freshwater (Harrison and Stiassny, 1999). Why should we care about the genetic structure of a species? Natural selection (‘survival of the fittest’) will act within populations (for an example of this effect acting on grayling, see Haugen and Vøllestad, 2000) while the genetic potential of the species to adapt to environmental change depends on the total genetic diversity represented among populations. Therefore, it is important to optimize both types of variation to maintain the full potential for evolutionary change within a species.
Identifying patterns of genetic diversity One of the challenges of conserving genetic diversity is determining how genetic variation is distributed within a species. Total genetic variation within a species can be
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partitioned into variation within and among populations, and at various other geographic levels, such as between oceanographic regions. At one extreme, there are fish such as salmon with relatively high genetic diversity between populations resulting from their homing to specific rivers at spawning time (e.g. Gharrett and Smoker, 1993). At the other extreme are species such as the great scallop (Pecten maximus), which have only 2% of their genetic variability between ‘beds’ and 98% between individuals within ‘beds’ (Heipel et al., 1998). It has been shown that, on average, marine fish have about 6% of the genetic variation distributed among populations, while anadromous species have 11% and freshwater 22% (Ward et al., 1994). These percentages are based on portions of the code that are believed to be ‘neutral’, i.e. not under selection, thus reflecting gene exchange between populations. Morphological variation and variation in genes under selection may be partitioned differently according to the environment. Often the genetic structure of a species is not what would be expected intuitively. A growing number of studies are revealing genetic structuring among populations that previously were thought to have been homogeneous (e.g. Merkouris et al., 1998; Shaklee and Bentzen, 1998; Ruzzante et al., 1999; Aubert and Lightner, 2000). Even amongst species with large population sizes and wide distributions, the presence of genetically distinctive populations has been revealed (Mork et al., 1985; Jörstad et al., 1991; Ruzzante et al., 1999; Shaw et al., 1999) and mechanisms have been proposed to explain them (Hedgecock, 1994a,b; Larson and Julian, 1999). Recently, our view of coral reef fishes was altered when damselfish were shown to have a proportion of offspring settling locally, contrary to previous concepts (Jones et al., 1999). Similarly, it was commonly believed that all European eel (Anguilla anguilla) migrate to the Sargasso Sea for reproduction and comprise a single large population. Wirth and Bernatchez (2001) have shown, using genetic tools, that there is significant population
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differentiation among eels, and ‘the reproductive biology of the European eel must now be reconsidered’.
Examples of population structuring in marine fish Atlantic herring is a fish that is characterized by a high level of biological complexity and population richness (Sinclair and Iles, 1988). In the Gulf of Maine and Scotian Shelf areas of the Western Atlantic, herring aggregate for a few weeks to spawn in a number of discrete locations, at different times of the year from April to November. Tagging studies have shown considerable migration and intermixing of spawning groups at other times of the year. However, spawners return to the same spawning location year after year, perhaps to their location of birth (Stephenson, 1991). Genetic analysis of these populations has shown significant population structure (McPherson et al., 2001). Equivalent results were found in the Pacific herring, which has a similar stock complexity (O’Connell et al., 1998). Cod has a lower degree of stock structure than herring. Nevertheless, on the eastern Scotian Shelf, at least four major groups have been identified, including spring and autumn spawning components (Fig. 14.2; DFO, 1998). In the northern cod stock complex, the collapse of the fisheries involved the loss of many spawning components (Fig. 14.3). Genetic analyses have identified discrete populations within the complex (Bentzen et al., 1996; Ruzzante et al., 1998) and suggest that bathymetry and oceanographic processes may govern stock complexity. For many groundfish in Atlantic Canada, it is necessary to reconstruct a picture of the natural population structure. Haddock, in particular, have undergone an extensive range contraction over the last century (Kenchington, 1996) and now occupy only a portion of their former range, and cusk have undergone both range contraction and fragmentation since at least 1970 (Fig. 14.4; DFO, 2000). In both, there may have been a parallel loss of genetic diversity.
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Fig. 14.2. September/October 1996 longline survey catches of cod (kg) in NAFO Divisions 4VsW. The cod resource on the eastern Scotian Shelf is a complex of spawning components including at least two major offshore groups and a chain of coastal spawning groups. In several of the spawning components (Sable/Western offshore and various inshore areas), there are also spring and autumn spawning fish (DFO, 1998).
Management Clearly, information on the genetic composition and structure of a species has important consequences for fisheries management. It is very important that the management area coincides with the population (Stephenson and Kenchington, 2000). The management of the blacklip abalone in Australia is an example where genetic information has supported established management areas and provided insight into the biology of the species. Partitioning the genetic variance showed concordance with each of the Victorian State abalone management zones and also showed the Port Phillip Bay stocks to be genetically separated from the stocks along the open coast (Hanna and Huang, 1995). Further, the data revealed evidence of limited larval dispersal and inbreeding in some populations,
alerting managers to their vulnerability to overexploitation. Genetic data can also be discordant with management boundaries. For example, Lankford et al. (1999) have shown that while the Atlantic croaker (Micropogonias undulatus) exhibits weak large-scale regional genetic differences between the Atlantic and Gulf of Mexico, there is no genetic differentiation among the Atlantic localities studied, including those north and south of Cape Hatteras, which previously was seen as a stock boundary. Spawning components are often identifiable in space and time, thus allowing them to be managed separately. When genetic data are not available, maintaining viable population sizes of the full range of these spawning components is a precautionary management approach to preserving genetic diversity.
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Fig. 14.3. Erosion of spawning components and severe range contraction in northern cod stock complex of NAFO Divisions 2J3KL. (a) Location of spawning stocks during the period 1948–1992. (b) Location of spawning cod during the period 1990–1993. (Figure copied with permission from Frank and Brickman, 2001, and originally adapted from Hutchings et al., 1993 (a) and deYoung and Rose 1993 (b).)
The Loss of Within-population Genetic Diversity A population can only acquire new genes either through the immigration of individuals from surrounding populations or through the process of random mistakes (mutation) of the code (Fig. 14.5). Apart from these two processes, the genetic diversity of a population is determined by its size and the selective forces acting on it. In general, mutations occur rarely, although there are cases where high levels of pollution or radiation have been shown to increase mutation rates. From an evolutionary perspective (tens of thousands of years) the process of mutation is the only way in which genetic variability is created, and without mutations there would be no biological diversity. Natural selection can maintain or deplete genetic variation (Fig. 14.5). Through selection, certain allelic combinations leave more offspring than others because they are better suited to the environment. When selection acts to destroy individuals with undesirable
alleles, or causes one allele to become fixed in all individuals, it becomes a negative influence on genetic diversity. In situations where individuals inheriting different forms of the same gene from their parents (one from each) are favoured, selection causes genetic variation to be maintained. A second process acting to reduce genetic variance is called genetic drift. In all small populations, the frequencies of particular alleles change randomly from one generation to the next. By chance, some of the alleles of the parents may not be passed on to their offspring. The smaller the population, the more dramatic the fluctuation of allele frequencies, and the faster the loss of genetic variation (Franklin, 1980). Small populations are also vulnerable to inbreeding, i.e. production of offspring from matings between close relatives. If a population is small and isolated, inbreeding is inevitable. In many species, inbreeding is coupled with reduced viability and reproduction, as well as increased occurrences of diseases and defects. Physiological and behavioural phenomena, referred to as allele effects (cf. Frank
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Fig. 14.4. Decadal-scale range contraction and fragmentation from 1970 to 2000 in cusk (Brosme brosme) on the Scotian Shelf, off the coast of Nova Scotia, Canada. Data are from the summer research vessel surveys conducted by the Department of Fisheries and Oceans, Canada, Maritimes Region (DFO, 2000).
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Fig. 14.5. Genetic processes that affect genetic diversity. Processes under the dark arrow act to increase genetic diversity, while processes above the white arrow either maintain or decrease genetic diversity. Population size is a key factor that determines which of these forces has the greater influence. Small population sizes facilitate the loss of genetic variation.
and Brickman, 2000), may also come into play in small populations, resulting in decreased reproduction. Population geneticists have tools available to them to calculate the degree to which these processes are operating within a species. Fisheries scientists have focused primarily on the number of fish or their weight, but, to understand the genetic structure, other factors must be considered. These include the sex ratio, whether different pairs produce different numbers of offspring, and whether the population has gone through a series of crashes previously or not. For example, for some species, genetic variation will be reduced if the sex ratio of breeders departs from 1:1. It is much better genetically to have a population of 50 males and 50 females than to have one of 10 males and 90 females, yet both have a total of 100 breeders. Similarly, the maximum genetic variation is produced in the population when all mating pairs produce equal sized families. In the case of the northern elephant seal, dominant bulls establish a harem and monopolize females, skewing the sex ratio through mating behaviour. This species was also hunted extensively in the 19th century and it is believed that the current population of over 120,000 arose from as few as ten animals in the 1890s (Hoelzel, 1999). As expected, current levels of genetic variation are low.
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How many fish do we have to leave behind? Recent theoretical work suggests that population sizes of the order of 1000–5000 spawners are required for long-term viability (Lynch and Lande, 1998). Previously suggested values of 500 or more breeders (Frankel and Soulé, 1981) appear to have been underestimated (but see Whitlock, 2000). These numbers seem small, but remember that these are successful breeders, i.e. fish that have offspring that survive to reproduce. We do not know how many fish in a spawning aggregation successfully mate or how many of their offspring survive to maturity. If at the same time we are fishing one sex over another (so that the sex ratio is not equal) or if the species is known to have a history of population crashes, we have to leave even more fish. Therefore, we have to hedge against these numbers, perhaps to an order of magnitude or more.
Fishing and genetic diversity The present rate of environmental change in marine ecosystems increases the importance of genetic variability in natural animal and plant populations to respond to these changes. At the same time, human activities may be reducing the genetic variability of these populations. New alleles are created over tens of thousands of years, while loss of alleles and allelic combinations can occur extremely rapidly, even within a single generation. Thorpe et al. (1995) reviewed the relative risks of various aspects of pollution, wild harvesting and aquaculture on genetic processes. Overharvesting and selective harvesting were described as having a high degree of impact. These can be affecting not only the target species but also by-catch and benthic organisms killed or disturbed by gear. For example, Hector’s dolphin, a small marine mammal native to New Zealand, has been reduced to about 4000 individuals and the species has undergone documented decline and range contraction. Recently, a time series of genetic diversity was constructed, extending from 1870, when the first specimens were preserved, to the present. The authors found
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a significant decline in genetic diversity and predicted that one of the two remaining populations will be extinct, and that the other will have lost all of its diversity in the genes examined in less than 20 years (Pichler et al., 1998; Pichler and Baker, 2000). Entanglement in gillnets since 1970 is the primary cause of decline, resulting in the loss of genetic variation (Dawson, 1991; Martien et al., 1999).
Selective fishing Fishing is a highly selective process acting on the size of the organism captured as well as the location of the populations (Law, 2000). The fishery may also favour capture of one sex over another, altering the sex ratio or sex-specific size frequency, or both, of the breeding population. Fish that have extensive ranges and undergo migration may be under different selection pressures in different parts of their range due to different fishing methods or regulations. In the extreme case of anadromous fish, these locations can be very different habitats (marine and freshwater). Fishing therefore has the potential to affect the genetic diversity and genetic structure of a species. The following text draws heavily on the excellent reviews of this topic by Law (2000) and Smith (1999), and readers are directed to those papers for a more comprehensive treatment. When a change in the physical characteristics of fish is observed, it is difficult to determine whether that change is caused by fishing or by other factors. This is because the environment can modify the expression of physical attributes, and fish, perhaps because they are cold blooded, are strongly influenced by environmental conditions. The physical characters we see are the result of three influences: (i) genetics; (ii) environment; and (iii) the genetic–environment interaction. Determining the relative contribution of each can only be done experimentally. Many changes in the characteristics of fish populations that appear to have been brought about by fishing have been described. It is not as easy to determine whether these changes are permanent, due to the loss of alleles, or if they are the result of fishing selection changing the frequency of alleles.
Changes in the characteristics of fish over time are well documented for a number of exploited species. Examples of traits that are thought to be under fishing selection include the following:
• • • • • • •
weight-at-age length-at-age age-at-maturity length-at-maturity spawning season number of eggs (salmon) size of eggs (salmon)
some of which are correlated (e.g. Rijnsdorp, 1993; Rowell, 1993; Millner and Whiting, 1996; Trippel et al., 1997; Smith, 1999; Law, 2000). A common feature among the groundfish of the Northwest Atlantic has been a decline in the mean size-at-age of fish over a period of 30 years (Smith, 1999). Figure 14.6 shows the dramatic change in size-at-age seen in haddock on the eastern Scotian Shelf in three ages (3, 5 and 7 years) from 1970 to 2000. During the 1970s, 7-year-old haddock were on average 60 cm long. By the 1990s, the mean size had fallen to about 40 cm. These changes initially occurred when numbers of fish were high and appear to be independent of abundance (Fig. 14.6). Similar trends are seen in other haddock stocks and in pollock (Frank et al., 1997; DFO, 2000). Fishing practices that remove the latematuring fish may also cause changes to the population (Smith, 1994). Individuals that mature early are able to reproduce at least once before capture and so perpetuate genes for this trait. The majority of Northwest Atlantic cod, pollock and haddock stocks have shown a decline in both age and length at sexual maturity since the 1970s (Trippel et al., 1997). The largest change was seen in the northern cod stock complex, which showed a 20% decrease in size-at-maturity during the early 1990s (Fig. 14.7; Trippel et al., 1997). Similar changes have been observed in the North Sea and in Pacific salmon, though in some cases (e.g. North Sea sole) the trend has been toward larger sizes (Law, 2000). In general, by removing the larger fish from a population consistently over time (30 or more years for some fish), the remaining
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Fig. 14.6. Changes in the mean length-at-age (3, 5 and 7 years) of haddock on the Eastern Scotian Shelf, and the mean number of fish per tow collected in the summer research vessel surveys from 1970 to 2000. (Figures and illustration provided by Dr K. Frank, DFO, Canada.)
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Fig. 14.7. Changes in the median age and length-at-50% maturity (in cm) of male and female northern cod collected from NAFO Subdivision 3Ps (1972–1994). Data are extracted from Trippel et al. (1997).
fish are slow-growing and early-maturing (Smith, 1999). These observations may be due to genetic changes induced by fishing, to environmental change or to elements of both. Temperature is known to be highly correlated with weight-at-age of cod (Brander, 1995), and changes may be attributed to changes in the ecosystem, including the abundance of prey and the degree of competition for food.
For genetic diversity to be affected by selective fishing, there must be a genetic difference between the fish caught and those left behind in the population (cf. Law, 2000). Selective breeding programmes for cultured fish (e.g. salmon; Jonasson et al., 1999) and invertebrates (e.g. scallops; Perez and Alfonsi, 1999) have shown that significant amounts of genetic heritability (the proportion of
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variation in a trait that is inherited from one generation to the next) exist for a number of characteristics, including yield-related traits important to fisheries (Law, 2000). This is the basis upon which aquaculturists, breeding from the largest or fastest growing fish each generation, can increase their production. Generally, characteristics such as the number or size of eggs have been found to be heavily influenced by environmental conditions. However, Smoker and colleagues (2000) recently have shown a very high degree of heritability in both the number and size of pink salmon eggs produced, which was seen as important to the persistence of populations in fluctuating environments. This is one of the few examples where heritability estimates have been determined from field studies. Collectively, this body of research has demonstrated clearly that there is a significant amount of genetic control over those traits selected for by fishing (Law, 2000). Are these changes permanent or reversible? The persistence of fishing-induced genetic changes will depend upon the other selective forces operating on the species, the proportion of genetic diversity affected and the reproductive biology of the species. In some cases, altering fishing practices may not readily reverse genetic change (Law and Grey, 1989). Consequently, fishing may be able to cause substantial evolution in traits of the exploited species (Law and Rowell, 1993) although the time scale over which it operates is unknown. In his recent review, Law (2000) concluded: There can be no question that fishing causes evolution of phenotypic traits of fish; the existence of additive genetic variation has been demonstrated beyond reasonable doubt, and directional selection pressures on this variation caused by fishing are substantial.
simple: maintain all spawning components (unless genetic information is available to support aggregation of units) at large enough population sizes to avoid unacceptable loss of genetic variability. However, even if population sizes are large, selection can result in the loss of alleles. Avoidance of fishery-induced selection is more challenging, but it may be useful to consider altering the size at first capture randomly by year to avoid strong selection for early maturing and slow growing fish. It is also important to note that while some stocks have made remarkable recoveries (including many whale species) they have done so from a narrow genetic base. This means that while everything may look good numerically, the population may have problems coping with environmental change or disease. Is it coincidental that the Caspian seal population fell to very low numbers in the late 1950s, recovered to a level of about 400,000, only to be hit with canine distemper and other viruses (cf. Stone, 2000)?
Acknowledgements I thank Drs Mike Sinclair and Trevor Kenchington for reviewing the first draft of this manuscript and offering helpful references and direction, and Drs Patrick O’Reilly, Trevor Kenchington, Howard Powles and Ken Frank for commenting on the penultimate draft. I also thank Dr Ken Frank for directing me to suitable figures to illustrate the points I wished to make, and Arran McPherson for offering references and unpublished data. Many of the ideas in this chapter were inspired by discussion within the ICES Working Group on the Application of Genetics to Fisheries and Mariculture.
References
Conclusions For many marine species, fishing probably has the greatest impact of any human activity on the loss of within-species diversity, both within and among populations. The key to avoiding loss of genetic diversity is quite
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shrimp Penaeus stylirostris of the Gulf of California, Mexico. Marine Biology 137, 875–885. Bentzen, P., Taggart, C.T., Ruzzante, D. and Cook, D. (1996) Microsatellite polymorphism and the population structure of Atlantic cod (Gadus morhua) in the northwest Atlantic. Canadian Journal of Fishery and Aquatic Science 53, 2706–2721. Brander, K.M. (1995) The effect of temperature on growth of Atlantic cod (Gadus morhua L.). ICES Journal of Marine Science 52, 1–10. Carlton, J.T. (1993) Neoextinctions in marine invertebrates. American Zoologist 33, 499–507. Conway Morris, S. (1995) New phylum from the lobster’s lips. Nature 378, 661–662. Costello, M.J., Emblow, C.S. and Picton, B.E. (1996) Long-term trends in the discovery of marine species new to science which occur in Britain and Ireland. Journal of the Marine Biology Association of the UK 76, 255–257. Culotta, E. (1994) Is marine biodiversity at risk? Science 263, 918–920. Dawson, S.M. (1991) Incidental catch of Hector’s dolphins in inshore gillnets. Marine Mammal Science 7, 283–295. Debrot, A.O. (2000) A review of records of the extinct West Indian monk seal, Monachus tropicalis (Carnivora: Phocidae), for the Netherlands Antilles. Marine Mammal Science 16, 834–837. DeYoung, B. and Rose, G.A. (1993) On recruitment and distribution of Atlantic cod (Gadus morhua) off Newfoundland. Canadian Journal of Fishery and Aquatic Science 50, 2729–2741. DFO (Department of Fisheries and Oceans, Canada) (1998) Eastern Scotian Shelf Cod. Stock Status Report A3–03. DFO (Department of Fisheries and Oceans, Canada) (2000) Updates on Selected Scotian Shelf Groundfish Stocks in 2000. Stock Status Report A3–35. Ehrlich, P.R. and Ehrlich, A.H. (1994) The value of biodiversity. Ambio 23, 219–226. Emlen, J.M. (1991) Heterosis and outbreeding depression: a multi-locus model and an application to salmon production. Fisheries Research 12, 187–212. Erdmann, M.V., Caldwell, R.L. and Kasim Moosa, M.K. (1998) Indonesian ‘king of the sea’ discovered. Nature 395, 24. Frank, K.T. and Brickman, D. (2000) Allee effects and compensatory population dynamics within a stock comple. Canadian Journal of Fishery and Aquatic Science 57, 513–517. Frank, K.T. and Brickman, D. (2001) Contemporary management issues confronting fisheries science. Journal of Sea Research 45, 173–187.
Frank, K.T., Mohn, R.K. and Simon, J.E. (1997) Assessment of 4TVW Haddock in 1996. DFO CSAS Research Document 97/107. Frankel, O.H. and Soulé, M.E. (1981) Conservation and Evolution. Cambridge University Press, Cambridge. Franklin, I.R. (1980) Evolutionary change in small populations. In: Soulé, M.E. and Wilcox, B.A. (eds) Conservation Biology: an Evolutionary– Ecological Perspective. Sinauer Associates, Sunderland, Massachusetts. Fricke, H., Hissmann, K., Scauer, J., Erdmann, M., Moosa, M.K. and Plante, R. (2000) Biogeography of the Indonesian coelacanths. Nature 403, 38. Funch, P. and Kristensen, R.M. (1995) Cycliophora is a new phylum with affinities to Entoprocta and Ectoprocta. Nature 378, 711–714. GESAMP (IMO/FAO/UNESCO-IOC/WMO/ WHO/IAEA/UN/UNEP Joint Group of Experts on Scientific Aspects of Marine Environmental Protection) (1997) Marine Biodiversity: Patterns, Threats and Conservation Needs. Report Study GESAMP No. 62. Gharrett, A.J. and Smoker, W.W. (1993) A perspective on the adaptive importance of genetic infrastructure in salmon populations to ocean ranching in Alaska. Fisheries Research 18, 45–58. Grassle, J.F. and Maciolek, N.J. (1992) Deep-sea species richness: regional and local diversity estimates from quantitative bottom samples. American Naturalist 139, 313–341. Hanna, P.J. and Huang, B. (1995) Application of Molecular Biology to Management of the Abalone Fishery. FRDC Report, 95/002. Harrison, I.J. and Stiassny, M.L.J. (1999) The quiet crisis: a preliminary listing of the freshwater fishes of the world that are extinct or ‘missing in action’. In: MacPhee, R. (ed.) Extinctions in Near Time. Kluwer Academic/Plenum Press, New York, pp. 271–333. Haugen, T.O. and Vøllestad, L.A. (2000) Population differences in early life-history traits in grayling. Journal of Evolution Biology 13, 897–905. Hedgecock, D. (1994a) Does variance in reproductive success limit effective population sizes of marine organisms? In: Beaumont, A.R. (ed.) Genetics and Evolution of Aquatic Organisms. Chapman and Hall, London, pp. 122–134. Hedgecock, D. (1994b) Temporal and spatial genetic structure of marine animal populations in the California Current. California Cooperative Oceanic Fisheries Investigations Reports 35, 73–81. Heipel, D.A., Bishop, J.D.D., Brand, A.R. and Thorpe, J.P. (1998) Population genetic
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differentiation of the great scallop (Pecten maximus) in western Britain investigated by randomly amplified polymorphic DNA. Marine Ecology Progress Series 162, 163–171. Hindar, K., Ryman, N. and Utter, F. (1991) Genetic effects of aquaculture on natural fish populations. Aquaculture 98, 259–261. Hissmann, K., Fricke, H. and Scauer, J. (1998) Population monitoring of the coelacanth (Letimeria chalumnae). Conservation Biology 12, 1–8. Hoelzel, A.R. (1999) Impact of population bottlenecks on genetic variation and the importance of life-history: a case study of the northern elephant seal. Biological Journal of the Linnean Society 68, 23–39. Huntsman, G.R. (1994) Endangered marine finfish: neglected resources or beasts of fiction? Fisheries 19, 8–15. Hutchings, J. (2001) Conservation biology of marine fishes: perceptions and caveats regarding assignment of extinction risk. Canadian Journal of Fisheries and Aquatic Science 58, 108–121. Hutchings, J.A., Myers, R.A. and Lilly, G.R. (1993) Geographic variation in the spawning of Atlantic cod, Gadus morhua, in the Northwest Atlantic. Canadian Journal of Fishery and Aquatic Science 50, 2457–2467. IUCN (World Conservation Union) (1996) 1996 IUCN Red List of Threatened Animals. IUCN, Gland, Switzerland, and Cambridge, UK. Jonasson, J., Stefansson, S.E., Gudnason, A. and Steinarsson (1999) Genetic variation for survival and shell length of cultured red abalone (Haliotis rufescens) in Iceland. Journal of Shellfish Research 18, 621–625. Jones, G.P., Milicich, M.J., Emslie, M.J. and Lunow, C. (1999) Self-recruitment in a coral reef fish population. Nature 402, 802–804. Jörstad, K.E., King, D.P.F. and Nævdal, G. (1991) Population structure of Atlantic herring, Clupea harengus L. Journal of Fish Biology 39 (Supplement A), 43–52. Kenchington, T.J. (1996) Long-term stability and change in the commercial groundfish longline fishing grounds of the northwest Atlantic. Fisheries Research 25, 139–154. Koslow, J.A., Boechlert, G.W., Gordon, J.D.M., Haedrich, R.L., Lorance, P. and Parin, N. (2000) Continental slope and deep-sea fisheries: implications for a fragile ecosystem. ICES Journal of Marine Science 57, 548–557. Kristensen, R.M. (1983) Loricifera, a new phylum with aschelminthes characters from the meiobenthos. Zeitschrift fur Zoologische Systematik und Evolutionforschung 21, 163–180.
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Lankford, T.E., Jr, Targett, T.E. and Gaffney, P.M. (1999) Mitochondrial DNA analysis of population structure in the Atlantic croaker, Micropogonias undulatus (Perciformes: Sciaenidae). Fisheries Bulletin 97, 884–890. Larson, R.J. and Julian, R.M. (1999) Spatial and temporal genetic patchiness in marine populations and their implications for fisheries management. CalCOFI Reports 40, 94–99. Law, R. (2000) Fishing, selection, and phenotypic evolution. ICES Journal of Marine Science 57, 659–668. Law, R. and Grey, D.R. (1989) Evolution of yields from populations with age-specific cropping. Evolutionary Ecology 3, 343–359. Law, R. and Rowell, C.A. (1993) Cohort-structured populations, selection responses, and exploitation of the North Sea cod. In: Stokes, T.K., McGlade, J.M. and Law, R. (eds) The Exploitation of Living Resources. Lecture Notes in Biomathematics, No. 99. Springer-Verlag, Berlin. Levin, P. and Schiewe, M. (2001) Preserving salmon biodiversity. American Scientist 89 (May–June issue). Lynch, M. and Lande, R. (1998) The critical effective size for a genetically secure population. Animal Conservation 1, 70–72. McKinney, M.L. (1997) Extinction vulnerability and selectivity: combining ecological and paleontological views. Annual Review of Ecology and Systematics 28, 495–516. McPherson, A.A., Stephenson, R.L., O’Reilly, P.T., Jones, M.W. and Tagart, C.T. (2001) Genetic diversity of coastal northwest Atlantic herring populations: implications for management. Journal of Fish Biology (Supplement) 59a, 356–370. Malakoff, D. (1997) Extinction on the high seas. Science 277, 486–488. Martien, K.K., Taylor, B.L., Slooten, E. and Dawson, S. (1999) A sensitivity analysis to guide research and management for Hector’s dolphin. Biological Conservation 90, 183–191. May, R.M. (1993) Reply to Poore and Wilson. Nature 316, 44–49. May, R.M., Lawton, J.H. and Stork, N.E. (1995) Assessing Extinction rates. In: Lawton, J.H. and May, R.M. (eds) Extinction Rates. Oxford University Press, Oxford. Merkouris, S.E., Seeb, L.W. and Murphy, M.C. (1998) Low levels of genetic diversity in highly exploited populations of Alaskan Tanner crabs, Chionecetes bairdi, and Alaskan and Atlantic snow crabs, C. opilio. Fishery Bulletin 95, 525–537.
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Millner, R.S. and Whiting, C.L. (1996) Long-term changes in growth and population abundance of sole in the North Sea from 1940 to the present. ICES Journal of Marine Science 53, 1185–1195. Mork, J., Ryman, N., Ståhl, G., Utter, F. and Sundnes, G. (1985) Genetic variation in Atlantic cod (Gadus morhua) throughout its range. Canadian Journal of Fisheries and Aquatic Science 42, 1580–1587. Morris, A.V., Roberts, C.M. and Hawkins, J.P. (2000) The threatened status of groupers (Epinephelinae). Biodiversity and Conservation 9, 919–942. Musick, J.A. (1998) Endangered marine fishes: criteria and identification of North American stocks at risk. Fisheries 23, 28–30. Musick, J.A. et al. (2000) Marine, estuarine and diadromous fish stocks at risk of extinction in North America (exclusive of Pacific salmonids). Fisheries 25, 6–29. Nehlsen, W., Williams, J.E. and Lichatowich, J.A. (1991) Pacific salmon at the crossroads: stocks at risk from California, Oregon, Idaho and Washington. Fisheries 16, 1–21. O’Connell, M., Dillon, M.C., Wright, J.M., Bentzen, P., Merkouris, S. and Seeb, J. (1998) Genetic structuring among Alaskan Pacific herring populations identified using microsatellite variation. Journal of Fish Biology 53, 150–163. Perez, J.E. and Alfonsi, C. (1999) Selection and realized heritability for growth in the scallop, Euvola ziczac (L.). Aquaculture Research 30, 211–214. Pichler, F.B. and Baker, C.S. (2000) Loss of genetic diversity in the endemic Hector’s dolphin due to fisheries-related mortality. Proceedings of the Royal Society, Series B 276, 97–105. Pichler, F.B., Baker, C.S., Dawson, S.M. and Slooten, E. (1998) Geographic isolation of Hector’s dolphin populations described by mitochondrial DNA sequences. Conservation Biology 12, 676–682. Pimm, S.L., Russell, G.J., Gittleman, J.L. and Brooks, T.M. (1995) The future of biodiversity. Science 269, 347–350. Poore, G. and Wilson, G.D.F. (1993) Marine species richness. Nature 361, 587–598. Powles, H., Bradford, M.J., Bradford, R.G., Doubleday, W.G., Innes, S. and Levings, C.D. (2000) Assessing and protecting endangered marine species. ICES Journal of Marine Science 57, 669–676. Reyes, J.C., Mead, J.G. and van Waerebeek, K. (1991) A new species of beaked whale Mesoplodon peruvianus sp. N. (Cetacea: Ziphiidae) from Peru. Marine Mammal Science 7, 1–24.
Ridley, M. (1999) Genome: the Autobiography of a Species in 23 Chapters. Harper Collins, New York. Rijnsdorp, A.D. (1993) Fisheries as a large-scale experiment on life-history evolution: disentangling phenotypic and genetic effects in changes in maturation and reproduction of North Sea plaice, Pleuronectes platessa L. Oecologia 96, 391–401. Rowell, C.A. (1993) The effects of fishing on the timing of maturity in North Sea cod (Gadus morhua L.). In: Stokes, T.K., McGlade, J.M. and Law, R. (eds) The Exploitation of Living Resources. Lecture Notes in Biomathematics, No. 99. Springer-Verlag, Berlin. Ruzzante, D.E., Taggart, C.T. and Cook, D. (1998) A nuclear DNA basis for shelf- and bank-scale population structure in Northwest Atlantic cod (Gadus morhua): Labrador to Georges Bank. Molecular Ecology 7, 1663–1680. Ruzzante, D.E., Taggart, C.T. and Cook, D. (1999) A review of the evidence for genetic structure of cod (Gadus morhua) populations in the Northwest Atlantic and population affinities of larval cod off Newfoundland and the Gulf of St Lawrence. Fisheries Research 43, 79–97. Ryman, N., Utter, F. and Laikre, L. (1995) Protection of intraspecific diversity of exploited fishes. Reviews in Fish Biology and Fisheries 5, 417–446. Shaklee, J.B. and Bentzen, P. (1998) Genetic identification of stocks of marine fish and shellfish. Bulletin of Marine Science 62, 589–621. Shaw, P.W., Pierce, G.J. and Boyle, P.R. (1999) Subtle population structuring within a highly vagile marine invertebrate, the veined squid, Loligo forbesi, demonstrated with microsatellite DNA markers. Molecular Ecology 8, 407–417. Sinclair, M. and Iles, T.D. (1988) Population richness of marine fish species. Aquatic Living Resources 1, 71–83. Slaney, T.L., Hyatt, K.D., Northcote, T.G. and Fielden, R.J. (1996) Status of anadromous salmon and trout in British Colombia and Yukon. Fisheries 21, 20–35. Smith, A.D.M. (1994) Management strategy evaluation – the light on the hill. In: Hancock, D.A. (ed.) Population Dynamics for Fisheries Management. Australian Society for Fish Biology, North Beach, pp. 249–253. Smith, P.J. (1999) Genetic resources and fisheries: policy aspects. In: Pullin, R.S.V., Bartley, D.M. and Kooiman, J. (eds) Towards Policies for Conservation and Sustainable Use of Aquatic Resources. ICLARM Conference Proceedings Series No. 59, pp. 43–62. Smoker, W.W., Gharrett, A.J., Stekoll, M.S. and Taylor, S.G. (2000) Genetic variation of fecundity and egg size in anadromous pink
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salmon, Oncorhynchus gorbuscha Walbaum. Alaska Fisheries Research Bulletin 7, 44–50. Ståhl, G. (1981) Genetic differentiation among natural populations of Atlantic salmon (Salmo salar) in northern Sweden. In: Ryman, N. (ed.) Fish Gene Pools. Ecological Bulletin (Stockholm) 34, pp. 95–105. Stephenson, R.L. (1991) Stock discreteness of Atlantic herring: a review of arguments for and against. In: Wespestad, V., Collie, J. and Collie, E. (eds) Proceedings of the International Herring Symposium, Anchorage, Alaska, 23–25 October 1990 (9th Lowell Wakefield Fisheries Symposium). University of Alaska, Fairbanks, pp. 659–666. Stephenson, R.L. and Kenchington, E. (2000) Conserving Fish Stock Structure is a Critical Aspect of Preserving Biodiversity. ICES CM 2000/Mini:07. Defining the Role of ICES in Supporting Biodiversity Conservation. Stone, R. (2000) Canine virus blamed in Caspian seal deaths. Science 289, 2017–2018. Swaby, S.E. and Potts, G.W. (1990) Rare British marine fishes – identification and conservation. Journal of Fish Biology 37, 133–145. Thorpe, J.E., Gall, G.A.E., Lannan, J. and Nash, C.E. (1995) Conservation of Fish and Shellfish Resources: Managing Diversity. Academic Press.
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Trippel, E.A., Morgan, M.J., Frechet, A., Rollet, C., Sinclair, A., Annand, C., Beanlands, D. and Brown, L. (1997) Changes in age and length at sexual maturity of northwest Atlantic cod, haddock and pollock stocks, 1972–1995. Canadian Technical Reports on Fisheries and Aquatic Science No. 2157. Vecchione, M., Mickevich, M.F., Fauchald, K., Collette, B.B., Williams, A.B., Munroe, T.A. and Young, R.E. (2000) Importance of assessing taxonomic adequacy in determining fishing effects on communities. ICES Journal of Marine Science 57, 677–681. Ward, R.D., Woodwark, M. and Skibinski, D.O.F. (1994) A comparison of genetic diversity levels in marine, freshwater and anadromous fishes. Journal of Fish Biology 44, 213–232. Whitlock, M.C. (2000) Fixation of new alleles and the extinction of small populations: drift load, beneficial alleles, and sexual selection. Evolution 54, 1855–1861. Wing, E.S. (1992) West Indian monk seal. In: Rare and Endangered Biota of Florida, Vol. 1, Mammals. University of Florida Press, Gainesville, Florida, pp. 35–40. Wirth, T. and Bernatchez, L. (2001) Genetic evidence against panmixia in the European eel. Nature 409, 1037–1040.
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The Effects of Fishing on Non-target Species and Ecosystem Structure and Function Henrik Gislason Danish Institute for Fisheries Research, Charlottenlund, Denmark
Abstract Marine fisheries landings increased through most of the 1900s, at the same time as their composition has shifted from larger, fish-eating species towards smaller, plankton-eating fishes. Fishing can affect the composition of the fauna by changing the relative abundance and size distribution of target and by-catch species, by affecting the habitat or by providing discards to scavenging populations such as seabirds. This can lead to changes in species interactions that can affect other parts of the ecosystem. In some cases, fisheries-generated reductions in populations of important forage fish has been reported to affect the growth, abundance and distribution of populations of fish, seabirds and marine mammals that depend on these species for food. Other studies have shown that fisheries-generated habitat changes have had knock-on effects on the local fauna. However, most of the cases where changes in species interactions have been linked to fishing come from relatively simple ecosystems, where a major part of the energy has to pass one or a few species positioned at an intermediate level in the food web. In the more complex systems, the effects of fishing are difficult to separate from natural changes in species abundance due to environmental changes in, for example, temperature and currents, or from man-made changes, such as increases in nutrients. For most of these systems, it is therefore unknown how fishing affects their overall structure and function. Although attempts have been made to develop overall indicators of the impact of fishing on marine food webs, the performance of these indicators has not yet been studied sufficiently to allow them to be used in fisheries management.
Introduction Marine fisheries have expanded considerably during the second half of the 20th century, and on a global scale the annual lendings of marine fish and shellfish now amounts to more than 80 million t (FAO, 2000). In heavily fished areas, such as the North Sea, this means that approximately one-third of the fish biomass is removed each year and, on a global
scale, most commercially targeted fish stocks are now either fully exploited or overfished. The expansion has been accompanied by a shift in the composition of the landings, from large, fish-eating fishes towards smaller, plankton-eating fishes and invertebrates (Pauly et al., 1998). Ecologists use the term ‘trophic level’ to characterize the number of links in the food web from the species in question to the primary producers (level 1) at the
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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bottom of the web. Figure 15.1 shows how the mean trophic level of the global marine landings has changed over the last 45 years. In the 1960s and early 1970s, the large landings from the fishery for Peruvian anchoveta, which feeds at low trophic levels, caused a temporary dip in the average trophic level of the global landings, but from 1972 onwards the mean trophic level has shown a steady decline. Part of this decline may have been caused by developments in markets and technology, but the fact remains that, on a global scale, fishing is now extracting fish further down the food chain than previously. Approximately 75% of the world’s fish catch is taken on the continental shelf or in coastal and estuarine areas where the primary production is high (Pauly and Christensen, 1995). Estimates of the transfer of energy through the food web suggest that between 25 and 35% of the energy fixed by the primary producers in these areas is needed to sustain the current fisheries. When such a high fraction of the production is removed from the system, it is likely that species that compete with the fishery for resources will be affected. It is also likely that fishing will influence production at the lower levels of the food web. It is, however, difficult to quantify these effects. Marine ecosystems are complex and subject to large natural fluctuations caused by changes in temperature and currents for example. The natural changes and the changes
Fig. 15.1.
caused by other human impacts, such as input of nutrients, interfere with the effects of fishing and make them difficult to isolate. Thus, even in situations where major exploited stocks have collapsed, the cause cannot always be attributed exclusively to fishing. In most cases, a combination of high fishing pressure and unfavourable environmental conditions has been involved. To show that a change in the abundance of a target species affects another species requires that both species are monitored, that a causal mechanism (e.g. predation or competition) can be established and, not least, that other likely causes can be excluded. For more than a century, fisheries biologists have struggled to separate the contribution of species interaction, the physical environment and the size of the parent stock, to fluctuations in recruitment, so far with only modest success. It is therefore not surprising that a full answer to the more general question of how fishing affects the structure and function of marine ecosystems cannot yet be provided and that many of the cases where fishing has been found to be involved in ecosystem change are based on circumstantial evidence rather than proof. The rarity of data from unexploited marine ecosystems constitutes an additional problem. In most cases, little information was collected before a fishery developed and, although historical and palaeoecological information can provide insights (e.g.
Mean trophic level of global landings (from Pauly et al., 1998).
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Baumgartner et al., 1992; Jackson, 1997; Quero, 1998), it is often impossible to know how a system looked prior to exploitation. Without knowledge about the unexploited situation, it is difficult to evaluate the present, and fisheries biologists run the risk of using a situation already influenced by fishing as their baseline for evaluating further change (Pauly, 1995). One of the first attempts to quantify how fishing affects the structure and function of marine ecosystems dates back to the Italian scientist, Vito Volterra, who tried to explain the apparent increase in sharks and other predators and the decline in their fish prey in the Adriatic Sea during the First World War (Gasca, 1996). His son-in-law, Umberto D’Ancona, had noticed a change in the landings in Venice, Trieste and Fiume after the war. The species composition had changed, even though the total landings had remained the same. According to D’Ancona, the cessation of the fishery during the war changed the species composition in favour of fish-eating fishes, and Volterra developed a simple theoretical model where the observations were explained as a result of changes in species interactions. Since then, a large number of case studies have examined how fishing affects the structure and functioning of marine ecosystems. In this chapter, I will review a number of these case studies, describe a couple of the indices that have been proposed to capture the overall impact of fishing, and end by briefly summarizing the general patterns that emerge.
The Impact of Fishing on Marine Ecosystems Apart from reducing the biomass of target and by-catch species, fishing can affect the structure and function of marine ecosystems through at least four mechanisms. First, changing the relative abundance of target and non-target species can lead to additional changes in other species mediated by alterations in predator–prey relationships or competition. Second, fishing is likely to affect populations of large slow growing and late maturing species to a larger extent than
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populations of small rapidly growing and early maturing species. This can lead to shifts in the relative abundance of species with different life history characteristics, independent of any changes in species interactions. Third, discarding of fish and offal can increase populations of scavenging species. Finally, gear in contact with the seabed can modify by removing stones and boulders or by damaging corals and other fauna or flora that provide a structural habitat, and this may influence the composition of the fauna. All of the four processes probably work in combination, and all may result in shifts in the relative abundance and growth of the various species in the community. However, changes in abundance and growth can also occur as a result of natural fluctuations in the physical environment or be generated by other human activities. Marine ecosystems are complex and variable. There are differences between marine ecosystems in the tropics, the temperate and the boreal parts of the world as well as between systems in the coastal zone and the deep sea. Many of the systems are subject to strong environmental forcing. As the following case studies demonstrate, it is often difficult to separate the changes that are caused by fisihing from those occurring naturally.
Coastal and estuarine systems Due to the growing interest in establishing marine protected areas, a large number of studies have been undertaken recently to demonstrate the effects of coastal marine reserves. Mosquera et al. (2000) reviewed empirical studies of the effect of marine reserves on the fish fauna and found that fish abundance was on average more than three times higher within the reserves than outside. Most of the difference was, however, due to an increase in large-bodied target species, while non-target species generally were equally abundant inside and outside the reserves. This suggests that any indirect effects of a change in the abundance of the target species had only a minor or no impact on the abundance of the non-target species.
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Similar observations have been made in coral reefs, where the indirect effects of removing fish predators on the abundance of their fish prey have been subtle (Jennings and Kaiser, 1998). Habitat changes can, however, be important. Coastal fishermen use various gears that can disturb the habitat either directly or indirectly, and effects have been noticeable where fishing has affected species providing a structural habitat such as corals, sea-grasses or seaweed. In coral areas, where destructive fishing methods, such as explosives, cyanide or heavy trawling gear, have destroyed the corals, the associated fauna has disappeared. In less impacted areas, the changes are often not as dramatic. On reefs where the abundance of fish-eating fish species has been reduced, herbivorous fishes will range farther from their coral shelter and/or become more abundant (McManus et al., 1992). This has been suggested to increase the areas on adjacent reef flats devoid of sea-grass (Jennings and Polunin, 1997; Miller and Hay, 1998). However, when fishing affects ‘keystone’ species either directly or indirectly, the effects on the reefs can be pronounced. Widespread changes were thus observed in the coral reefs of Jamaica when the algal-grazing sea urchin Diadema antillarum suffered mass mortalities from a species-specific disease from 1982 to 1984. This resulted in a massive bloom of fleshy macro-algae that overgrew the corals and produced a dramatic decline in coral cover (Hughes, 1994). The change in coral cover was brought about by a sequence of events that included a hurricane that struck the island in 1980, and overfishing of the sea urchin’s natural fish predators and the algal grazing fishes with which they competed. This left Diadema in control of algal growth. By removing the competitors and predators of the sea urchins, fishing set the stage for the shift to algae that followed the sea urchin collapse. Similar effects have been observed elsewhere. On the Great Barrier Reef, the removal of a number of fish species that feed on young starfish has been implicated in outbreaks of a coral-eating starfish (Acanthaster) (Bradbury and Seymour, 1997), and in Kenya the removal of triggerfish by fishing has caused an increase in
the abundance of a burrowing sea urchin (Echinometra), leading to extensive bio-erosion of the reefs (McClanahan and Muthiga, 1988; McClanahan et al., 1996). In colder waters, productive kelp forests support a highly diverse fauna. Kelp forests are variable, and influenced by grazing, storms and El Niño events. Among the animals that feed on kelp, sea urchins are by far the most important (Dayton, 1985). When sea urchins are abundant, they are able to graze down the kelp and create ‘barren grounds’, resulting in a dramatic change in the fauna. The abundance of sea urchins is influenced by recruitment, predation, disease and immigration of adults: factors that vary greatly over time and space. The abundance and types of sea urchin predators also vary from area to area. In the North Pacific, along the coasts of Alaska and Canada, sea otters function as keystone predators. When sea otters are abundant, they reduce the abundance of sea urchins. In the absence of sea otters, dense populations of sea urchins develop and graze down the kelp (Estes and Duggins, 1995). Along the coast of California, the situation is more complex, but two sea urchin predators, spiny lobster and the sheephead, a labrid fish, are likely to play a role similar to sea otters further north. Both of these predators are subject to intense fishing, and it has been suggested that fishing may lead to more abundant grazing outbreaks (Tegner and Dayton, 2000). A similar interaction between fishing and sea urchin grazing may exist in other areas where commercially exploited predators affect sea urchin abundance. In the northwestern Atlantic, a range of large, commercially exploited fish prey on sea urchins (Steneck, 1997); off South Africa, rock lobster is the most important sea urchin predator (Anderson et al., 1997); and, in northern New Zealand, spiny lobsters and snappers prey on sea-urchins (Babcock et al., 1999).
Semi-enclosed seas Many semi-enclosed seas, such as the Baltic, the Black Sea and the Mediterranean, support important fisheries. The fauna in these areas
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have undergone marked changes over time. However, fishing is unlikely to be the only driving force. In many areas, nutrient enrichment also seems to be involved (Caddy, 1993). The most dramatic example of the combined effects of nutrient enrichment, overfishing, decreases in freshwater input and the introduction of alien species has been observed in the Black Sea (Shiganova and Bulgakova, 2000). In the 1970s, increases in nutrient input produced an increase in primary and secondary production at the same time as fishing reduced the biomass and changed the species composition of the fish fauna. The larger, fish-eating fishes declined, while small pelagics increased. In the early 1980s, the increasing production at lower trophic levels led to an explosive development of jellyfish and to a rapid growth of the population of an invading ctenophore Mneniopsis. During the 1980s and in the early 1990s, jellyfish and Mneniopsis reduced the biomass of zooplankton to a level that led to a reduction in the recruitment of the fish populations, in particular anchovy, after which the fish populations and fisheries collapsed. In the late 1990s, the situation
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started to reverse after the invasion of another ctenophore, Beroe. This species feeds on Mnemiopsis and apparently has reduced its biomass in the most recent years. Despite general increases in nutrient input, quite as dramatic changes have not been observed in other semi-enclosed seas. In the Baltic Sea, the large reduction in cod biomass following intensive exploitation in the late 1970s and 1980s was accompanied by an increase in the biomass of sprat. The fish fauna in the Baltic is dominated by three commercially exploited species: cod, herring and sprat. The interactions between these species are well studied (Sparholt, 1994). Cod is the major predator of sprat and herring, and the reduction in cod biomass in the 1980s coincided with an increase in the biomass of sprat, and was followed by a subsequent increase in the growth rate of cod (Fig. 15.2). Since sprat also feeds on the eggs and larvae of cod, it has been suggested that the fishery has driven the system from a cod-dominated state with low biomasses of sprat, to a sprat-dominated state in which the recruitment to the cod stock is limited by sprat predation (Rudstam et al., 1994). However, the story is complicated by
Fig. 15.2. Changes in spawning stock biomass of cod and sprat in the Baltic together with cod weight at age 4 (modified from Gislason, 1999).
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the fact that the changes have been observed in a period where nutrient concentrations increased concurrently with a decline in the salinity and oxygen content of the water in spawning areas of cod over the deep parts of the Baltic, partly as a result of a reduced inflow of saline waters from the Atlantic. Thus, both changes in environmental conditions and a high fishing pressure reducing the spawning biomass of adult cod are likely to be implicated.
Boreal shelf areas The Barents Sea north of Norway and Russia is subject to large variations in climatic conditions, with changes between cold and warm periods (Loeng, 1989). Capelin, juvenile herring and polar cod are the most important plankton-eating fish, while cod and a variety of seabirds and marine mammals constitute the main predators (Hamre, 1994). After hatching, juvenile herring from the AtlantoScandian herring stock drift along the west coast of Norway and eventually enter the Barents Sea, where they stay until they mature. When the Atlanto-Scandian herring
stock collapsed in the late 1960s due to heavy fishing and deteriorating conditions, the inflow of juvenile herring to the Barents Sea ceased. In the 1970s, the cod population, a predator on both herring and capelin, continued a long-term decline caused by fishing, but the capelin biomass was high despite heavy fishing. Herring is a predator of larval and juvenile capelin, and when recruitment to the Atlanto-Scandian herring stock improved and the influx of juvenile herring increased again, the predation on capelin increased. The increased predation, low individual growth rates and heavy fishing made the capelin stock collapse in the mid-1980s (Gjøsæter, 1995, 1998). Following the capelin collapse, the growth and fecundity of cod decreased and cod cannibalism increased (Bogstad and Mehl, 1997) (Fig. 15.3). Seabirds and marine mammals feeding on capelin were also strongly affected. Harp seals migrated down the coast of Norway in 1987 and high unintentional catches occurred in fishing nets along the coast, while guillemots suffered winter kills of thousands of individuals in the Barents Sea (Livingston and Tjelmeland, 2000). The capelin stock recovered as a result of the large 1989 yearclass, only to collapse again when the large 1991
Fig. 15.3. Changes in average biomass of cod and biomass of capelin in the Barents Sea together with cod weight at age 6 (data from Bogstad and Mehl, 1997).
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and 1992 yearclasses of herring entered the area. This time, however, the guillemot population did not decline, perhaps because the birds were able to switch to herring as an alternative food (Anker-Nilssen et al., 1997). The changes in the relatively simple ecosystem of the Barents Sea demonstrate the importance of the links between capelin, cod and herring. The overfishing of herring in the adjacent Norwegian Sea and changes in the inflow of warm water from the Atlantic (Skjoldal et al., 1992) triggered a sequence of changes in which fishing seems to have played a significant role. In the Bering Sea between Alaska and Russia, recent declines in populations of marine mammals and seabirds have attracted attention. The various hypotheses explaining these declines are discussed in a report from the USA National Research Council (NRC, 1999). As in the Barents Sea, the environmental conditions are variable, with interdecadal changes in sea surface temperature and ice cover. The biomass and composition of the demersal fish community has undergone large changes over time (Livingston and Tjelmeland, 2000). Over the last 20 years, the stock of walleye pollock has increased, and in the beginning of the 1990s it accounted for half the biomass of demersal fish. The causes for the increase are not well known. It may be related to a number of warm years, where pollock cannibalism was reduced because the adults and juveniles were separated spatially (Wespestad et al., 2000), but it may also be linked to increases in the food supply to pollock caused by a reduction in the abundance of other fishes and whales competing for the same resources as pollock. In the 1950s, 1960s and early 1970s, whales were exploited intensively. At the same time, trawl fisheries severely reduced the stocks of eastern Bering Sea shelf flatfishes and herring as well as the stocks of rockfishes found on the upper continental slope. In addition, ocean climate changed as sea surface temperatures increased and ice cover reduced. The increase in pollock biomass was accompanied by an apparent decline in a number of smaller fish species that serve as food for pollock. Seabirds, seals and other top predators declined as well. In the Pribiloff Islands, juvenile and
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adult female fur seals declined from the mid-1970s to the early 1980s. Harbour seals declined rapidly in the late 1970s in the Gulf of Alaska and have remained low since then. Steller sea lions declined in the eastern Aleutians in the early 1970s and in the central and western Aleutians and western Gulf of Alaska in the 1980s. The evidence available to explain these declines is mainly circumstantial, but for Steller sea lions they have been explained as a result of a fisheries-generated lack of food. Steller sea lions eat pollock in addition to other forage fish, and as the pollock fishery overlapped with the foraging areas of the sea lions it may have reduced the biomass of pollock in these areas sufficiently to reduce the growth of the mammals. An alternative hypothesis is that Steller sea lions have declined as a result of a reduction in the availability of highenergy prey fish such as herring and sand eel (Rosen and Trites, 2000). The decline in sea lions and seals may have had knock-on effects in nearshore kelp forests, where declines in sea otter populations have been observed recently (Estes et al., 1998). These declines have been suggested to be a result of increases in killer whale predation on sea otters, and hypothesized to be linked to changes in the foraging area of the killer whales caused by the reduction in the abundance of the sea lions and seals that they normally would eat.
Temperate shelf areas The North Sea is one of the most intensively studied marine systems in the world, and several analyses of long-term trends in the species composition of fish and other biota are available. For fish, the analysis of longterm survey time series reveal that the composition of the demersal fish fauna of the northern North Sea has remained surprisingly stable. Greenstreet et al. (1999) studied the species composition in bottom trawl survey data from 1925 to 1996. Comparing different periods, they found only small changes in the relative species composition. The most apparent was an increase in Norway pout, a small gadoid, for which an industrial fishery developed in the 1960s. Examination of the
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size composition of the catches revealed a shift towards smaller fish in the whole assemblage, but this shift was not significant for the non-target species. The fish fauna in the southern North Sea has shown a larger change. Rijnsdorp et al. (1996) compared catches from bottom trawl surveys in 1906–1909 with data from 1990–1995, and found a reduction in overall abundance as well as a decrease in diversity caused by an increase in dominance of a few of the target species. Some of the exploited species have also exhibited changes in growth and size at maturity. During the Second World War, the North Sea fishery declined. Studies of North Sea plaice show that growth was reduced during the war and in the period immediately after, coinciding with a three-fold increase in the plaice stock. Length at first maturity has decreased in both male and female plaice since 1900 (Rijnsdorp and Leeuwen, 1992). While part of the decrease in length at first maturity can be explained by phenotypic plasticity and differences in water temperature, the authors suggest that the remaining part could be linked to genetic selection of a smaller size at first maturity caused by size-selective fishing. The results referred to above considered the long-term changes. Over shorter time spans, considerable changes in the fish fauna have been observed. One of the most apparent was the gadoid outburst that started in the beginning of the 1960s (Cushing, 1980). Over the next 20 years, the main commercially exploited gadoid species – cod, whiting, haddock and Norway pout – all produced one or more outstanding yearclasses. The outburst has been explained by fisheries-generated changes in species interactions (Andersen and Ursin, 1977), while others have correlated gadoid recruitment to various environmental parameters such as salinity anomalies (Dickson et al., 1988a), the prevalence of westerly (Aebisher et al., 1990) or northerly winds (Dickson et al., 1988b), and wind stress and winter cooling (Svendsen and Magnusson, 1991). At the same time as gadoid recruitment increased, the mackerel and herring stocks declined due to heavy fishing. In the second half of the 1960s, the stock of North Sea mackerel dropped from almost 3 million t to 0.5
million t, while herring declined from 2 to 1 million t. Andersen and Ursin (1977) built an extensive model of the North Sea ecosystem in an attempt to explain the gadoid outburst. The model included primary production, two functional groups of zooplankton, three benthos groups, detritus and 11 interacting fish species. The model was able to mimic the general increase in the gadoid biomass as a result of reductions in mackerel and herring predation on juvenile gadoids, but did not allow the authors to exclude the possibility of a general upward recruitment trend due to other environmental factors. In a later analysis of the timing of the events, Hislop (1996) concluded that none of the explanations offered have provided a satisfactory explanation. The first large yearclass of haddock occurred in 1962 at a time when the stocks of both herring and mackerel were high. According to Hislop (1996), we are still – almost 40 years after it began – no closer to understanding the causes of the outburst. Currently, some 1 million t of fish, offal and benthic invertebrates is discarded annually by commercial fisheries in the North Sea (Tasker et al., 2000). A number of seabird species feed on these discards, and many of these have shown large increases over most of the 20th century. The increases have been linked to improved food resources, in particular discards, but are also related to the decrease in seabird hunting that took place in the early part of the century. The birds are unable to eat all of the discarded material. A part will sink outside their feeding range or be composed of flatfish and other items that are difficult to swallow. However, a link between increases in discards and overall changes in the abundance of scavenging species of fish and invertebrates on the seabed has not yet been established. When the mackerel and herring stocks declined in the North Sea, a fishery for sandeels developed. Mackerel and herring include sandeel in their diet, and an increase in sandeel biomass was predicted by the model of Andersen and Ursin (1977). However, no data on sandeel abundance from the 1960s and early 1970s are available to show whether such an increase did in fact take place. On Georges Bank, on the other side of
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the Atlantic, a similar decrease in mackerel and herring was accompanied by an increase in the abundance of sandeel larvae (Sherman et al., 1981). Fogarty et al. (1991) studied the interaction between sandeel, herring and mackerel, and found a significant inverse relationship between an index of the abundance of the latter two species and sandeel recruitment over the time period 1970–1986. The changes in mackerel, herring and sandeel stocks were not the only things that happened on Georges Bank. During the 1960s, the total fish biomass was more than halved as a result of increasing exploitation by distant water fleets (Fogarty and Murawski, 1998). After the establishment of the 200-mile limit in 1977, exploitation initially declined, but then increases in the domestic fleet led to a new increase in the exploitation of gadoids and flatfish, resulting in a reduction of their biomass to historically low levels. The reduction in the biomass of demersal target species was accompanied by an increase in small sharks and rays and, due to reductions in pelagic fishing effort, to an increase in the abundance of herring and mackerel in the late 1980s. Recently, fishing has reduced the abundance of the smaller sharks and rays again. The increase in sharks and rays on Georges Bank in the 1970s and 1980s has been attributed to reduced competition caused by the decrease in the biomass of gadoids and flatfish with which they compete for food (Fogarty and Murawski, 1998). The reduced populations of demersal fish have recently exhibited high individual growth rates (Hunt, 1996) consistent with the hypothesis that competition for food is important in the system. The reductions in herring and mackerel biomass in the late 1960s and 1970s and the increase in sandeel abundance has been used to explain changes in the distribution of whales on the Georges Bank and in the Gulf of Maine (Kenney et al., 1996).
Tropical shelves In the 1960s, intensive fisheries developed in the Gulf of Thailand (Pauly, 1988; Christensen, 1998). Overall catch rates
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declined from about 300 kg h−1 in 1961 to 54 kg h−1 in 1983 (Suvapepun, 1991). Declines were particularly dramatic for long-lived fish (e.g. rays and sawfish) as well as for several families of smaller fish, while squids increased in abundance. Shrimps initially declined but then increased somewhat again. Pauly (1985) found an inverse relationship between the mortality of small squid and the biomass of fish, and suggested that the increase in squid was caused by the removal of their predators. Harris and Poiner (1991) studied a similar system in the southeastern Gulf of Carpentaria, Australia, but found comparatively little change over time. Comparing the composition of the fish fauna before the start of the prawn fishery with surveys made 20 years later, the majority of the taxa examined showed no significant change in abundance. The modest levels of fishing compared with the Gulf of Thailand (Blaber et al., 1990) may explain the different response. On the Saharan Bank, in West Africa, catches of cephalopods were low until the 1960s, when landings increased rapidly. Before the 1960s, the landings were composed mainly of demersal fish, and when the species composition changed towards cephalopods this was interpreted as a change in the ecosystem mediated by an overexploitation of sea breams (Sparidae) (e.g. Caddy and Rodhouse, 1998). However, a recent analysis of survey data shows that the changes in the relative species composition of the fauna were much less dramatic than the landings data would suggest (Balguerias et al., 2000). Comparing survey data from 1942 with recent data, the most apparent change was a major decrease in the total catch rate in the surveys, by a factor of seven. This decrease was most severe for long-lived species of sharks and rays, but finfish and commercially important crustaceans such as lobsters and large crabs also declined in relative abundance, at the same time as other crustaceans and benthic cephalopods increased. The authors attribute the increase in benthic cephalopods to reductions in predator populations and increases in the food available to scavengers due to increased discarding.
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Sainsbury (1991) and Sainsbury et al. (1997) studied the changes in the fish fauna on the northwest shelf of Australia. A trawl fishery for demersal fish developed in the beginning of the 1970s, targeting mainly breams, lizardfish, emperors and snappers. Later, a trap fishery for the latter two species evolved. Research surveys from 1960 onwards showed changes in species composition but little change in overall fish biomass. The abundance of emperors and snappers declined, while the biomass of lizardfish and breams increased. At the same time, the quantity of large epibenthos (mainly soft corals and sponges) declined, possibly due to fishing. To resolve the question of whether the changes in the fish fauna were due to changes in species interactions, or to the removal of large epibenthos, or whether they could be explained by differences in the life history parameters of the four species, an experiment was designed in which a part of the area was closed to trawling. Later examination of data collected from the area suggested that the removal of soft corals and sponges by the fishery was the most likely explanation for the observed changes in species composition. Thus, in this case, the changes in the fish fauna seem to have been caused by habitat modification rather than by changes in predatory or competitive interactions between the species.
Upwelling systems Upwelling systems are subject to large, environmentally driven fluctuations. One of the most noticeable changes in such a system took place off Peru in 1972. The area along the Peruvian coast is highly productive due to wind-driven upwelling of nutrient-rich water, and the production of pelagic fish in the area varies in response to the upwelling intensity. In years with weak winds and little upwelling, smaller catches are taken, and vice versa. The changes in the strength of the wind are influenced by the so-called El Niño–Southern Ocean oscillation (ENSO), driven by changes in atmospheric pressure systems. In 1972 and 1973, the upwelling intensity was particularly low, and this had a major impact on the fishery of anchovies. In 1970, the fishery landed 13 million t of anchoveta, but the landings fell to 2–3 million t in subsequent years, and to 1 million t in 1980. Following the decline of the anchoveta landings, sardine landings increased, but in 1995 anchovies recovered again. The reduction of the anchovy population had a large impact on species of seabirds such as the guanay cormorant, whose population declined due to starvation and reduced breeding. Figure 15.4 shows how the increase in the landings of anchoveta in the 1960s was
Fig. 15.4. Landings of anchoveta and sardines in Peru and Chile and population size of Guanay cormorants (data from FAO landings statistics and Crawford and Jahncke, 1999).
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accompanied by a reduction in the number of breeding guanay cormorants. Similar changes in seabird mortality and breeding success have been observed in other upwelling systems (Crawford and Dyer, 1995; Crawford and Jahncke, 1999). Shifts between anchovy- and sardinedominated periods have been observed in many upwelling systems throughout the world (Schwartzlose et al., 1999). The shifts seem to be triggered by the formation of one or a couple of extraordinary yearclasses. This may happen while the other species is still abundant, and therefore does not necessarily represent a response to a reduction in the stock size of a competing species. Rather, a combination of environmental changes and fishing often seems to be involved. Excessive fishing may reduce the abundance of the dominating species to the point where a couple of low yearclasses will lead to a collapse of the stock. Fishing may also concentrate on abundant yearclasses and thus prevent the least abundant fish from becoming dominant. Records of scale deposits in anoxic sediments thus show that large population fluctuations have occurred in the Californian upwelling system over several millennia and long before any fishing started (Baumgartner et al., 1992). Furthermore, many of the recent fluctuations exhibit a global synchrony, suggesting that climate may be the major driving force (Klyashtorin, 1997). However, in a review of the impact of ten major fisheries for small pelagic species, Beverton (1990) concluded that fishing was the main cause of the collapses observed in most, but not all cases, and the relative role of climate and fishing in the collapse of exploited stock of small pelagic species remains controversial (Cury et al., 2000).
Deep-sea ecosystems Most of the fisheries exploiting continental slope and deep-sea resources have a relatively short history, and the indirect effects that these fisheries may cause are largely unknown. Deep-sea fish species are characterized by slow growth, extreme longevity, low fecundity and high age at first maturity,
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and any indirect effects of their removal are therefore likely to require considerable time before they may result in measurable change (Koslow et al., 2000). Deep-sea fishing has, however, been shown to have a potentially severe impact on deep-sea corals and other benthos on the seamounts where many of the commercially exploited fish species congregate. The ability of the benthos to regenerate has not been investigated, but is most likely to be low. No data are available to demonstrate how the altered habitat will affect the remaining fauna.
Indices of fisheries-generated changes in ecosystem structure and functioning The general wish to develop precautionary management systems that take the ecosystem effects of fishing into account has generated a search for useful reference points for ecologically dependent species, as well as for overall trophic level balance (Christensen, 2000; Gislason et al., 2000). In some instances, reference points for forage species have been developed that take the food requirements of their natural predators into account (e.g. Constable et al., 2000; ICES, 2001). The development of such reference points requires that strong trophic linkages can be identified and that biologically significant indicators can be monitored, such as the condition, diet composition, breeding success or abundance of the predator. Indicators that take overall trophic level balance into account have not yet been fully developed. A number of indices have been proposed, but their ability to reflect changes in trophic level balance awaits further investigation and reference points have not yet been identified. In the following, I will give a brief description of two of the proposed indicators, the slope of size spectra and the so-called Fishing-in-Balance (FIB) index. Size is an important determinant of trophic interactions in marine systems. Cod larvae will be feeding on plankton prior to settlement, juvenile cod on crustacea and other small invertebrates, while adult cod will eat larger invertebrates and increasingly more
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fish as they grow. It is therefore not unreasonable to assume that size by and large determines an individual’s trophic position in the food web. If this is so, then the shape of the overall size distribution should reflect how energy is transferred between trophic levels (e.g. Borgman, 1987; Thiebaux and Dickie, 1993). Pope and Knights (1982) compared the size composition of demersal fish caught by bottom trawl surveys in the North Sea and at the Faeroe Islands, and found that a straight line fitted log numbers per size class versus size in both cases. Subsequent comparisons of size spectra from demersal fish communities from various parts of the world has confirmed that the log of the numbers per size group often is linearly related to the size of the fish and have suggested that the slope of the spectrum is related to fishing intensity (Pope et al., 1987; Murawski and Idoine, 1992; Gobert, 1994; Bianchi et al., 2000). Figure 15.5 shows how the slope of the size spectrum in the North Sea has changed over time. There are now relatively fewer large fish in the system than there were at the beginning of the time series. Similar reductions in the slope have been observed in a number of other systems (Bianchi et al., 2000). Gislason and Rice (1998) modelled the North Sea size spectrum and found that the slope was inversely proportional to overall
fishing mortality. However, the relationship was found to be sensitive to changes in species growth. Almost identical changes in slope were found for spectra simulated by singleand multi-species models, suggesting that changes in predator–prey relationships may be relatively unimportant compared with the other life history parameters of the included species. The finding that the slope of the size spectrum is inversely proportional to fishing would make it suitable as an overall indicator of fishing impacts. However, as concluded by Bianchi et al. (2000), the usefulness of the size spectrum slope for management purposes currently is limited, as there is insufficient empirical and theoretical background for the changes observed. It is also unclear how suitable reference points could be defined. The FIB index represents another attempt to capture the effect of fishing on the energy transfer in the food web (Pauly et al., 2000). The FIB index attempts to capture the effect of predator removals. Biological production increases by roughly a factor of ten moving one trophic level down in the food web. If fishing reduces the mean trophic level by one, we would assume that this would generate a tenfold increase in landings, provided the fishery was able to harvest the surplus production. The FIB index of a given year, y, is calculated relative to the FIB index in the first
Fig. 15.5. Changes in the slope of the size spectrum of North Sea fish as caught during the International Bottom Trawl Survey (ICES, 1996).
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year, 1, of a time series of catch data from the equation: FIB = log
[(Catch y × TE TL −1 ) /] y
[(Catch1 × TE TL −1 )] l
where TL is the average trophic level and TE is the mean energy transfer efficiency between trophic levels (Christensen, 2000). The FIB index will stay constant if a decrease in the average trophic level is matched by a sufficiently large increase in the catch. Christensen (2000) applied the FIB index to the Gulf of Thailand and the North Atlantic. For the latter area, the index increased as the fishery expanded over the period from 1950 to
Fig. 15.6. Changes in the FIB index of the Gulf of Thailand landings (from Christensen, 2000).
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1976, and then decreased from 1976 onwards. For the Gulf of Thailand, the index initially rose, but then remained at a constant level for the remaining period (Fig. 15.6). This would indicate that the fishery in the Gulf of Thailand is in balance as the catches increased in the expected way even though the average trophic level of the catch declined from approximately 3.4 to 3.2. In the North Sea, the FIB index shows an overall increase up to 1970, before it levels off (Fig. 15.7), while the global FIB index increases up to 1987 before it stabilizes (Fig. 15.8). It is still too early to judge the utility of the FIB index as a measure of trophic balance. The index tries to capture the extent to which fishing disrupts the flow of energy from the lower trophic levels. The index can be quantified easily from landing statistics, and trophic levels from published ECOPATH models but, just as for the size spectrum slope, it is not known exactly when a reduction in the index is a cause for alarm. Assume, for instance, that the fishing effort in the North Sea is halved, resulting in a 50% reduction in the catch of all species the following year. In this case, the average trophic level of the catch remains constant, but since the catches are halved the FIB index will decline in the short term and
Fig. 15.7. Changes in the FIB index of the North Sea landings (data kindly supplied by Villy Christensen).
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Fig. 15.8. Changes in the FIB index of global marine landings (data kindly supplied by Villy Christensen).
not increase until either the catch or the average trophic level increase again.
A brief summary of the changes observed It is well documented that fishing can lead to shifts in community structure and composition by removing fish with different life histories. Fishing imposes mortality on target and by-catch species, and life history theory predicts that populations of large, slow-growing, late-maturing species should decline more in response to fishing than small, earlymaturing species (Kirkwood et al., 1994). Comparing the relative decline of large and small species within closely related groups, Jennings et al. (1999a,b) thus found a significantly greater decline in the abundance of the larger species, both in reef fishes in Fiji and in demersal fishes in the northern North Sea. Many species of sharks and rays are large, grow slowly and mature at a high age, and the virtual extirpation of some of these species from large regions (Stevens et al., 2000) provides a good example of the importance of life history characteristics for the ability to withstand fisheries-generated mortality. Reductions in populations of forage fish have been reported to affect the growth, reproductive rate, and population size of their natural predators. For instance, the condition
and growth of cod in the Barents Sea and around Iceland changed in response to changes in the abundance of capelin (Mehl and Sunnanå, 1991; Stefánsson et al., 1998; Yaragina and Marshall, 2000), and cod growth in the Baltic seems to be related to the abundance of sprat (Gislason, 1999). There have also been well documented cases where natural or fisheries-induced reductions in forage fish abundance have led to collapses of seabird populations. The most prominent examples come from upwelling areas, where changes in pelagic fish stocks can have a large impact on the seabird populations feeding upon them (Crawford and Dyer, 1995; Crawford and Jahncke, 1999), there are also examples from other areas, such as the starvation of seabirds in colonies along the Norwegian coast following the fisheries-induced collapse of the Atlanto-Scandian herring stock (Anker-Nilssen et al., 1997). For mammals, the reductions in the growth of Steller sea lions (NRC, 1996) and the lack of food that resulted in a mass migration of harp seals along the coast of Norway in 1987 (Livingston and Tjelmeland, 2000) may be linked to fishing. Predator removal has been suggested as leading to increases in prey populations. In a range of ecosystems, predation has been estimated to remove between 2 and 35 times more of the total fish production than fishing (Bax, 1998). It would therefore be surprising if prey populations did not respond when fishing
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reduced the abundance of their predators. However, prey releases following predator removal do not appear to be quite as common as intuition would suggest. In the coastal zone, sea urchin populations on coral reefs and in kelp beds have responded to changes in predator abundance, but reductions in the abundance of fish-eating fish species on coral reefs have not lead to increases in the abundance of their fish prey (Roberts, 1995; Jennings and Kaiser, 1998). Comparisons of the fish fauna inside and outside areal closures in coastal tropical and subtropical areas also suggest that indirect effects are of minor importance (Mosquera et al., 2000). Systems found in deeper waters are not as easily manipulated, and the lack of controlled experiments makes most of the evidence from these areas circumstantial. The increases in populations of forage fish such as the sprat population in the Baltic (Rudstam et al., 1994), sandeel on Georges Bank (Fogarty et al., 1991) and capelin in the North West Atlantic (Carscadden et al., 2001) have been attributed to fisheries-generated reductions in predator populations. However, environmental changes may also be involved. Competition has been suggested less often as a likely explanation for changes in species abundance. There have been dramatic shifts in species composition in the fisheries for small pelagics, where one species apparently replaced another, but a closer look at the available time series data shows that the subordinate species often started to increase before the dominant one began to decrease. In many cases, the shifts are correlated over large distances, suggesting that change in global climate is involved (Schwartzlose et al., 1999). Perhaps the best example of competition comes from Georges Bank, where fishing greatly reduced the biomass of demersal fish and this was followed by an increase in small sharks and rays. However, as Fogarty and Murawski (1998) comment, it is unknown whether resources such as food and space were in fact limiting for the sharks and rays prior to the change. In their review of the ecosystem effects of fishing, Jennings and Kaiser (1998) concluded that little evidence exists to suggest that fishing has resulted in compensatory replacements of one species by another.
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The experimental studies by Sainsbury (1991) and Sainsbury et al. (1997) of areal closures on the North West Shelf of Australia demonstrate that habitat modification can be important. The removal of soft corals and sponges by the fishery was the most likely explanation for the changes observed. Similar results emerge from studies of fish assemblages on coral reefs. In deeper waters, it generally has been difficult to link changes in fish assemblage structure to fisheries-generated habitat changes. However, a recent study argued that there were links between low survival of juvenile cod to reductions in habitat complexity caused by mobile fishing gear (Lindholm et al., 2001). A large proportion of the discards from fishing vessels is eaten by seabirds. For some of the species involved, this seems to have resulted in population increases, with subsequent changes in the species composition of the seabird community (Tasker et al., 2000). The fate of the fraction of the discards that sinks through the water column is uncertain, but studies suggest that mid-water scavengers, such as cetaceans and sharks, may benefit. The proportion that ends on the seabed adds to the benthos, together with fish that have been damaged or killed by the gear but escaped before it was brought on deck (Jennings and Kaiser, 1998). This source of food is eaten by a number of scavenging invertebrates and fish. While the effects of discarding on the seabird population are reasonably well understood, the long-term effects on mid-water and demersal scavengers are unknown, as are the consequences of the changes in the interactions with their natural prey. Most of the evidence for fisheries impacts on ecosystem structure comes from systems where biota important for the structure of the habitat have been affected either directly or indirectly (e.g. coral reefs, soft corals and sponges) or where strong trophic linkages occur, such as in the so-called ‘wasp-waist’ systems, where the major part of the energy has to pass through one or a few species positioned at an intermediate level in the food web (e.g. capelin and herring in the Barents Sea; sardines and anchovies in upwelling systems). This is not to say that food web effects
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can be safely neglected elsewhere. The availability of data and scientific effort differs from one system to another, and this could bias the conclusion. Coastal ecosystems are much more accessible for direct observation and experimentation than systems in deeper water. Species richness increases as one moves from boreal and temperate systems towards the tropics, at the same time as the availability of data and scientific effort decreases. The common use of regressions to relate ecosystem change to fishing and the existence of diffuse effects in complex systems (Yodzis, 2000) probably means that the likelihood of identifying a relationship decreases as the complexity of the systems increases, at the same time as the number of possible explanatory variables grow. However, the relative scarcity of reported effects from the more complex systems may also reflect a real difference in stability. Theoretical studies have shown that increasing diversity will increase food web stability provided the interactions between the species are weak (McCann, 2000), and this could very well be one of the underlying overall themes.
Acknowledgements Thanks to Villy Christensen for providing figures on changes in the FIB index for the Gulf of Thailand, the North Sea and for Global Marine Fisheries landings, as well as to Tom Fenchel, Louize Hill, Morten Vinther and Jake Rice for providing useful comments and suggestions on draft version of this chapter.
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NRC (1999) Sustaining Marine Fisheries. National Academy Press, Washington, DC. Pauly, D. (1985) The population dynamics of short-lived species, with emphasis on squids. NAFO Scientific Council Studies 9, 143–154. Pauly, D. (1988) Fisheries research and the demersal fisheries of Southeast Asia. In: Gulland, J.A. (ed.) Fish Population Dynamics. John Wiley and Sons, London, pp. 329–348. Pauly, D. (1995) Anecdotes and the shifting baseline of fisheries. Trends in Ecology and Evolution 10(10), 430. Pauly, D. and Christensen, V. (1995) Primary production required to sustain global fisheries. Nature 374, 225–257. Pauly, D., Christensen, V., Dalsgaard, J., Froese, R. and Torres F., Jr. (1998) Fishing down marine food webs. Science 279, 860–863. Pauly, D., Christensen, V. and Walters, C. (2000) ECOPATH, ECOSIM, and ECOSPACE as tools for evaluating ecosystem impact of fishing. ICES Journal of Marine Science 57(3), 697–706. Pope, J.G. and Knights, B.J. (1982) Comparison of the length distributions of combined catches of all demersal fishes in surveys in the North Sea and at Faeroe Bank. In: Mercer, M.C. (ed.) Multispecies Approaches to Fisheries Management. Canadian Special Publication of Fisheries and Aquatic Science No.59, pp. 116–118. Pope, J.G., Stokes, T.K., Murawski, S.A. and Idoine, J.S. (1987) A comparison of fish sizecomposition in the North Sea and on Georges Bank. In: Wolff, W., Soeder, C.-J. and Drepper, F.R. (eds) Ecodynamics – Contributions to Theoretical Ecology. Springer-Verlag, Berlin, pp. 146–152. Quero, J.C. (1998) Changes in the Euro-Atlantic fish species composition resulting from fishing and ocean warming. Italian Journal of Zoology 65 (Supplement), 493–499. Rijnsdorp, A.D. and van Leeuwen, P.I. (1992) Density-dependent and independent changes in somatic growth of female North Sea plaice (Pleuronectes platessa L.) between 1930 and 1985 as revealed by back-calculation otoliths. Marine Ecology Progress Series 88(1), 19–32. Rijnsdorp, A.D., van Leeuwen, P.I., Daan, N. and Heessen, J.L. (1996) Changes in abundance of demersal fish species in the North Sea between 1906 and 1990–1995. ICES Journal of Marine Science 53, 1054–1062. Roberts, C.M. (1995) Effects of fishing on the ecosystem structure of coral reefs. Conservation Biology 9(5), 988–995. Rosen, D.A.S. and Trites, A.W. (2000) Pollock and the decline of Steller sea lions: testing
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the junk-food hypothesis. Canadian Journal of Zoology 78, 1243–1258. Rudstam, L.G., Aneer, G. and Hilden, M. (1994) Top-down control in the pelagic Baltic ecosystem. Dana 10, 105–129. Sainsbury, K.J. (1991) Application of an experimental management approach to management of a tropical multispecies fishery with highly uncertain dynamics. ICES Marine Science Symposium 193, 301–320. Sainsbury, K.J., Campbell, R.A., Lindholm, R. and Whitelaw, A.W. (1997) Experimental management of an Australian multispecies fishery: examining the possibility of trawl-induced habitat modification. In: Pikitch, K., Huppert, D.D. and Sissenwine, M.P. (eds) Global Trends: Fisheries Management. American Fisheries Society, Bethesda, Maryland, pp. 107–122. Schwartzlose, R.A., Alheit, J., Bakun, A., Baumgartner, T.R., Cloete, R., Crawford, R.J.M., Fletcher, W.J., Green-Ruiz, Y., Hagen, E., Kawasaki, T., Lluch-Belda, D., Lluch-Cota, S.E., MacCall, A.D., Matsuura, Y., NevárezMartínez, M.O., Parrish, R.H., Roy, C., Serra, R., Shust, K.V., Ward, M.N. and Zuzunaga, J.Z. (1999) Worldwide large-scale fluctuations of sardine and anchovy populations. South African Journal of Marine Science 21, 289–347. Sherman, K., Jones, C., Sullivan, L., Smith, W., Berrien, P. and Ejsymont, L. (1981) Congruent shifts in sand eel abundance in western and eastern North Atlantic ecosystems. Nature 291, 486–489. Shiganova, T.A. and Bulgakova, Y.V. (2000) Effects of gelatinous plankton on Black Sea and Sea of Azov fish and their food resources. ICES Journal of Marine Science 57(3), 641–648. Skjoldal, H.R., Gjøsæter, H. and Loeng (1992) The Barents Sea ecosystem in the 1980s: oceanic climate, plankton, and capelin growth. ICES Science Symposium 195, 278–290. Sparholt, H. (1994) Fish species interaction in the Baltic Sea. Dana 10, 131–162. Stefánsson, G., Skúladóttir, U. and Steinarsson, B.Æ. (1998) Aspects of the ecology of a Boreal system. ICES Journal of Marine Science 55, 859–862. Steneck, R.S. (1997) Fisheries-induced changes to the structure and function of the Gulf of Maine ecosystem. In: Proceedings of the Gulf of Maine Ecosystem Dynamics Scientific Symposium and Workshop. RARCOM Report 91–1. Regional Association for Research in the Gulf of Maine, Hanover, New Hampshire, pp. 151–165. Stevens, J.D., Bonfil, R., Dulvy, N.K. and Walker, P.A. (2000) The effects of fishing on sharks, rays, and chimaeras (Chondrichthyans), and the
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implications for marine ecosystems. ICES Journal of Marine Science 57, 476–494. Suvapepun, S. (1991) Long-term ecological changes in the Gulf of Thailand. Marine Pollution Bulletin 23, 213–217. Svendsen, E. and Magnusson, A.K. (1991) Climate variability in the North Sea. ICES Marine Science Symposia 195, 144–158. Tasker, M.L., Camphuysen, C.J., Cooper, J., Garthe, S., Montevecchi, W.A. and Blaber, S.J.M. (2000) The impacts of fishing on marine birds. ICES Journal of Marine Science 57, 531–547. Tegner, M.J. and Dayton, P.K. (2000) Ecosystem effects of fishing in kelp forests communities. ICES Journal of Marine Science 57, 579–589. Thiebaux, M.L. and Dickie, L.M. (1993) Structure of the body-size spectrum of the biomass
in aquatic ecosystems: a consequence of allometry in predator-prey interactions. Canadian Journal of Fisheries and Aquatic Science 50, 1308–1317. Wespestad, V.G., Fritz, L.W., Ingraham, W.J. and Megrey, B.A. (2000) On relationships between cannibalism, climate variability, physical transport and recruitment success of Bering Sea walleye pollock (Theragra chalcogramma). ICES Journal of Marine Science 57, 272–278. Yaragina, N.A. and Marshall, C.T. (2000) Trophic influences on interannual and seasonal variation in the liver condition index of Northeast Arctic cod (Gadus morhua). ICES Journal of Marine Science 57, 42–55. Yodzis, P. (2000) Diffuse effects in food webs. Ecology 81(1), 261–266.
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Anthropogenically Induced Changes in the Environment: Effect on Fisheries Katherine Richardson Department of Marine Ecology, University of Aarhus, Aarhus, Denmark
Abstract In considering ‘responsible fisheries’, the focus usually is on potential effects of fisheries on ecosystems and habitats (i.e. effects of fisheries on the environment). For most, the term ‘responsible fisheries’ implies a need for a change in fishing practices to protect or improve the state of the environment. The state of the environment, however, also inevitably affects fish and, thus, fisheries. Many societal activities influence the state of aquatic environments. Thus, in moving towards ‘responsible fisheries’, changes in other societal activities rather than fisheries alone may also need to be considered. The indirect (i.e. non-fishing) effects on fish and fisheries can be divided into two types: those that affect ecosystem structure or population processes such that recruitment to the fishable stock is reduced, and those that affect the quality (and, hence, marketability) of the fish product. Environmental changes that can influence recruitment include land use changes that may alter habitats for fish. Damming and re-routing of streams and rivers may reduce access to spawning grounds for fishes that migrate between salt and freshwaters. Erosion (leading to increased turbidity) and eutrophication also lead to changes in habitats and food availability that can affect recruitment. Intentional and non-intentional introductions of new species to a region can alter ecosystems to the point that fisheries are severely affected. All of these influences are well described for individual stocks or local regions. However, a global assessment of the quantitative impact of such changes on fisheries is lacking. Chemical contamination of aquatic ecosystems can influence the physiology of organisms and accumulate in body tissues. Thus, both the recruitment and the marketability of fish products may be affected by contaminants. Many studies dealing with the potential toxicity of contaminants to physiological processes at the cell and organism level have been carried out. However, few studies have dealt with the effect of contaminants at the population level and thus attempted to quantify the effects of environmental contamination on fisheries. Monitoring of contaminant concentrations in fish meat is, in many regions of the world, standard protocol as part of public health protection measures, and some fisheries, especially in fresh and semi-enclosed marine waters, have been restricted as a result of such contamination. Although, in most cases, contaminant concentrations in wild fishes have been found to be below the levels considered to be safe for human consumption, recent studies have shown that, for example, dioxin contaminant levels in the muscle of wild fishes in the EU is higher than those found in meat produced in commercial agriculture. As knowledge concerning the effects of contaminants on human physiological processes increases, the contaminant concentrations considered as safe for human consumption are being reconsidered and, in some cases, reduced. Thus, the fact that wild fish meat is among the most contaminated with respect to dioxin of the common meat protein sources for the human population suggests that environmental effects on fisheries will be an area of increasing concern in coming years.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Introduction Fish and shellfish for human and animal consumption are taken through harvest of natural stocks, harvest of ‘ranched’ (released) stocks, and a variety of aquaculture activities. All of these types of fishing activities affect the environment and, for most, the term ‘responsible fisheries’ implies efforts to ensure that the environmental interactions and consequences of fishing activities are brought into or kept within levels deemed acceptable by current standards of society. Thus, there are both natural science (identification of environmental effects of fisheries) and social science (quantification of the acceptable social norms) components to the consideration of ‘responsible fisheries’. There is, however, yet another aspect of natural science investigation that is relevant to, but often overlooked in, the consideration of responsible fisheries: changes brought about by human activities in the state of the environment and the influence of these changes on the health and size of fish stocks and/or the marketability of fish products. That the state of the environment affects both the survival and recruitment of fish stocks is well known, and considerable research effort has been devoted to a description of the effects of fluctuations in, for example, temperature, salinity, food and predator abundance on fish stocks. The assumption in considering the influence of these fluctuations in environmental variables on fish stocks is, as a rule, that the root causes of the observed fluctuations are not under anthropogenic control. There is, however, increasing evidence that human activities are influencing climate as well as other aspects of the environment important for the recruitment and survival of fish stocks (for a review, see Steffen and Tyson, 2001). Responsible fisheries should, then, include ‘responsible (i.e. socially acceptable) care or conservation’ of the stocks to be fished. Thus, the influence on fish stocks of anthropogenically induced changes in the state of the environment is an obvious and important consideration when developing a strategy for carrying out ‘responsible fisheries’.
The state of the environment can influence fisheries in two fundamentally different ways: either by changing the state of the stock to be fished (numbers, condition and/or health), or by affecting the marketability of the fish products, themselves (i.e. contaminant content in excess of public health standards for human consumption). Of course, natural as well as anthropogenically induced changes in the state of the environment influence the status of fish stocks. Thus, a major challenge in quantifying human impact on the status of fish stocks is differentiating the anthropogenically induced from the naturally occurring changes in the state of the environment. In some cases, especially those involving construction projects (i.e. dams) that have impeded the passage of diadromous species between marine and freshwaters (e.g. salmon fishes in many regions of the world), it is relatively easy to quantify the human influence on fish stocks and fisheries. However, in most cases, differentiating the natural and anthropogenically induced environmental change signals from one another is not yet possible. In the case of the influence on fisheries of the introduction via human activities of contaminants to aquatic systems, separating the anthropogenically induced from the natural signal in environmental change is more straightforward. Here, however, the scientific challenge is quantifying the effect of contaminants at the level where fisheries may be affected (i.e. the population level). Tens of thousands of studies have been carried out during recent decades in which the effects of various contaminants on the physiology of fish and shellfish have been examined. However, in almost all cases, the effects being examined have been at the level of the cell or individual organism. Studying effects of contaminants at these levels is essential in order to understand the mechanisms by which contaminants may influence fish stocks. However, it is not possible to extrapolate results directly from studies carried out at the level of the cell or individual to the fisheries-relevant level of populations. Again, this is, at least in part, because of the fact that the population is being influenced by environmental factors other than the contaminant exposure alone.
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The purpose of this chapter is to review the current state of knowledge with respect to the effect on fishes and fisheries of anthropogenically induced changes in the environment. The primary focus is contaminants and their effects on fish and fisheries. However, the influences of other human activities on fish and fisheries, as well as some efforts to quantify the effects in economic terms of selected activities or events on specific fisheries, are also briefly reviewed in order to present a complete outline of the potential influence of non-fisheries-related human activities on fisheries and to provide the reader with relevant references for further study of this complex aspect of responsible fisheries.
Land Use Changes and Their Impacts on Fisheries The largest and, therefore, economically most important commercial fisheries in the world are marine. The oceans also extend over large regions of the earth and are characterized by large water volumes and rapid water circulation. This means that the footprint of human activities on the environment will be eroded relatively easily in marine environments. Thus, marine regions and the most important world fisheries will be the last of the aquatic ecosystems to be obviously affected by human activities. Nevertheless, there is now evidence from many coastal and semi-enclosed sea areas that changes in land use practices are influencing the state of the marine environment and, potentially, fisheries (e.g. Caddy, 1993). This realization has put the consideration of anthropogenically induced changes in the environment on the global agenda. However, examples of the influence of land use change on fish and fisheries in freshwater systems are plentiful and well documented. Loss of or changes in fish stocks resulting from human activities have been documented in thousands of lakes and rivers (Maitland, 1995). In considering the potential influence of human activities on marine
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fisheries, much can be learned from the historical experience in freshwater systems (e.g. Caddy, 1993). From freshwater systems, it is well documented that land use changes can seriously affect fish stocks and fisheries (see, for example, Maitland, 1995). River obstructions such as dams can block migration routes for diadromous species and change habitats in such a manner as to make them unsuitable for fish stocks (i.e. increase sedimentation in spawning habitats). Land drainage schemes can change (or even eliminate) the water content of ponds and lakes and alter flow and siltation rates in adjacent waterways. Land use changes that increase erosion to surrounding waters can affect not only siltation rates but also the turbidity of the water, that may influence predator–prey interactions crucial to the survival of fishes. Farming activities lead to the runoff of nutrients and pesticides to adjacent water bodies. This nutrient enrichment can lead to eutrophication. In freshwater systems, this phenomenon has been well studied, and it is now realized that eutrophication is also occurring in many coastal marine areas (see Jørgensen and Richardson, 1996). In its mildest form, it affects fish and fisheries by changing the availability and type of food for the fishes. In some cases, mild eutrophication may increase food availability for foodlimited fish stocks and actually increase the fish available for harvest by the fishery (e.g. Nielsen and Richardson, 1996). Thus, anthropogenically induced changes in the environment need not necessarily be negative for fisheries. In more severe cases of eutrophication, however, hypoxia and anoxia develop, making the environment uninhabitable for some or all fishes. There is also some evidence (see Richardson, 1997) that nutrient enrichment of coastal areas may have increased the incidence of algal blooms. Some of these blooms contain ichthyotoxins that have a direct impact on fish. Others contain toxins that endanger the human consumer of fish or shellfish products. Thus, the presence of these latter algal toxins in fish or shellfish can affect the marketability of these products.
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Introduction of Non-indigenous Species: Impact on Fisheries Non-indigenous species can be introduced intentionally (i.e. with the intent of developing a new fishery) or non-intentionally (with ballast water; through biofouling of ship hulls, etc.; the building of canals; or in association with organisms migrating or actively transferred between different regions). Most non-indigenous species do not survive when introduced to a new environment. However, there are numerous examples of introductions that have had unexpected and dramatic consequences on fisheries. The most often quoted example of fishery effects following the introduction of a new species is that of the Black Sea, where the ctenophore, Mnemiopsis leidyi, was introduced (presumably with ballast water) in 1992. Russian scientists documented that in some years following the introduction, up to about 95% of Black Sea wet weight biomass was comprised of this organism (Travis, 1993). In the Azov Sea, fish (primarily anchovy) catches dropped by an estimated 200,000 t, with the loss to the fishery estimated at US$250 million (Travis, 1993). Another important and well-documented example of an accidentally introduced alien species influencing a fishery is the introduction, in the 1970s, of Gyrodactylus salaries into Norwegian rivers with salmon released from infected hatcheries. The Gyrodactylus acts as a parasite on salmon and has seriously reduced catches in infected rivers (Johnsen and Jensen, 1991). This parasite continues to pose a problem for the sports fishery on salmon in Norway. As it affects a sports fishery, the economic consequences of the introduction of this species are far reaching and include the tourism industry and the local communities that rely on this tourism. Intentional introductions of alien species can also have unpredicted effects on the ecosystems into which they are introduced, including effects on other fish species and fisheries. Some of the best documented examples of intentional introductions are those of the Nile perch (Lates niloticus) and the Nile tilapia (Oreochromis niloticus) and other tilapia
to the large African lakes, Lake Victoria and Lake Kyoga, in the 1950s and 1960s. These species were introduced with the purpose of establishing a fishery on these species and, indeed, successful fisheries were established on these species. Concomitant with the development of the new fisheries, however, the stocks of many of the indigenous fishes collapsed. These collapses have been widely attributed to predation by and competition with the introduced species. This interpretation has, however, been questioned by some authors (e.g. Ogutu-Ohwayo, 1990; Kudhongania et al., 1992). Both of these studies suggest that lack of effective fisheries management, the use of destructive gear and overfishing may also have had a profound effect on the state of the indigenous stocks. It has also been pointed out that even before the introductions, eutrophication was altering the lakes (Kaufman, 1992) and that eutrophication effects may also have contributed to the demise of the indigenous stocks. Thus, while it is not possible to isolate the effect of the introduction of the new species on the indigenous stocks in these lakes, the introduction of the new species was certainly one of a number of anthropogenic modifications of the environment that, ultimately, led to the collapse of the indigenous stocks. Far fewer introductions of fish species have occurred in marine waters than in freshwaters. Nevertheless, at least 120 fishes have been introduced to marine waters foreign from their source of origin (Baltz, 1991). Many of these introductions were intentional, and introductions designed to improve fisheries in marine waters have been carried out since the late 19th century. The Soviet Union was, during the 20th century, most active in terms of attempting introductions of fish species to try to create new fisheries. In all, 42 attempts at introducing 29 different species were made by the Soviet Union. Of these attempts, large populations of the introduced species developed in 11% of the cases, and small populations in 5%. In many cases, however, the introductions had a serious impact on the endemic fish species. This was particularly true in closed ecosystems (Baltz, 1991).
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An interesting modification on the theme of introduction of alien species is when a genetically modified (either by selective breeding in captivity or by actual gene manipulation) organism is released to the wild and begins interacting and breeding with its wild counterpart (e.g. salmon in the River Vosso, Norway; Sægrov et al., 1997). Technically speaking, this example does not represent the introduction of a non-indigenous species as the salmon are already a part of the river ecosystem. However, the replacement of wild stock with the reared salmon does have potential repercussions for the fishery as the genetic composition of the reared salmon may not be as robust over time for environmental variability as the wild stocks. Thus, recruitment in mixed wild–reared stock potentially could differ from that which would be achieved by the unadulterated wild stock.
Impact on Fisheries of Conservation Measures Changes in the relative abundance of organisms in the various trophic levels of the marine food web will, clearly, influence the function of the web. In recent years, large predators that feed on fish, such as marine mammals and birds, have been the object of considerable conservation interest. Protection measures have been taken in the EU, for example, with respect to whales, seals and many marine birds. In a number of cases, these protection measures have been highly successful, and the previously diminished or endangered stocks of top predators are now thriving. Examples are some seal populations in the Baltic, and cormorants along the Baltic and Kattegat coasts. As these large predators often feed exclusively on fish, an increase in population size will certainly imply a greater predation pressure on fish stocks affected by these predators. The magnitude of the increased predation pressure on fish stocks due to increased predator abundance has not been determined. However, some – especially coastal fishers working, for example, in areas in close proximity to large
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cormorant colonies – are convinced that the competition for fish resources between these large predators and fishers is intense. Thus, for these fishers, at least, the conservation of species at the higher trophic levels of marine food webs represents an anthropogenically induced environmental effect on fisheries. In the wake of the Earth Summit Meeting in 1992 and the subsequent ratification by many countries (178 as of August 2000) of the Biodiversity Convention, whereby countries commit themselves to the preservation of the Earth’s biodiversity, many countries are establishing or considering the establishment of marine protected areas, where mechanical intervention in the ecosystem (i.e. fishing activities) is restricted and nature is allowed to develop according to its own premises. Although there are anecdotal reports of improved fisheries in waters abutting such protected areas, most fishers perceive the establishment of marine protected areas as an anthropogenically induced environmental effect on (restriction of) fisheries.
Conflict with Other Users of Aquatic Ecosystems: Impact on Fisheries A number of fishers are concerned about the possible influence on fish behaviour and/or survival of disturbance caused by nonfishing users of aquatic environments. For example, seismic activities carried out in association with oil and gas prospecting have attracted particular attention from fishers in some areas, as activities that may potentially interfere with fish survival and behaviour. The significance of seismic activity for fisheries has not been quantified, although some workers have reported mortality of larval fishes located in close proximity to the air guns used to generate acoustic signals (Dalen and Knudsen, 1987). Sand and gravel dredging, either for the purpose of extracting the sand and gravel for use on land or as a part of construction projects (i.e. establishment of artificial islands or bridges) at sea can also affect fisheries by, for example, increasing water turbidity,
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frightening fish away and/or releasing contaminants to the water column from disturbed sediments. DeGroot (1979) estimated the cost to the Dutch shrimp and sole fisheries of a proposed establishment of some artificial islands in the North Sea at 10 million guilders.
Contaminants in the Environment: Impact on Harvestable Marine Organisms General responses to contamination Contaminants are substances that are toxic to living organisms. They may occur naturally in the environment or they may be chemicals produced through human activities. Both naturally occurring and anthropogenically introduced contaminants can influence fisheries. For example, Shilts and Coker (1995) report on a commercial trout fishery in a remote area which had to be closed due to high levels of mercury from a geological source (i.e. mercury not anthropogenically introduced into the environment). Some contaminants are required at low levels in living organisms (i.e. metals such as copper, zinc and iron) and only become toxic at high concentrations. Metals such as cadmium, lead and mercury have no biological function and generally are toxic at much lower concentrations than copper, zinc or iron. Contaminants are often referred to as ‘xenobiotics’ or ‘micropollutants’. Of the contaminants produced by humans, the most toxic are generally the chlorinated or brominated compounds. Examples of these include pesticides (DDT, lindane), brominated flame retardants, dioxins and PCBs (polychlorinated biphenyls). These compounds are very persistent in the environment – which explains why chlorinated and brominated compounds are found in Arctic animals – far from the sites of production and entry into the environment. Two different approaches can be used to examine the effects of contaminants on fishes. The effect of individual contaminants on fish can be examined in toxicity studies carried out in the laboratory, where the fish are exposed
to controlled concentrations of the contaminants under investigation. Such studies can help elucidate the responses to specific contaminants at the level of the cell or individual. However, in nature, fish are not exposed simply to single contaminants but to a variety of natural and anthropogenically induced environmental stresses. Therefore, laboratory exposure studies are seldom very informative with respect to responses to contamination at the population level (i.e. the level relevant to fisheries). For understanding responses at this level, field studies are much more useful. Several general observations can be made concerning the results from field studies comparing fish from areas of relatively high and low contamination, respectively. The first concerns fish health. Individual fish have always been susceptible to disease and infection. However, there are increasing reports of disease outbreaks involving large numbers of fish. Many studies (e.g. Malins et al., 1984, 1985; Couch and Harshbarger, 1985; Murchelano and Wolke, 1991; Vethaak and ap Rheinallt, 1992; Chu and Hale, 1994; Myers et al., 1994; Vethaak and Jol, 1996) have shown correlations between contamination levels or exposure to water taken from contaminated regions. As noted, cause and effect are not necessarily demonstrated by such correlation studies, but toxicity studies carried out under controlled laboratory conditions (see below) are now identifying physiological mechanisms whereby the immunological system of fishes is influenced by some contaminants, thus confirming a link between fish health and contaminant exposure. Another general observation that can, in some cases, be made at the population level is endocrine disruption, which occurs when natural or synthetic chemicals interfere with the normal hormonal activity in wildlife. Among the synthetic chemicals that may cause endocrine disruption are PCBs and organochlorine pesticides. The most publicized example of endocrine disruption is that of the influence on some marine gastropods of tributyltin (TBT), a herbicide contained in anti-fouling paints. The TBT induces imposex in these snails and, ultimately, may reduce or remove the population’s ability to reproduce
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itself. This phenomenon was first discovered in the North Sea on gastropods not used in fisheries or aquaculture. However, in, for example, southeast Asia, where gastropods are used more extensively in aquaculture, endocrine disruption caused by TBT contamination potentially could directly affect the production from these aquaculture facilities. For finfish, endocrine disruption has been most noted and studied in freshwater fishes. Downstream of some sewage treatment plants in the UK, male fish have been observed to produce the yolk precursor protein, vitellogenin (Harries et al., 1996, 1997). Natural and synthetic steroid human hormones that are found in sewage effluent appear to be an important source of potentially endocrine-disrupting compounds in the environment (Desbrow et al., 1998). The occurrence of endocrine disruption in the natural environment has been realized only recently, and its ecological significance, if any, has not yet been quantified. Another general response to contamination (and one which may be related to endocrine disruption) is alteration in reproductive behaviour. Relatively few studies have considered reproductive behaviour in relation to contamination: Jones and Reynolds (1997) reviewed the existing literature and found that, out of the about 20,000 reported scientific investigations on the effects of contaminants on fishes that have been carried out over the last two decades, only about 0.1% have considered the effect of contaminants on reproductive behaviour. Behaviour abnormalities that have been noted in association with contaminants or thermal pollution include effects on courtship (changes in the frequency of displays or the duration of courtship or male-like activity in females) and parental care (changes in nest-building activity, defence of the young or the parental care role). Bioassays where test fishes are monitored for changes in reproductive behaviour following exposure to potentially contaminated water have been suggested as possible supplementary tests in routine monitoring for contaminants (see review by Jones and Reynolds, 1997). Finally, it should be noted that, in some freshwater systems, a general acidification
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resulting from industrial contamination of the atmosphere and subsequent acidification of precipitation has had serious ramifications for fishes and fisheries. Arctic char may be particularly sensitive to acidification (Hesthagen and Sandlund, 1995). The whole process of lake acidification demonstrates that fish and fisheries need not be in the immediate vicinity of an environmental contamination source to be affected. The industrial emissions responsible for the acidification of rain are often located far from the site of the resulting acidification. Marine systems are less prone to acidification than freshwater systems because of the carbonate buffering system in seawater.
Specific types of contamination and their effects
Pesticides Pesticides are usually organochlorines and can be extremely toxic to fish. An estimated 50,000 kg of dead fish were reported in 1985 in the Miranda River in South America as a result of pesticide exposure, although, in that case, the pesticide may have been introduced deliberately to the river as part of a political gesture (Alho and Vieira, 1997). Pesticides are considered to be one of the main reasons for the demise of the commercial fishery in the Azov Sea. In the late 1980s, pesticide input to this sea was about 100,000 t year−1 and consisted primarily of compounds from the DDT group (Semenov et al., 1998). Pesticides have a recognized effect on biology, although the mechanism of the effect may be different for each stage of an organism’s development (Kobayashi et al., 1991). Pesticide exposure (to chlordane, DDT, lorsban and lindane) decreases the protein content in shrimp and increases their respiration (Reyes et al., 1996). Concern about the potential effects of pesticides on the production of fish products is particularly relevant for countries, such as Egypt, where aquaculture activities are carried out using runoff water from agricultural activities and, indeed, considerable research concerning the effects of pesticides on fishes used in
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aquaculture has been carried out here (e.g. Shalaby et al., 1995a,b). TBT is a well known herbicide used in anti-fouling paints. However, organotins are also used as stabilizers in PVC and can enter the environment in connection with PVC production. Organotins are ubiquitous in the marine environment and can threaten coastal fisheries (e.g. Davis et al., 1999; Manning et al., 1999). Fish and crustaceans have enzymes that may facilitate a more rapid elimination of TBT than molluscs and other organisms (Lee, 1991), and TBT may be less prone to bio-accumulation than some other organotin compounds. Nevertheless, organotin concentrations in fishes collected in The Netherlands suggest that the survival of these fishes might be threatened as a result of the concentrations they had accumulated (Stäb et al., 1996). Other studies have also suggested that fish population exposed to naturally occurring concentrations of TBT may suffer effects that could influence the fishes at the population level. The maximum concentrations of TBT in Chesapeake Bay have been reported to be sufficient to lead to larval mortality and, hence, influence recruitment in striped bass (Pinkney et al., 1990). Potential risks of mortality for both fish and zooplankton have been calculated at the concentrations of TBT predicted in marinas during and after TBT clean-up operations (Traas et al., 1996). TBT exposure has also been related to a reduction in the efficiency of the immune response in fish. Exposure of flounder to TBT concentrations similar to the highest found in nature caused mortality after 7–12 days. This was associated with gill lesions, reduced immune activity and a reduction in the volume of the thymus (Grinwis et al., 1998, 2000). These same authors have suggested that TBT might have a causal role in the prevalence of lymphocystis viral infections in nature. Also, shellfish have been identified as showing a reduction in immune efficiency following exposure to TBT. Exposure increased susceptibility of oysters to a parasite pathogen (Perkinsus marinus). In addition, animals exposed to TBT may succumb at lower levels of pathogen infection than non-TBT-exposed animals (Fisher et al., 1999).
Oil contamination In areas with chronically high concentrations of oils, mortalities of fish and shrimp fry have been recorded (Ramamurthy, 1991). Oil spills provide a unique opportunity to examine responses of fish populations to specific contamination events rather than to chronic exposure of a contaminant that is more common. There often are also resources available for monitoring of biological impacts following an oil spill. Perhaps the most dramatic effect of an oil pollution event on fisheries occurred in the Saudi Arabian shrimp fishery following the 1991 Gulf War. In the Gulf, the effect of oil spills was most noticeable along the Saudi Arabian coast. In the summer of 1991 (i.e. following the oil spills occurring in connection with the war in February 1991), the landings of shrimp fell dramatically and the landings contained only a small proportion of sexually mature adults (Mathews et al., 1993). In the subsequent period, spawning stock biomass fell to approximately 1.8% of pre-war level and the total biomass to less than 1.5% of the pre-war biomass (landings in 1989, 4000 t; 1992, 25 t). A man-made recruitment failure of the shrimp resulting from the oil spills is suspected. However, due to lack of monitoring data and research activity, the mechanism(s) leading to the recruitment failure has(have) not been identified (Mathews et al., 1993). Other oil spills have also provided opportunities to investigate the influence of oil contamination on fish and shellfish, at both the individual and the population level. Squire (1992) considered possible effects of a 1969 oil blowout in the Santa Barbara Channel on three pelagic fisheries: northern anchovy, jack mackerel and Pacific bonito. Although the latter two species exhibited a reduced abundance in the area during the months immediately following the oil blowout, Squire concludes that pelagic fish species are relatively lightly affected by such a contamination event. Following the Exxon Valdez spill in 1989 in Prince William Sound, the suitability of the sediments as spawning grounds for salmon was monitored for a number of years. Leaching of polycyclic aromatic hydrocarbons (PAHs)
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from the substrates in several rivers exceeded the concentrations believed to be lethal to salmon embryos up to and including the year 1993 (Murphy et al., 1999). The fish and shellfish fisheries in proximity to the Sea Empress oil spill in Wales (1996) were closed as a precautionary measure for consumers. Associated monitoring showed that fish and crustacea only took up small concentrations of PAHs, but molluscs took up whole oil into their tissue. This and a number of earlier studies indicate that filter feeders accumulate oil more readily than other feeding types (Law et al., 1999). The responses of both salmon and dab were recorded in the aftermath of the Braer oil spill at the Shetland Islands in Scotland in 1993. Salmon, and to some extent dab, responded immediately to the spill by producing detoxification enzymes. The fact that dab were less inclined to produce these enzymes may suggest that the PAHs harboured in the sediment following the incident were not directly bio-available to the fish. There was also some evidence of liver pathology in the dabs taken from the most contaminated sites in the year following the spill. However, a cause-and-effect relationship between the oil contamination and the liver pathology was not established (Stagg et al., 1998). Finally, it has been suggested that offshore drilling may be a possible source of hormone mimics or endocrine disrupters in the marine environment, as PAHs and alkylphenols are released through these activities. Newer studies, however, suggest that, in the North Sea, accumulation of endocrine disrupters in wild fish is more likely to be as a result of estuarine exposure to these chemicals rather than from offshore oil extraction activities (Lye, 2000).
Contamination of Fisheries Products and Public Health Standards Heavy metal contamination was the first form of contamination recognized as a threat to human health and, in many areas, fish products are monitored routinely for heavy
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metal contamination. Likewise, many areas have established public health standards that fish products, by law, must meet. Many examples can be cited from around the world of fish products containing heavy metals in concentrations that exceed established public health standards. For example, choosing examples randomly to show the geographic extent of the problem: 50 and 35% of fishes from two rivers in the Pantanal (floodplain western Brazil and Bolivia) have mercury levels above the recommended standards (Alho and Vieira, 1997). Eels in some eastern English rivers exceeded standards for metal contamination but, as eels generally are not eaten in England, this is not considered to be a problem (Barak and Mason, 1990). Molluscs in Ria Formosa Lagoon, Portugal, contain metal concentrations considered unsafe for human consumption (with mass mortalities attributed to environmental degradation, but mechanism not specifically described) (Bebianno, 1995). As indicated, these examples are far from unique. Most countries can identify areas near harbours or industrial effluent outlets where metal contamination is a problem, especially for shellfish and more stationary fish species that remain in the area of contamination. In some areas, such as the Aswan Lake, metal concentrations in fish are used as bioassay indicators for contamination (Rashed, 1999). Mercury contamination in fish is a problem being awarded increasing public awareness. Recent studies have indicated that children of mothers in the Faeroe Islands with a heavy body burden of mercury (in the case of Faeroese women, resulting primarily from the consumption of whale meat) demonstrate a statistically significant reduction in some neurological processes (Grandjean et al., 1998). In the USA, health authorities have warned pregnant women and women of childbearing age who might become pregnant not to eat more than 12 oz of certain types of fatty marine fish per week because of mercury contamination (Anon., 2001). Thus, despite the fact that the dangers associated with consumption of mercury-contaminated fish have been known for decades and legislation has been in effect to reduce the input of
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mercury and other metals to the aquatic environment, the perceived public health risks associated with eating wild fish are, due to mercury contamination, greater than they have ever been. This is, of course, due in part to a greater awareness of effects of these metals at low concentrations, but can also be attributed to a greater accumulation of metals in some marine organisms than could be predicted on the basis of contamination levels alone. Similar concerns are now also being raised concerning the dioxin concentration in wild and cultured fishes. Dioxins enter the environment with industrial waste and as products of incineration, and they are considered to be carcinogenic. A recent report from the EU (http:/europa.eu.int/comm./ food/fs/) indicates that contamination by dioxin of fish products (both wild and farmed) in the EU is much greater than in commercially farmed terrestrial sources of meat protein. The contamination is greatest in fatty fishes and fish meal, and for fishes coming from the Baltic and North Seas. The fact that the aquaculture industry relies on fish meal as a staple in the feeds used in culturing activities almost certainly explains the high concentrations recorded in cultured fishes. A European official has been quoted in the popular press (Simons, 2000) as follows: ‘Nobody is saying we can’t eat fish anymore, but consumers must be made aware that fish contributes significantly to the intake of dioxins’ and ‘. . . if you eat fish every day, you are likely to have a problem.’ In contrast, radioactive contamination of fishes seems unlikely to pose a threat to the public health quality of fish products. Radionuclides originating from anthropogenic activities enter the marine environment from nuclear weapons testing, global fallout, releases from nuclear facilities, dumping of radioactive waste and accidental delivery (nuclear submarine and aircraft accidents at sea, as well as terrestrial accidents such as the Chernobyl incident). Considerable public concern and attention have been focused on radioactive contamination of the sea and its resources (see, for example, Klungsøyr et al., 1995). Not surprisingly, the concentrations of
anthropogenically introduced radioactivity (137Cs) are highest in those seas nearest the site of the Chernobyl accident (Baltic, Irish, Black Seas and the northeast Atlantic). Nevertheless, a recent review suggests that, even in these areas, the contribution of anthropogenically introduced radioactivity to the total radioactive contamination of marine fishes is very small (100–1000 times lower) and that the radioactivity consumed with marine fishes is well below the established ‘safe’ levels for consumption (Livingston and Povinec, 2000).
Quantifying the Consequences of Anthropogenic Activities on Fisheries Marine resource availability is, of course, dependent on many factors. These include climate, availability of food and abundance of predators. Much research has been focused on identifying the influence of factors such as temperature, food availability and ecosystem structure (including predator abundance) on recruitment to the fishable stocks. Much less effort has been devoted to assessing the impact of anthropogenic activities (both terrestrial and marine) on fish stocks and fisheries. Clearly, however, the status of fish and shellfish stocks is not independent of the conscious and unconscious use by humans of aquatic environments. Quantifying the effect(s) of environmental degradation due to anthropogenic activities is not always straightforward. Boreman (1997) discussed methods of comparing the impacts of habitat degradation due to pollution as opposed to fishing of fish populations. He concludes that the science of aquatic environmental impact assessment on fish stocks is clearly not yet at the same level as traditional stock assessments, and admonishes administrators not to use additional harvest restrictions as an alternative to trying to understand further the environmental influences on fish stock sizes. Lipton and Strand (1997) provide a very accessible overview for the non-economist
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of the factors to be taken into account and the theory behind estimating the economic costs for fisheries of a pollution-related event. These authors point out that a major problem in making such estimates is the lack of any baseline data to describe what the earnings would have been from the lost fishery. Another problem they identify is that the economic loss with respect to a polluted fishery will also depend on the availability of alternative fishing sites. Using standard economic modelling techniques, the losses will be greater for those fishers who have no alternative fishing area than those who have, even though the cost in biological terms will be the same following a pollution event regardless of proximity to alternative fishing grounds. Using standard economic models, several authors have attempted to place a market value on the economic losses experienced by fisheries as a result of specific environmental catastrophes or pollution events (e.g. Kahn, 1987; Clites et al., 1991). Estimating the economic consequences of oil spills and accidents has attracted considerable attention from the economic community, and a number of models for making these estimates have been developed (e.g. Reed et al., 1984; Spaulding et al., 1985; Gringalunas et al., 1986, 1988). Heen and Andersen (1994) describe a general approach for identifying regional impact of environmental incidents such as oil spills on, for example, fish farms. Cohen (1995) estimated the economic losses of the 1989 Exxon Valdez spill to the south-central Alaskan Fishery (salmon, shellfish, herring, halibut and groundfish). The upper estimates of cost established in that study were US$108 million during the first year and US$47 million during the second year. Kahn and Buerger (1994) considered the effects of two different forms of environmental stresses on the striped bass fishery of the Hudson river: general environmental deterioration of the river; and PCB contamination of the bass. They concluded that the environmental deterioration of the river generated annual losses of US$2.3–7.7 million and the PCB contamination between US$0.745 and 3.7 million.
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Conclusions Anthropogenically induced changes in aquatic environments are having an effect on natural fish populations and, ultimately, on fisheries. Usually, however, these effects are not possible to quantify. This often is because natural changes or fluctuations in the environment are occurring simultaneously with the anthropogenically induced changes. Fish are exposed to a variety of environmental stressors at all times, and identifying the effect of individual stressors on the fishable population is seldom feasible. This makes the signal from the anthropogenically induced changes very difficult to separate from the signals of other stressors. This is especially true in cases where the anthropogenically induced effect is related to ecosystem interactions that are poorly described or understood. It is more feasible to estimate the (economic) consequences of individual environmental incidents resulting from human activities (such as oil spills) on specific fisheries, and a number of economic models have been developed for that purpose. The effects of human activities on fish and fisheries are most obvious and easiest to elucidate in enclosed freshwater systems and in estuarine systems where the impact of human activities is greatest. However, there is now considerable evidence that wild fishes in marine stocks – even in what previously has been considered open waters, relatively free from human impact – are also influenced by changes in the environment brought about by human activities. Much of this evidence is in the form of chemical contamination. Some disturbing signs of endocrine disruption are being observed in wild fish stocks. It is not clear what chemical(s) is(are) responsible for this disruption or whether this endocrine disruption has ecological consequences. However, elucidating the causes and consequences of this disruption must be a top priority for further research. Increased concern has been expressed recently about the relatively high concentrations of some well-known contaminants (i.e. mercury and dioxin) in wild fishes.
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Pregnant women are being advised against eating large quantities of some fishes in some areas due to mercury contamination. As knowledge concerning the effects of low-dosage exposure to certain chemicals increases, the concern about contamination in wild fishes may be likely to increase. There are, then, influences of changes in the environment brought about by human activities on both freshwater and marine fisheries. Responsible fisheries management must recognize these influences on fish stocks, in addition to considering the direct effects of harvest on stocks.
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progress of stages. Water Science Technology 23, 487–496. Kudhongania, A.W., Twongo, T. and OgutuOhwayo, R. (1992) Impact of the Nile perch on the fisheries of Lake Victoria and Kyoga. Hydrobiologia 232, 1–10. Law, R.J., Kelly, C.A. and Nicholson, M.D. (1999) Polycyclic aromatic hydrocarbons (PAH) in shellfish affected by the Sea Empress oil spill in Wales in 1996. Polycyclic Aromatic Compounds 17, 229–239. Lee, R.F. (1991) Metabolism of tributyltin by marine animals and possible linkages to effects. Marine Environmental Research 32, 29–35. Lipton, D.W. and Strand, I.E. (1997) Economic effects of pollution in fish habitats. Transactions of the American Fisheries Society 126, 514–518. Livingston, H.D. and Povinec, P.P. (2000) Anthropogenic marine radioactivity. Ocean and Coastal Management 43, 689–712. Lye, C.M. (2000) Impact of oestrogenic substances from oil production at sea. Toxicology Letters 112–113, 265–272. Maitland, P.S. (1995) The conservation of freshwater fish: past and present experience. Biological Conservation 72, 259–270. Malins, D.C., McCain, B.B., Brown, D.W., Chan, S.L., Myers, M.S., Landahl, J.T., Prohaska, P.G., Friedman, A.J., Rhodes, L.D., Burrows, D.G., Gronlund, W.D. and Hodgins, H.O. (1984) Chemical pollutants in sediments and diseases of bottom-dwelling fish in Puget Sound, Washington. Environmental Science Technology 18, 705–713. Malins, D.C., Krahn, M.M., Brown, D.W., Rhodes, A.J., Myers, M.S., McCain, B.B. and Chan, S.L. (1985) Toxic chemicals in marine sediment and biota from Mukilteo, Washington: relationships with hepatic neoplasms and other hepatic lesions in English sole (Parophrys vetulus). Journal of the National Cancer Institute 74, 487–494. Manning, C.S., Lytle, T.F., Walker, W.W. and Lytle, J.S. (1999) Life-cycle toxicity of bis(tributyltin) oxide to the Sheepshead minnow (Cyprinodon variegatus). Archives of Environmental Contamination and Toxicology 37, 258–266. Mathews, C.P., Kedidi, S., Fita, N.I., Al-Yahya, A. and Al-Rashid, K. (1993) Preliminary assessment of the effects of the 1991 Gulf War on Saudi Arabian prawn stocks. Marine Pollution Bulletin 27, 251–271. Murchelano, R.A. and Wolke, R.E. (1991) Neoplasms and non-neoplastic liver lesions in winter flounder, Pseudopleuronectes americanus, from Boston harbour, Massachusetts. Environmental Health Perspectives 90, 17–26.
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Murphy, M.L., Heintz, R.A., Short, J.W., Larsen, M.L. and Rice, S.D. (1999) Recovery of pink salmon spawning areas after the Exxon Valdez oil spill. Transactions of the American Fisheries Society 128, 909–918. Myers, M.S., Stehr, C.M., Olson, O.P., Johnson, L.L., McCain, B.B., Chan, S.L. and Varanasi, U. (1994) Relationships between toxipathic hepatic lesions and exposure to chemical contaminants in English sole (Pleuronectes vetulus), starry flounder (Platichthys stellatus), and white croaker (Genyonemus lineatus) from selected marine sites on the Pacific coast, USA. Environmental Health Perspectives 102, 200–215. Nielsen, E. and Richardson, K. (1996) Can changes in the fisheries yield in the Kattegat (1953–1992) be linked to changes in primary production? ICES Journal of Marine Science 53, 988–994. Ogutu-Ohwayo, R. (1990) The decline of the native fishes of Lake Victoria and Kyoga (East Africa) and the impact of introduced species, especially the Nile perch, Lates niloticus, and the Nile tilapia, Oreochromis niloticus. Environmental Biology of Fishes 27, 81–96. Pinkney, A.E., Matteson, L.L. and Wright, D.A. (1990) Effects of tributyltin on survival, growth, morphometry and RNA–DNA ratio of larval striped bass, Morone saxatilis. Archives of Environmental Contamination and Toxicology 19, 235–240. Ramamurthy, V.D. (1991) Effects of oil pollution on bio-ecology and fisheries on certain enclosed coastal regions of the Arabian Sea. Marine Pollution Bulletin 23, 239–245. Rashed, M.N. (2001) Cadmium and lead levels in fish (Tilapia nilotica) tissues as biological indicator for lake water pollution. Environmental Monitoring and Assessment 68, 75–89. Reed, M., Spaulding, M.L., Lorda, H.W. and Saila, S.B. (1984) Oil spill fishery impact assessment modelling: the fisheries recruitment problem. Estuarine Coastal and Shelf Science 19, 591–610. Richardson, K. (1997) Harmful or exceptional phytoplankton blooms in the marine ecosystem. Advances in Marine Biology 31, 301–385. Reyes, J.G.G., Jasso, A.M. and Lizarrago, C.V. (1996) Toxic effects of organochlorine pesticides on Penaeus vannamei shrimps in Sinaloa, Mexico. Chemospere 33(3), 567–575. Semenov, A.D., Sapozhnikova, E.V. and Gribanova, S.E. (1998) Pesticide pollution and its role in the reduction of fish stock in the Sea of Azov. Russian Journal of Ecology 29(6), 435–437. Shalaby, A.A., El-Tantawy, M.A. and El-Kenawy, D.A. (1995a) Toxicity of four pesticides on
freshwater fish, the Nile Tilapia and the common carp, with special references to their residues in water and tissues. Zagazig. Journal of Agricultural Science 22(6), 1521–1534. Shalaby, A.A., El-Tantawy, M.A. and El-Kenawy, D.A. (1995b) The effect of some pesticides on transaminases and phosphatases of Oreochromis niloticus and Cyprinus carpio. Zagazig. Journal of Agricultural Science 22(6), 1535–1550. Shilts, W.W. and Coker, W.B. (1995) Mercury anomalies in lake water and in commercially harvested fish, Kaminak Lake area, district of Keewatin, Canada. Water, Air, and Soil Pollution 80, 881–884. Simons, M. (2000) Europe is told it may not be safe to eat fish, either. The New York Times Sunday, 17 December. Spaulding, M.L., Reed, M., Anderson, E., Isaji, T., Swanson, J.C., Saila, S.B., Lorda, E. and Walker, H. (1985) Oil spill fishery impact assessment model: sensitivity to spill location and timing. Estuarine Coastal and Shelf Science, 20, 41–53. Squire, J.L. (1992) Effects of the Santa Barbara, Calif., oil spill on the apparent abundance of pelagic fishery resources. Marine Fisheries Review, 54(1), 7–14. Stagg, R., Robinson, C., McIntosh, A.M., Moffat, C.F. and Bruno, D.W. (1998) The effect of the ‘Braer’ oil spill, Shetland Islands, Scotland, on P4501A in farmed Atlantic salmon (Salmo salar) and the common dab (Limanda limanda). Marine Environmental Research, 46(1–5), 301–306. Steffen, W. and Tyson, P. (2001) Global Change and the Earth System: a Planet under Pressure. International Geosphere Biosphere Programme, IGBP Science 4. IGBP Secretariat. Royal Swedish Academy of Sciences, Stockholm. Stäb, J.A., Traas, T.P., Stroomberg, G., van Kesteren, J., Leonards, P., van Hattum, B. and Brinkman, U.A.Th. (1996) Determination of organotin compounds in the foodweb of a shallow freshwater lake in the Netherlands. Archives of Environmental Contamination and Toxicology, 31, 319–328. Sægrov, H., Hindar, K., Kålas, S. and Lura, H. (1997) Interactions between salmon culture and wild stocks of Atlantic salmon: the scientific and management issues. ICES Journal of Marine Research 54, 1166–1172. Traas, T.P., Stäb, J.A., Roel, P., Kramer, G., Cofino, W.P. and Aldenberg, T. (1996) Modelling and risk assessment of tributyltin accumulation in the food web of a shallow freshwater lake. Environmental Science and Technology 30, 1227–1237.
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Travis, J. (1993) Invader threatens Black, Azov Seas. Science 262, 1366–1367. Vethaak, A.D. and ap Rheinallt, T. (1992) Fish disease as a monitor for marine pollution: the case of the North Sea. Reviews in Fish Biology and Fisheries 2, 1–32.
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Vethaak, A.D. and Jol, J.G. (1996) Diseases of flounder (Platichthys flesus) in Dutch coastal and estuarine waters, with particular reference to environmental stress factors, Part 1, Epizootiology of gross lesions. Diseases of Aquatic Organisms 26, 81–97.
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The Performance of Fisheries Management Systems and the Ecosystem Challenge Jon G. Sutinen and Mark Soboil
Department of Environmental and Natural Resource Economics, University of Rhode Island, Kingston, USA
Abstract This Chapter has three objectives. First, it presents a modest update of the evidence used in the 1997 study by the Organisation for Economic Co-operation and Development (OECD) that attempted to show which management measures are effective in conserving marine fisheries and producing significant economic and social benefits. In its original report, the OECD found that competitive total allowable catch (TAC) management results in a race-to-fish, with all its attendant effects; and that individual fishing quotas are an effective means of controlling exploitation, of mitigating the race-to-fish, of generating resource rent and increased profits, and of reducing the number of participants in a fishery. In addition, the OECD evidence indicated that time and area closures have not been effective in ensuring resource conservation, though conservation might well have been poorer without them. The update indicates that most of the original results are upheld. The second objective is to report on recent trends in policy since 1995, with a focus on ecosystem-based management policies. These include large marine ecosystem (LME) programmes, marine protected areas (MPAs) and the development of alternative rights-based regimes. Thirdly, the chapter examines the governance challenges of ecosystem-based fisheries management. We argue that the political marketplace that produces fisheries management policies tends to be biased against conservation and long-term economic benefits. The chapter concludes with recommendations for reforming our fishery governance institutions.
Introduction In this chapter, we investigate the prospects for ecosystem-based management of fisheries. Specifically, we attempt to answer the question: ‘What features of present fisheries management systems are useful for attaining a broader set of conservation objectives than currently are being used?’
We proceed by first reviewing the evidence of the experiences of the Organisation for Economic Co-operation and Development (OECD) member countries with management regimes that have not been ecosystem based. This includes a modest update of the evidence used in the study by the OECD (1997) that produced evidence on which management measures are effective in conserving marine
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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fisheries and producing significant economic and social benefits. Secondly, we summarize a few recent trends in fishery management policy that were revealed by our update of the evidence. Thirdly, we examine some of the governance challenges of ecosystem-based fisheries management. The chapter concludes with recommendations for reforming fishery governance institutions.
The OECD Results In 1997, the OECD published a study of fisheries management experiences indicating which management measures were effective in conserving marine fisheries and producing significant economic and social benefits. The data for the study were drawn from over 100 fisheries in 24 OECD Member Countries. This is the only study we know that systematically compares individual fishing quotas (IFQs) with more conventional approaches to fisheries management. Since the original OECD study assembled information up to about 1994, we attempted to update the experiences with fishery management in OECD member countries for the period 1995–2000.1 The update is not comprehensive. In other words, we have not been able to obtain evidence on all of the fisheries included in the original OECD report. In addition, we were not able to obtain evidence on the economic, social and administrative outcomes for many fisheries in the update. OECD (1997) analysed three categories of fisheries management measures: output controls, input controls and technical measures. Output controls include total allowable catch (TAC) (total quotas), IFQs and vessel catch limits.2 Input controls include limited
licences, individual effort quotas, and gear and vessel restrictions. Technical measures included size and sex selectivity measures that restrict the size and sex of fish that can be taken and landed, and time and area closures limiting the time and place where fishing units can operate. The analytical framework used in the study assumes that regulations imposed on fisheries affect the performance of the fisheries. The OECD measured this performance in terms of biological, economic and social outcomes. Management measures, such as quotas, closed areas and seasons, and gear restrictions, tend to change the way fishing activities are conducted and, in turn, affect outcomes (stock sizes, landings, incomes, etc.) in the fishery. Actual outcomes, of course, are determined not only by the set of measures imposed but also by the biological, economic, social and institutional characteristics of the fishery (and by influences exogenous to the fishery). To update the evidence reported in OECD (1997), we applied the methodology described in Sutinen (1999) for a selected set of fisheries. In the original study, the performance of specific management measures imposed by the respective institutions was assessed through a step-wise process. The first step involved developing a set of expected consequences, the second step confronted these expectations with evidence, and the third step assessed the theory on which the expected consequences were based.
Individual Fishing Quotas OECD (1997) presented persuasive evidence that shows that IFQs are an effective means of
1 Similarly to the original study, we chose to update fisheries for which a reasonable quantity of good quality information was readily available. The principle sources of information were the Internet, and national fishery web sites and their respective annual stock assessment reports. We collected other information from the OECD Review of Fisheries Reports (1998–2000). Similarly, the majority of the evidence described the biological changes with respect to using different management measures. Evidence to measure the economic, social and administrative performance of a particular management measure was lacking. 2 Vessel catch limits differ from IFQs in that they restrict the amount a fishing unit can catch per trip or short period (such as a week) but do not necessarily restrict the number of fishing units or number of fishing trips or periods.
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controlling exploitation, of mitigating the race-to-fish and most of its attendant effects,3 of generating resource rent and increased profits, and of reducing the number of participants in a fishery.4 IFQs have been effective in limiting catch at or below the TAC level determined by management authorities. OECD reports that catch was maintained at or below the TAC in 23 out of the 31 IFQ fisheries for which information was available. The TAC over-runs that did occur were due to inadequate monitoring and enforcement. Where overexploitation occurred, it was due to poor data that allowed the TAC to be set too high. The OECD evidence demonstrates that IFQs alleviate the race-to-fish and the resulting problems of overcapacity, excess effort, waste, unsafe harvesting practices, gear conflict and loss, and reduced product quality. Two of the most notable cases are the Canadian halibut and sablefish fisheries. Seasons that had been reduced to a few days under competitive TACs and limited entry were increased to most of the year almost immediately. Elimination of the race-to-fish has not been universal, however. For example, in The Netherlands sole and plaice and the Norwegian cod fisheries, IFQs failed to eliminate the race-to-fish.5 The race-to-fish in these fisheries occurred because the fishery could be closed down when the national quota was met, even if individual fishing quotas had not been filled. In Iceland, the option to choose between individual effort and catch quotas in the demersal fishery led to an increase in investment. A race-to-fish occurs in the
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New Zealand flatfish fishery in years of low abundance. Most of the fisheries experiencing a race-to-fish used time or area closures independent of the attainment of TAC, which may have influenced these outcomes. There is a worldwide trend towards the use of IFQs. A growing number of governments are bringing their fisheries under this form of rights-based management. In our update of the OECD evidence, we found that managers have introduced individual transferable quotas in a number of fisheries during the last few years. This is especially evident in fisheries that formerly were under competitive TAC management. In Australia, managers introduced IFQs to the remainder of the trawl quota in the Southeast Trawl fishery, and introduced IFQ management for school and gummy shark stocks in the Southern Shark fishery. A few countries, including New Zealand and Poland, introduced legislation that provided a mechanism either to introduce an IFQ system or to move most of the remaining commercial fish species into a quota managed system. Iceland introduced IFQs for Atlantic wolffish and witch, and Canada introduced them to their Pacific herring large-seine fleet. Denmark introduced the experimental use of annual vessel quotas in the Baltic cod fishery, and for their herring and mackerel fisheries in the North Sea. Norway introduced a unitquota system to their purse seine fleet and cod trawl fleet. In addition, a company-wide quota system was introduced to the groundfish fishery. In Chile, IFQs have been applied to the red shrimp and cod fisheries, but not to the more heavily exploited pelagic fisheries.
3 The race-to-fish causes fishing seasons to be shorter than optimal for maximum economic performance, and causes fish to be caught and landed that are too small and/or of inferior quality. Producers are induced to make excessive investments in their vessels and gear. The excessively large quantities of fish that are landed in short periods also induces the build up of excessively large processing, storage and distribution facilities to handle the periodic peak loads. Sellers and consumers find supplies of fish that are abundant for short periods and scarce for long periods; or the product is processed for a long shelf life, which generally reduces the quality and price of the products on the market. The race-to-fish also aggravates the incidence of unsafe harvesting practices, and gear conflict and loss. 4 The report by the National Research Council (1999) drew upon much of the evidence contained in OECD (2000). 5 However, the recent evidence from 1998 and 1999 now shows that the number of fishermen employed fell in the Norwegian cod trawler fleet, and that overall profitability had increased substantially.
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We argue that fishery managers are introducing IFQs because they work well. IFQs have a proven record of accomplishment of promoting sustainable management of fisheries and producing wealth. In our view, it is easier to explain why IFQs are used than why they are not used in more fisheries. This is considered further below. The OECD study also demonstrates that IFQs present problems with the initial allocation of quota and with enforcement and compliance. Of the 55 IFQ fisheries reviewed by OECD, quota allocation problems were documented in ten fisheries, with no counter examples. The initial allocation of quota is the major impediment to the adoption of IFQs in most fisheries. Potential participants in an IFQ programme commonly express concern that they will not receive their fair share in the initial allocation of quota. The exceptions are fisheries with a relatively small number of producers who are relatively homogeneous. The struggle to find a fair and just allocation of harvest rights is difficult, time-consuming and adversarial. The allocation struggle also extends to the processing sector. For example, there is a debate over whether processors in Alaska should be awarded rights to processing shares in any future IFQ fisheries. Allocation of the access to fish, or of the rights to catch fish, is a problem that plagues all forms of fisheries management, whether based on IFQs or traditional methods. Allocation is the constant topic of meetings and decisions made by fishery managers, and even the subject of legislative deliberations. There is a trade-off related to allocation and IFQs that the management system too frequently is unwilling to make. The trade-off is between the ongoing cost of allocation problems with conventional management measures and the high up-front cost of initial allocation with IFQs. Without a market to handle allocation issues, the management system pays the price of allocation struggles on a continuing basis. While the initial
allocation of IFQs is extremely difficult, the ‘pain’ is all up front and once-and-for-all. This is especially true for transferable IFQs, since thereafter a market emerges to handle the re-allocation of a quota that is needed for the fishery to evolve.6 If managers escape the high up-front cost of the initial allocation of transferable IFQs, they must face the continuing distraction of dealing with allocation instead of conservation. Actual solutions to the initial allocation problem have taken a variety of forms. This variety is probably because there is not universal agreement on what constitutes a fair and just allocation. Each solution is the result of a negotiation and bargaining process. The important aspect of the solution is the process – the process by which the solution is found. An open and transparent process is needed to ensure institutional legitimacy, credibility and trust. Higher enforcement costs and/or greater enforcement problems occurred in 17 fisheries compared with five that experienced improvements. Enforcement proved particularly difficult in high-value fisheries, in multi-species fisheries and in transborder fisheries. Although enforcement costs frequently increased under individual vessel quotas, there often was an increased ability and willingness of fishermen to pay these increased costs. Support from industry for increased enforcement is common. IFQ holders recognize that illegal fishing by others damages the value of their quota rights and therefore they have an incentive to aid authorities with enforcement. The rents generated by IFQs provide governments with a source of revenue to cover the costs of enforcement and administration. In the many IFQ fisheries in Australia, Canada, Iceland and New Zealand, industry pays for administration and enforcement with fees levied on quota owners. In some cases, quota holders voluntarily paid for added enforcement, such as in the New Zealand lobster fishery. In addition, IFQ management
6 If the IFQs are not transferable, management authorities also will have to re-visit the allocation issue repeatedly.
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has led to increased cooperation between fishermen and enforcement authorities in several cases, including the New Zealand fisheries in general, and the USA wreckfish fishery. Fishermen reported improved compliance in the Canadian halibut fishery.7 We have learned a great deal over the last 20 years of IFQ management. Despite the many and serious problems that have confronted IFQs, fishery managers are finding ways to mitigate or remove these impediments to IFQs. We see evidence that managers can find designs of IFQ programmes that satisfy first principles (such as creating an exclusive harvest right) and still address the concerns of fairness and justice. Where no solutions are immediately evident, countries should craft the legislation to encourage innovation and experimentation.
How do IFQs compare with other fishery management measures?8 In its assessment of other management measures, OECD (1997) concluded that none of the other (non-IFQ) management measures performs well when used without IFQs, i.e. they do not effectively control exploitation and mitigate the race-to-fish. However, they do not present as many social and administrative difficulties as IFQs. The findings are summarized below.
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Total allowable catch quota OECD (1997) found that competitive TAC management causes a race-to-fish, with the attendant effects of overcapitalization, shortened seasons, market gluts, increased harvesting and processing costs particularly evident. Competitive TAC management generally has not prevented overexploitation of the fishery resource effectively – though it has been successful in some fisheries. The update of selected experiences supports the original findings that competitive TAC management results in a race-to-fish with all its attendant effects, and generally has not prevented overexploitation of resources effectively. For example, in the EU, recent stock assessment reports indicate that most of the fishery stocks that are managed with competitive TACs are heavily fished. Management authorities have reduced the TACs substantially for a number of stocks over the last couple of years. Since 1995, TACs have been introduced to several fisheries not mentioned in the original OECD report. In the EU, TACs were introduced to numerous species in the North Sea fisheries, as well as for tuna and swordfish. Japan also subjected six of its commercially fished species to a TAC system, complementing the effort and limited access systems and technical measures already in place. There were also some modifications to a couple of fisheries managed with competitive TACs. For example, the Canadian Pacific
7 Other problems with IFQs that were identified included: under-reporting of catch and data degradation (documented for 12 fisheries, but improvements were made in six fisheries); industry resistance to IFQs in eight fisheries, but the opposite was true in five fisheries; several cases where quotas were consolidated (documented in 12 fisheries, but five showed contrary evidence), and rules were in place to limit consolidation; little evidence that smaller vessels are eliminated when individual vessel quotas are introduced (two fisheries where elimination occurred and five where it did not); divisions among social classes were documented only for the Icelandic fisheries. 8 The OECD study represents one of the few attempts, if not the only one, to assess comprehensively the performance of the full suite of management measures. The study found considerable evidence, and excellent scholarly studies of IFQs, limited licences and TAC measures. However, there is great paucity of evidence on the performance of the other management measures (size and sex selectivity, closures, effort quotas, vessel catch limits and gear and vessel restrictions). While the theory of how these measures are supposed to work is well developed, the supporting empirical evidence is missing. The actual application of these methods appears to be conducted more on faith than on a sound factual basis.
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herring fishery continued to use competitive TAC management, but separate TACs were established for the spring and autumn components in the inshore fleet.
Limited licences OECD (1997) found that overcapitalization and increased harvesting costs occur with limited licences, but the evidence is confounded by the presence of TACs in many of the reported cases. There have been some initial allocation problems, but the amount of evidence is too small to draw a firm conclusion. Limited licences have not stemmed the tendency to overexploit the fishery resource.
Size and sex selectivity OECD (1997) found that size and sex selectivity measures do not mitigate the race-to-fish and result in increased enforcement costs or other problems, as demonstrated by the evidence. There is only weak evidence that the average size of fish landed increases and that discards increase. Our update uncovered some evidence that is contrary to the 1997 report in this respect. Management authorities have changed size and sex selectivity measures, including increasing the minimum size regulation in the Victorian abalone and scallop fisheries in Australia. Recent assessments have shown that, in the Victorian scallop fishery, average production had in fact fallen and become sporadic.
Closures OECD (1997) concluded that that time and area closures have not been effective in assuring resource conservation, though conservation might well have been poorer without them. Our update revealed that there have been several time and area closures introduced since 1995. Seasonal closures in the Great Australian Bight trawl fishery were implemented within the marine mammal protection area of the Great Australian Bight marine park; and demersal trawling was prohibited within the benthic protection strip area.
Impacts on the fishery by the marine park are still uncertain. Recent assessments show that there is some variability within the fishery, although it appears to be due to changes in the aggregating behaviour of the fishes. In addition, in the EU, seasonal closures were introduced for the first time to the Mediterranean bluefin tuna fishery. Area and season lengths have also increased in several fisheries, including the Tasmanian abalone fishery and the Icelandic capelin fishery. In the USA, three large closed areas on Georges Bank and in the Gulf of Maine, initially implemented to improve the recruitment of groundfish, have aided in the recovery of yellowtail flounder and dramatically improved the abundance of large scallops. Scallop biomass tripled in the first 20 months after closure of the areas on Georges Bank.
Individual effort quotas OECD (1997) found that individual effort quotas (e.g. days-at-sea and trap quotas) result in overcapitalization, increased harvesting costs and increased enforcement problems.
Vessel catch limits OECD (1997) found that vessel catch limits (as distinguished from IFQs) increase enforcement costs and problems. We found in our recent survey that none of the fisheries that continued to use vessel catch limits improved resource conservation. Average production in the Victorian scallop fishery, for example, had fallen and, in 1996, the Australian Victorian Government announced that it would buy-out the licences issued in Port Phillip Bay. The Canadian Pacific groundfish trawl fishery declined further, and currently is at historical low levels.
Why do IFQs perform so well? IFQs provide important benefits that other approaches do not. IFQs effectively constrain exploitation within set limits, mitigate the race-to-fish, and reduce overcapacity and
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gear conflicts, while improving product quality and availability. Producers benefit, consumers benefit and, when the resource rent is used to pay for the cost of management, the general public benefits. In addition, there are environmental benefits that often are overlooked. For example, reducing the 300,000 traps in Area 2 of the North American lobster fishery is expected to reduce entanglements with whales substantially, while at the same time realizing the same yield. Why do IFQs and other rights-based approaches have the potential to achieve this much? Fishery economists and most social scientists are not surprised that IFQs perform so well in comparison with other management measures. IFQs solve numerous problems by providing exclusive harvesting rights. Other ‘rights-based’ management measures have the potential to do the same. None of the traditional management measures provides exclusive rights and, therefore, cannot solve the problems created by non-exclusive use of the resource. In fisheries without exclusive harvesting rights, no fisherman has the right to exclude other fishermen from harvesting any part of the resource. From an individual fisherman’s perspective, leaving fish to grow and reproduce is done at the risk of losing the fish to other fishermen. There is therefore no incentive to conserve the resource for future use, since no fisherman has exclusive use. The non-exclusive nature of fisheries resources is the fundamental cause of overexploitation in modern fisheries. Without an exclusive right to harvest a quantity of fish, competition to catch fish before others do then causes ‘race-to-fish’, resulting in fishing seasons that are shorter than optimal for maximum economic performance, landings of inferior quality, and excessive investments in vessels and gear. The non-exclusive nature of harvesting fisheries resources also leads to conflicts among user groups. Since no fisherman has the right to exclude another from access to the resource, two or more fishermen can interact at the same time and place in a fishery. They impose external costs on each other in the form of gear or other losses. Mobile gear (such as trawls) may fish in the same area as fixed
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bottom gear (such as traps), causing damage to one or both of the gears. Large, efficient vessels can operate in a fishery on which small-scale fishermen are heavily dependent, draining the stock available for capture by the smaller fishermen. Failure to consider these external costs when deciding where and how to fish causes inferior economic performance in the fishery. Processors, distributors, wholesalers, retailers and consumers are also affected by the non-exclusive nature of harvesting. The raceto-fish can result in large quantities of fish being landed during short periods, requiring the building up of excessively large processing, storage and distribution facilities to handle the periodic peak loads. Wholesalers, retailers and consumers find supplies of specific fish are abundant for short periods and scarce for long periods; or, the product is processed for long shelf life, generally reducing the product quality and price on the market. Several attributes appear to contribute to the success of IFQ management. Fisheries with a limited number and well-defined group of participants are more easily brought under and managed with IFQs. Fisheries that were under limited entry or had a very small number of participants provide a welldefined user group, and initial allocations are made easier. The quota holders have already developed a sense of property in the fishery in many cases that contributes to acceptance of individual fishing quotas and, at least in some cases, may contribute to improved compliance and collaboration with enforcement. Fisheries with a homogeneous fleet are more easily put under IFQs. Allocations are decided more easily and the fishery typically has less adjustment to go through. OECD (1997) reported that user participation in the development and implementation of fishery management plans is a critical element for successful management. Co-management arrangements are one of the more promising avenues for greater user participation. Yet there are several outstanding questions concerning co-management. How, for example, should co-management be implemented? What rights and responsibilities should remain with government and which with users? Is more user participation
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better than less in all fisheries? Is user participation feasible and desirable in the cases of straddling stocks and highly migratory species? If producers’ organizations are needed for co-management, how can and should fishermen be organized, especially when they have no history of organization? Canada, Denmark, The Netherlands, Norway, Sweden, the UK and other countries have devolved fishing rights and responsibilities to producers and their organizations. These countries have found that the comanagement approach reduces administrative costs and greatly improves compliance with management regulations. Japan has built on its lengthy tradition of rights-based management and now has the world’s most extensive and sophisticated fisheries co-management system (OECD, 1997). Other countries have much to learn from these experiences. Of all the management measures available to managers, rights-based management measures (such as IFQs) seem to have the greatest chance of correcting the fundamental problem of non-exclusive harvesting rights and of reducing conflicts among users, producing superior economic performance while conserving fishery resources.
Multi-species fisheries and ecosystem-based management In theory, most management measures are expected to provide some degree of conservation benefits in the form of maintaining or rebuilding resource stocks to desired levels. Unfortunately, in practice, no known management measure assures optimal resource conservation. Achieving optimal conservation is complicated by several factors or conditions, including multi-species, bycatch and discards, and wide fluctuations in resource stocks and markets. Despite the complex challenges presented by multi-species fisheries, the OECD evidence shows that IFQs outperformed all other management measures. This is not to say, however, that only IFQs are needed in multi-species fisheries. Rather, when other management measures (such as mesh size
regulations) are used in combination with IFQs, performance was superior. When not used with IFQs, performance suffered. Fisheries that harvest multiple species are more difficult and costly to manage effectively than single-species fisheries. A high proportion of multi-species groundfish fisheries in OECD countries experienced poor resource conservation and economic performance. Relatively non-selective trawl gear is used in these fisheries, which results in high by-catch and discard rates, further weakening management’s control on exploitation patterns (unless by-catch and discarded catch are monitored adequately). Multi-species fisheries complicate all forms of fishery management. In multispecies fisheries where several species are caught jointly, it is very difficult to achieve the optimal fishing mortality for all species. Almost any change in management measures will favour one species at the expense of another. Good conservation on all stocks simultaneously may not be feasible in such cases. There is widespread consensus on the importance of accounting for multi-species interactions in fisheries analysis and management, but only a limited amount has been accomplished to date. The theory for developing models to explain and analyse interactions is well developed. Biological and economic empirical evidence, however, is inadequate. Attempts to model multi-species fisheries in several countries are ongoing and are already providing information for the management process in some fisheries. IFQs seem to offer high promise, relative to non-rights-based approaches, for wrestling with the challenge of managing complex marine ecosystems (Arnason, 2000). Researchers currently are exploring other rights-based approaches, but no experiments or tests of these approaches are underway. By-catch is inevitable in many multispecies fisheries. Incentives play a major role in determining the amounts of by-catch. An individual fisherman will try to control by-catch as long as the benefits outweigh the costs to him. Effective management recognizes this and creates or modifies incentives to lessen the impact of by-catch.
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There is some anecdotal evidence suggesting that substantial discarding at sea and under-reporting of landings have increased since the implementation of IFQs. However, a study done for OECD found no discernible increase in discards under an IFQ system compared with the previous limited effort management scheme. In our update, we found no information either to support or to refute any increase in by-catch under IFQs. However, the development of by-catch action plans was introduced in a few IFQ fisheries, including the South East trawl fishery and the Torres Strait prawn fishery. In Canada, individual vessel by-catch limits were implemented to supplement the IFQ-managed groundfish trawl fishery, resulting in substantial by-catch mortality reductions. Some countries have developed tools to counteract discarding. These tools include setting TACs by species such that different TACs can be filled approximately simultaneously while employing standard harvesting technologies; simple and well advertised discard rules; flexible monitoring and surveillance designed to deal with the most pressing problems at each point in time; and addressing alleged violations quickly and effectively with penalties high enough to deter such practices.
Selected Trends in Fisheries Management Policy Since 1995 Our review of recent fisheries management experiences also revealed some trends in management policy that were not evident in the original OECD study. This section reports on some of these trends in policy since 1995, with a focus on ecosystem-based management policies. These include closures and marine protected areas (MPAs), large marine ecosystem (LME) programmes and the development of alternative rights-based regimes.
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Closures and marine protected areas9 There is a growing demand for the establishment of various forms of MPAs the world over. Management authorities are implementing MPAs to protect, maintain or restore natural and cultural resources in coastal and marine waters. They are also using MPAs to conserve biodiversity, manage natural resources, protect endangered species, reduce user conflicts, provide educational and research opportunities, and enhance commercial and recreational activities (Salm et al., 2000). The World Commission on Protected Areas (of the International Union for the Conservation of Nature and Natural Resources – IUCN) is collaborating with the Global Environmental Facility (GEF) and its partners to design projects highlighting innovative approaches to marine biodiversity conservation and community development. Projects are being implemented in Samoa, Tanzania and Viet Nam that will employ communitybased management of marine biodiversity and sustainable use of marine resources. In Australia, MPAs are emerging as an important plank in a number of government initiatives to promote sustainable fishery management. For example, their Oceans Policy identified principles for the ecologically sustainable use of their fishery resources. Moreover, it provides an infrastructure to examine the need for, and options to achieve, integrated oceans management, and ensure the long-term protection of their marine environment. In April 1998, Australia’s Commonwealth government proclaimed its second largest marine park, covering 2.3 million ha and located in the Great Australian Bight. It provides protection for the southern right whale, the Australian sea lion and a large range of benthic flora and fauna species. It comprises an area designated for mammal protection adjacent to a State inshore park. A second area, 20 nautical miles wide, and
9 For the purposes of this discussion, we use the definition of MPAs that was developed by the World Conservation Union: ‘any area of the intertidal or subtidal terrain, together with its overlying water and associated flora, fauna, historical and cultural features, which has been reserved by law or other effective means to protect part or all of the enclosed environment’ (IUCN, 1988).
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extending from the State park boundary to the limit of the Australian exclusive economic zone (EEZ), is designated to conserve benthic flora and fauna. In 1999, the first deep-sea seamount Marine Reserve was declared off the southern coast of Tasmania. This MPA consists of two vertical zones. The upper zone from the surface to 500 m is a managed resource zone. The zone deeper than 500 m is a no-take zone. Later in the year, the waters around the southeastern region of Macquarie Island were also declared an MPA. In addition, in Canada, the Oceans Act provided a management system based on the principles of ecosystem-based management. It provided a framework for establishing MPAs, which, according to the Oceans Act, would include being used for the conservation and sustainable management of their fishery resources. In 1998, Canada established its first pilot MPAs at Race Rocks and Gabriola Passage. The same year, a pilot MPA was also announced for the Sable Gully on the edge of the Scotian Shelf, just north of Sable Island. It provides habitat for a wide diversity of marine life, including 200 bottlenose whales, a vulnerable species that lives in the Gully year round, and some of the best examples of northern coral. Recently, two new pilot MPAs in the offshore waters of the Pacific Ocean were established at Endeavour Hot Vents Area and the Bowie Seamount. Endeavour Hot Vents lie in the offshore waters of the northeast Pacific, about 250 km southwest of Vancouver Island. Endeavour Hot Vents Area is the world’s first pilot MPA for offshore hydrothermal vents. Bowie Seamount is located 180 km west of the Queen Charlotte Islands in the northeast Pacific. It is an ancient subsea volcano, rising over 3000 m above the sea bottom. In 2000, President Clinton signed an Executive Order regarding MPAs in the USA. The Executive Order would help protect the significant natural and cultural resources within the marine environment for the benefit of present and future generations by strengthening and expanding the nation’s system of MPAs.
Only a few studies demonstrate the replenishment of fish stocks on fishing grounds via export from closures. However, in fairness, it appears that few closures have been established solely for this purpose. Interestingly though, several international agencies, including IUCN (1994), state that one of the aims of MPAs is to ensure the sustainable utilization of species. There are, however, numerous studies that have shown that the density of some species within closures have increased, the average size of individual of a given species can change, and that fish biomass can increase significantly (NRC, 2001). A few studies demonstrate the potential to replenish exploited fish stocks through the dispersal of larval or adult fish from the closed areas into regions where fishing is allowed. Two studies, one in the Philippines and the other in Kenya, documented increases in the populations of large adult fish in protected areas and subsequent population enhancements in adjoining regions (McClanahan and Kaunda-Arara, 1996; Russ and Alcala, 1996). The benefits of protecting juvenile fish for their export as young adults to fishing grounds has also only been documented vaguely. Single-species closures used for plaice in the North Sea and mackerel in southwest England have resulted in increased yields by enhancing juvenile survival (Horwood et al., 1998). On Georges Bank and in the Gulf of Maine, more than 5000 square miles were closed to bottom trawling and dredging in December 1994 in response to the critical decline of groundfish stocks. The intent was to improve recruitment of groundfish by reducing by-catch of juveniles and preventing the disturbance of juvenile habitat in the closed areas. Evidence has shown that the three large closed areas have aided in the recovery of yellowtail flounder and dramatically improved the abundance of large scallops. Scallop biomass tripled in the first 20 months after closure of the areas on Georges Bank (NRC, 2001). A side effect of the closure, however, was the displacement of fishing effort that resulted in the overfishing of other spawning components.
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Large marine ecosystems The World Bank and the GEF have adopted the LME approach to marine ecosystem research and management, viewing it as ‘an effective way to manage and organize scientific research on natural processes occurring within marine ecosystems [and] to study how pollutants travel within these marine systems . . .’ (World Bank, 1995: Annex A). There currently are 11 active LME projects, funded at US$2750 million, involving 62 countries. The concept of LMEs is a science-based method for dividing the world’s oceans, developed 15 years ago by Kenneth Sherman and Lewis Alexander (1986). LMEs are geographic areas of oceans that have distinct bathymetry, hydrography, productivity and trophically dependent populations. The geographic limits of most LMEs are defined by the extent of continental margins and the seaward extent of coastal currents. Among the 52 LMEs identified to date are the Benguela Current, Gulf of Guinea, Bay of Bengal, Yellow Sea, Northeast USA Continental Shelf, Gulf of Alaska, Gulf of Mexico and Eastern Bering Sea, to name a few. Some LMEs are semi-enclosed seas, such as the Caribbean, Mediterranean and Black Seas. LMEs can be divided further into subsystems such as the Gulf of Maine, Georges Bank, Southern New England and the Mid-Atlantic Bight in the case of the Northeast USA Continental Shelf (Sherman et al., 1988). Approximately 95% of all fish and other living marine resources produced are taken from the world’s 51 LMEs. Unfortunately, many LMEs currently are stressed from overexploitation of marine resources, habitat degradation and pollution (Sutinen, 2000). The LME management approach links the management of drainage basins and coastal areas with continental shelves and dominant coastal currents. The approach: (i) addresses the many-faceted problem of sustainable development of marine resources; (ii) provides a framework for research monitoring, assessment and modelling to allow prediction and better management decisions; and (iii) aids in focusing marine assessments and management on sustaining
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productivity and conserving the integrity of ecosystems. The assessment, monitoring and governance challenges of LMEs are enormous. The report by Sutinen (2000) presents a methodology for determining what is known of the socio-economic and governance aspects – the human dimensions – of LME management. The report describes a basic framework for identifying the salient socio-economic and governance elements and processes of an LME. Methods for monitoring and assessing the various elements and processes are also discussed. LME management increases the need for intergovernmental and intersectoral management. Government agencies will have to identify barriers to interagency coordination and develop alliances and partnerships with non-federal agencies and private sector stakeholders (Hennessey, 1997). Management agencies must learn to cope with the uncertainty associated with the complexity of ecosystems as natural systems, and the organizational and institutional complexity of the implementation environment (Acheson, 1994; Hennessey, 1997). A major impediment to successful management is imperfect fit between the spatial and temporal scales of government jurisdictions and ecosystems. Ways to connect ‘nested’ ecosystems through ‘networked institutions’ at federal, state, local and nongovernmental organization (NGO) levels will have to be found (Hennessey, 1997). How these institutions adapt to deal with the complexity of the ecosystem and the complexity of the governance system in order to achieve an optimal mix of benefits and costs is a fundamental issue (Creed and McCay, 1996).
Other rights-based management approaches A couple of creative rights-based alternatives to IFQs have been employed in the USA. One of the approaches is Community Development Quotas (CDQs), and the other is fishery cooperatives. Both of these programmes are in Alaska.
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In the USA, the North Pacific Fishery Management Council implemented a CDQ programme in December 1992. The CDQ programme allocates a portion of the annual fish harvest of certain commercial species directly to a coalition of villages in the Bering Sea region. The programme was an attempt to accomplish rural development in rural coastal communities in western Alaska. In its first year, the council allocated 7.5% of the TAC catch of Bering Sea Pollock to six CDQ groups, organized from 56 eligible communities (recently expanded to 57). They managed their harvest quotas and allocated the returns. The quotas are transferable, and thus those fishing partners authorized by the communities in exchange for royalties can also harvest a portion of this TAC. In 1996, an amendment to the Magnusson Act extended the CDQ programme to include halibut, sablefish, crab and assorted groundfish managed under Federal Fish Management plans. The CDQ programme has had a positive economic impact on Western Alaskan communities. During the first 4 years of operation, the six CDQ groups collected over US$92 million in gross revenues from fishing partners. The CDQ programme has enhanced the employment of western Alaskans in the commercial fishing industry, as well as their average income. Evidence suggests that the programme has also contributed to improve understanding of business administration, corporate structure and procedures, and technical skills of village residents (NRC, 1999). Also in the USA, the Pollock Conservation Cooperative (PCC) was formed in December 1998, to promote the rational and orderly harvest of pollock by the catcher/processor sector of the Bering Sea and Aleutian Islands trawl fisheries off Alaska, through the mutual cooperation of PCC members. The PCC is made up of eight companies that own 19 catcher/processors eligible under the American Fisheries Act (AFA) to harvest and process pollock in the directed pollock
fishery. Under the PCC, each company is contractually allocated a percentage of the directed fishery catch specified under the AFA. The cooperatives for the factory trawler sector began in 1999, and the cooperatives for the mother ship and inshore processor sectors began in 2000. For the Bering Sea–Aleutian Island area (BSAI) and the Gulf of Alaska (GOA) fisheries as a whole, the annual discard rate for groundfish decreased from 14.6% in 1995 to 9.4% in 1999, after a very large reduction in 1998 and a small increase in 1999. The 43% reduction in the overall discard rate in 1998 is the result of prohibiting pollock and Pacific cod discards in all BSAI and GOA groundfish fisheries beginning in 1998. The ex-vessel value of the domestic landings in the FMP fisheries, excluding the value added by at-sea processing, decreased from US$585 million in 1995 to US$531 million in 1996, increased in 1997 to US$615 million, decreased again to US$416 million in 1998, and increased to US$488 million in 1999 (Hiatt and Terry, 2000).
The Political Economy of Fisheries Management Despite the promising trends towards more use of rights-based management measures, our fishery management institutions do not have a good record of conserving and managing fisheries. Garcia and de Leiva Morena (Chapter 1, this volume) report that 25% of the world’s fish stocks for which data are available are underexploited, 47% are fully exploited, and the remaining 28% of fish stocks are overexploited. In addition, most of these fisheries suffer from overcapacity,10 product waste, user conflicts, low incomes and other negative economic benefits. In our view, it is clear that our fishery management institutions have failed to
10 Overcapacity is the condition in which a fishing fleet has the ability to produce more than the fishery resource can sustain or more than a desired reference point (e.g. maximum sustainable yield).
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conserve resources and improve the economic health of fishing communities. Why have our management institutions not done better? According to FAO (2001a), the poor record is due to problems that include:
• • • •
uncertainty about the status and dynamics of the stock; a tendency to give priority to short-term social and economic needs at the expense of the longer term sustainability of the stock; poorly defined objectives; and institutional weaknesses, particularly in relation to the absence of long-term rights amongst the different key stakeholders and decision-making structures and processes.
We focus here on the last three problem areas for fisheries management. The first is not necessarily a cause of management failure. After all, managers in the private sector routinely succeed in the face of severe uncertainties. We view the last three problem areas as manifestations of the incentive structure found in most fisheries governance regimes. Unless the incentive structure in our collective decision making organizations is repaired, it will remain biased against the conservation of fisheries resources.
Fig. 17.1.
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Fisheries governance system The fisheries governance system in most countries consists of both formal and informal linkages among four components of the system. The legislature passes fisheries laws that authorize the implementation of fisheries policies and programmes by a fisheries agency. In turn, the fisheries agency commonly establishes a fisheries management authority to develop fishery management plans that specify the set of management measures that are applied to the fisheries under its jurisdiction. Stakeholders (fishing producers, communities and environmental advocates) usually have a formal role – from advising to decision making – in the management plan development process. The resultant plans, if approved, are then implemented by the fisheries agency. The solid arrows in Fig. 17.1 represent these formal linkages. In addition to the formal linkages, there are informal linkages – represented by the dashed arrows in Fig. 17.1. As voters who help elect members of the legislature, fisheries stakeholders frequently take their problems and concerns to their elected representatives. If a stakeholder group feels that the fishery management process has not treated it properly, they will ask their elected representative
Fisheries governance system.
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to assist them. The assistance often is in the form of influencing the fisheries agency and/or fisheries management authority. Sissenwine and Mace (Chapter 21, this volume) refer to this as the ‘end run’ phenomenon. We now examine the political dynamics of the fisheries governance system in more detail.11 The political process controls government and, in most Western democracies, political decisions are produced by a legislative process. Voters elect representatives to guide government policies and actions; agencies are formed; and bureaucrats are hired to implement government policies. These three groups are the major players in the political process, and government policies and actions result from complex interactions among these players. Voters – especially groups of voters with special interests – express their demand for government policy and action. Elected representatives supply legislation (policy) and government bureaucrats implement the programmes and rules specified in the legislation. In the fisheries context, the principle products of this political marketplace are fisheries laws and regulations related to conservation and management, safety, environmental protection, etc. Voters are the consumers of the political process, demanding political products. Voters and groups of voters (that form to pursue their special interests) demand public-sector action to reduce inefficiencies and to redistribute income, usually in a self-interested redistribution. Votes, campaign contributions and lobbying are the currency by which these demands are expressed. Politicians are the elected administrators and legislators in federal, state and local government, including members of executive and legislative branches. Politicians are motivated in part by the need to be elected or remain in power by supplying the political goods that are demanded by voters. Therefore, politicians tend to select
positions that maximize the probability of re-election. Bureaucrats work at national, state or provincial, and local levels as hired officials. Agency employees implement laws and regulations, and develop programmes. Bureaucrats are also motivated in part by self-interest. They naturally resist downsizing of their budget and number of employees, and commonly attempt to increase the size of their budget and number of employees. To achieve this objective, they often appeal to politicians with programmes that would be favoured by voters.12 Political equilibrium is reached as voters, politicians and bureaucrats make choices to achieve their own objectives. Both socially desirable and undesirable outcomes are possible, depending on the underlying incentives of these groups. Governance failure (also known as socially undesirable outcome) is due to a number of inter-related causes, including:
• • • • • •
special interest effects; rational voter ignorance; bundling of issues; shortsightedness effects; de-coupling of costs and benefits; and bureaucratic inefficiencies.
Special interest effects occur when a relatively small number of voters make large individual gains at the expense of a large number of citizens who bear small individual losses. Rent seeking occurs when individuals and groups attempt to use the political process to redistribute income from others to themselves. Special interests gain disproportionate power relative to their numbers because they can provide campaign funds, publicity and delivery of voters who are passionate about a particular issue. Meanwhile, rational voter ignorance occurs because it is seldom worth the cost for the typical voter to acquire the information needed to make a fully informed voting decision. In addition, the choice of a
11 The arguments in this section are based on extensive public choice literature (e.g. Buchanan and Tullock, 1962; Olson, 1964; Niskanen, 1971; Buchanan, 1980; Wolf, 1988). Applications of public choice to fisheries can be found in Andersen et al. (1998) and Sutinen and Upton (2000). 12 For a study of this phenomenon outside of fisheries, see Johnson and Libecap (1994).
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single voter is seldom decisive when the overall number of voters is large. This further decreases the voter’s motivation to acquire more information, while in many cases the individual may not bother to vote at all. These factors induce the politician to favour special interests. The packaging or bundling of the candidate’s positions further accentuates special interest effects and rational voter ignorance. Members of the general public who are relatively disinterested in a specific issue are unlikely to vote on the basis of that issue alone. It is likely that many other issues are of greater importance to him or her, especially when the impact on their welfare is small. Yet members of an interest group are likely to vote strictly according to the issue, especially when it has a significant effect on their welfare. A given political candidacy will be accepted or rejected on the basis of the entire package of positions and not on the basis of a single special interest issue. Since voters can only express their will through a legislator who represents a bundle of political goods, the political process becomes imprecise with regard to voter preferences. For example, it has been estimated that the typical citizen makes only one public choice decision for each thousand made in the private sector. In addition, politicians often package issues in a complex manner so that most voters will be unaware of the true costs that programmes will impose upon them. However, special interests are likely to be well informed regarding the underlying costs and benefits of a policy that is specific to their interests. Politicians tend to be shortsighted because they face short re-election cycles, of 2, 4 or 6 years.13 They are concerned about the consequences of policies and programmes before the next election. The long-term consequences tend to carry little weight in the calculations of the politician. Politicians often exhibit shortsightedness. They regularly enact special legislation and appropriations for fisheries, and periodically attempt to influence directly the contents of fishery management
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plans. Shortsightedness is a natural attribute of a politician. Shortsightedness also is present on the demand side. Fishing interests in most managed fisheries tend to be shortsighted about fishery management policy. In open-access fisheries, fishermen have no secure claim on future outcomes in their fishery, i.e. they have no assurance that they will reap the benefits that might accrue from their short-term sacrifices. Fishermen in rights-based fisheries, on the other hand, are expected to be less shortsighted. Fishermen also tend to be shortsighted because of the great uncertainty they face regarding future fishery policies, fish stocks and markets. Fishermen are simply being rational in their shortsightedness. The shortsightedness on both demand and supply sides combines to favour legislation that provides easily identified current benefits at the expense of future costs that are complex and difficult to identify. Conservation, which requires short-term sacrifice in exchange for long-term gains, tends to be disfavoured in this environment. Another characteristic that strongly influences fishery policies and outcomes is de-coupled benefits and costs. Political products have benefits and they have costs. For many fishery products, those who benefit are not those who pay the cost of a product. For other products, benefits accrue at a different point in time from the costs. An example of de-coupled benefits and costs are government-financed vessel buyback programmes, such as the US$25 million vessel/permit buyout programme in the USA Northeast fisheries. The beneficiaries are the fishermen whose vessels are purchased by the programme and those remaining in the fishery. The costs, on the other hand, are borne by the general taxpayer. The beneficiaries do not pay in proportion to the benefits they receive; and the payers do not benefit in proportion to what they pay. Government agencies do not face incentives to produce goods and services efficiently. By cultivating the political influence
13 We use the term shortsightedness to describe the tendency by people to ignore, or give little weight to, future consequences, especially consequences in the medium to distant future.
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of powerful politicians and groups of constituents, bureaucrats create opportunities for themselves to lead larger government agencies. While bureaucrats compete for tax revenues, promotions, higher incomes and greater power (just as employees do in the private sector), they do not face incentives to increase the value and decrease the costs of their outputs. Public employees cannot increase their income by improving the efficiency of the agency, and their job performance is usually difficult to measure (at least in terms of the contribution to the agency’s output). As a result, they tend to be less conscious of costs, especially since they are spending other people’s money. There is no need to compare revenues with costs; there is no measure of inefficiency and no pressure to reduce it. The incentives inherent in government agencies lead to inefficient production of government goods and services. In addition, government is often the sole provider of the good or service. The exclusive right of production is often mandated by law. Education and postal services in the USA are exceptions. In general, the lack of constant competition for customers leads to inefficiency in government production. Unlike the private sector, there is no systematic mechanism to weed out governmental inefficiencies. In the private sector, inefficient firms do not survive – they go bankrupt. In the public sector, agencies with high costs or that cannot meet their targets are often rewarded with increased funding. Agencies that reduce costs and do not spend their budget allocation are penalized with the threat of a smaller budget the following year. These two characteristics, shortsightedness of the principal actors and de-coupled benefits and costs of fishery products, have a powerful influence on the choice of fishery management policies. The presence of shortsightedness and de-coupled costs and benefits works against adoption of effective conservation policies. The structure of the fishery management system tends to disfavour effective conservation policies because they concentrate short-term costs upon resource users in exchange for benefits in the future that would
not necessarily accrue to those users who make the sacrifice. There are many examples in which the political marketplace favours fishery policies and programmes where benefits are distributed to a few and the costs are borne by taxpayers. Fishery policies and programmes with short-term benefits, and costs to be borne in the future are also favoured. The political marketplace disfavours policies and programmes for which costs are concentrated on a few and benefits accrue to many; and policies and programmes with short-term costs in exchange for future benefits. The fishery political marketplace can be expected to produce effective conservation policies only when those who sacrifice in the present can expect to receive benefits in the future.
Challenges of ecosystem-based fisheries management FAO (2001b) (Chapter 22, this volume) observes that as management expands its focus from target stock to ecosystem, all of these problems [identified above] increase in an exponential way and biological uncertainty becomes ecological uncertainty, which is even more complex. The number of competing users increases, as do the resulting conflicts of interest. Objectives become more complex and conflicting, and the number of stakeholders is expanded to include all the users of all the different ecosystem components. Of course, this expanding complexity is a result of recognizing the reality of the inter-dependence of all ecosystem components, instead of the false assumption that stocks are independent.
In addition, ecosystem-based fisheries management will greatly expand the number of government agencies involved, and the need for interagency collaboration, coordination and cooperation. The resulting governance system becomes a multidimensional web of complex linkages. Hennessey (1997) aptly refers to this web as the ‘ecology of governance’.
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Above, we developed a model of fisheries governance in the context of a single fishery, or a simple set of fisheries, and conclude that the governance dynamics tend to be biased against conservation of the fishery resources and the generation of long-term sustainable economic benefits. As the dimensionality of fisheries governance expands to accommodate ecosystem-based management, we fear that the chances of governance failure will increase. The unfortunate implication is that attempts to implement ecosystem-based management programmes may actually slow progress towards achieving a future of sustainable fisheries. The chances of governance failure will increase because the lack of strong property rights in the marine environment breeds shortsightedness among most stakeholders. Since the number of stakeholders will multiply under ecosystem-based management, there probably will be more efforts by more interest groups to align themselves with the short-term interests of elected representatives. If this is borne out, the bias against conservation of the marine environment will only deepen.
Fisheries governance reforms How then might we correct our fisheries governance institutions and dampen, if not eliminate, the bias against conservation that may only deepen with ecosystem-based fisheries management? One obvious way to avoid government failure is to privatize the fishery resource. Privatization of government-owned and operated industries has been carried out in several countries for several industries, including railways, utilities, communications and energy industries. Complete privatization of fisheries is rare. However, there is a recent trend to devolve more management responsibility down to users, in effect privatizing many management functions. See OECD (1997) for descriptions and analysis of some co-managed fisheries. We use the term ‘self-governance’ to describe a fishery in which users of the resource,
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without the participation of government, take all of the principal management decisions. If all external effects can be internalized by selfgovernance (and there are no other conditions that cause market failure), a fishery has the potential to operate efficiently. While today there is a trend towards co-management, we see no pronounced trend towards selfgovernance of fisheries. Numerous obstacles lie in the way of privatizing a fishery, and complete self-governance may not be desirable from an efficiency perspective. A partial step towards self-governance would be to devolve to users and others with strong interests in a fishery a greater share in the rights and responsibilities of setting management policies and bearing the full consequences of those policies. This would couple the future benefits more strongly with the current sacrifices needed for effective conservation. In other words, users should be assured of reaping the benefits in proportion to their sacrifices. Such action would better harmonize the interests of users and managers with the nation’s interest. Another reform is to implement the principle of beneficiaries paying in proportion to the benefits they receive. In our current system, too many policies and programmes provide benefits for a select few, and impose widespread costs. The costs of fishery management need to be recovered from the beneficiaries of that management. This would mean collecting fees from users of fishery resources. Properly designed and implemented cost recovery can have sizeable beneficial effects on the performance of fishery management by minimizing the opportunities for the political marketplace to produce fishery products with de-coupled costs and benefits. Cost recovery in fisheries appears to be spreading. Australia collects 100% of the attributable costs associated with management and research through levies. In 1996–1997, this policy resulted in US$10 million being recovered to cover management costs and US$11 million for research and development costs. Canada recovered user fees for the management of its fisheries. In 1997, US$38 million was recovered from commercial fishers. It included licence fees for access to the fishery. In addition, fisheries are expected to cover the
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costs of dockside monitoring and observers at sea. New Zealand recovers costs associated with fisheries management services and conservation services carried out for the benefit of the commercial sector. The principle used for the cost recovery is that costs incurred by the government as a result of the existence of the commercial fishing industry should be recovered from the commercial sector. In 1997, this policy resulted in the recovery of US$23 million – a 9% increase over the previous year. Iceland recovered US$1.8 million from ITQ owners to cover the costs of managing ITQ regulations. Even the USA recently has begun developing a cost recovery programme for its IFQ fisheries. Lastly, fishery managers need to be protected from the shortsighted tendencies of elected representatives. Political interference is common in fishery management throughout the world. A mechanism must be found that is both consistent with democratic principles and that limits political involvement in fishery management to the strategic level. Elected officials need to leave to the designated management institutions the dayto-day responsibility for developing and implementing fishery management plans.
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Creed, C.F. and McKay, B.J. (1996) Property rights, conservation and institutional authority: policy implications of the Magnuson Act reauthorization for the Mid-Atlantic region. Tulane Environmental Law Journal 9 (2), 245–256. DFO (Department of Fisheries and Oceans) (2000) DFO Science Stock Status Reports 2000. Canadian Stock Assessment Secretariat, Ottawa, Ontario. FAO (2001a) Basic principles of ecosystem management. Section 3.2.8 in World Fisheries and Aquaculture Atlas. FAO, Rome. FAO (2001b) Towards Ecosystem-based Fisheries Management. A background paper prepared for the Reykjavik Conference on Responsible Fisheries in the Marine Ecosystem, Reykjavik, Iceland, October 2001. Hennessey, T.M. (1997) Ecosystem management. In: Soden, D., Lamb, B. and Tennert, J. (eds) Ecosystems Management: a Social Science Perspective. Kendal/Hunt, Dubuque, Iowa. Hiatt, T. and Terry, J. (2000) Stock assessment and fishery evaluation report for the groundfish fisheries of the Gulf of Alaska and the Bering Sea/Aleutian Island area: economic status of the groundfish fisheries off Alaska, 1999. National Marine Fisheries Service (NMFS), Seattle, Washington. Horwood, J.W., Nichols, J.H. and Milligan, S. (1998) Evaluations of closed areas for fish stock conservation. Journal of Applied Ecology 35, 893–903. IUCN (International Union for the Conservation of Nature and Natural Resources) (1988) Resolution 17.38 of the 17th General Assembly of the IUCN. IUCN, Gland, Switzerland and Cambridge, UK. IUCN (1994) Guidelines for Protected Area Management Categories. IUCN Gland, Switzerland, and Cambridge, UK. Johnson, R. and Libecap, G. (1994) The Federal Service System and the Problem of Bureaucracy: the Economics and Politics of Institutional Change. University of Chicago Press, Chicago. McClanahan, T.R. and Kaunda-Arara, B. (1996) Fishery recovery in a coral-reef marine park and its effect on the adjacent fishery. Conservation Biology 10, 1187–1199. Niskanen, W. (1971) Bureaucracy and Representative Government. Aldine Press, Chicago, Illinois. NRC (National Research Council) (1999) The Community Development Quota Program in Alaska. National Academy Press, Washington, DC. NRC (2001) Marine Protected Areas: Tools for Sustaining Ocean Ecosystems. National Academy Press, Washington, DC.
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OECD (Organisation for Economic Co-operation and Development) (1997) Towards Sustainable Fisheries: Economic Aspects of the Management of Living Marine Resources. OECD, Paris. OECD (2000) Review of Fisheries in OECD Countries. Volume I: Policies and Summary Statistics. Volume II: Country Statistics. OECD, Paris. Olson, M. (1964) The Logic of Collective Action. Harvard University Press, Cambridge, Massachusetts. Russ, G.R. and Alcala, A.C. (1996) Marine reserves: rates and patterns of recovery and decline of large predatory fish. Ecological Applications 6 (3), 947–961. Salm, R.V., Clark, J. and Siirila, E. (2000) Marine and Coastal Protected Areas: a Guide for Planners and Managers. IUCN – The World Conservation Union, Washington, DC. Sherman, K. and Alexander, L. (eds) (1986) Variability and Management of Large Marine Ecosystems. AAAS Selected Symposium 99. Westview Press, Boulder, Colorodo. Sherman, K., Grosslein, M., Mountain, D., O’Reilly, J. and Theroux, R. (1988) The continental shelf
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ecosystem off the northeast coast of the United States. In: Postmas, H. and Zijlstra, J. (eds) Ecosystems of the World. 27: Continental Shelves. Elsevier, Amsterdam, The Netherlands. Sutinen, J.G. (1999) What works well and why: evidence from fishery management experiences in OECD countries, ICES Journal of Marine Science 56, 1051–1058. Sutinen, J.G. (2000) A Framework for Monitoring and Assessing Socioeconomics and Governance of Large Marine Ecosystems. NOAA Technical Memorandum NMFS-NE–158. Northeast Fisheries Science Center, Woods Hole, Massachusetts. Sutinen, J.G. and Upton, H.F. (2000) Economic perspectives on New England fisheries management. Northeast Naturalist 7, 361–372. Wolf, C. Jr (1988) Markets or Governments: Choosing Between Imperfect Alternatives. MIT Press, Cambridge, Massachusetts. World Bank (1995) Monitoring and Evaluation Guidelines for the World Bank – GEF International Waters Projects. The World Bank, Environment Department, Washington, DC.
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The Role of Harvest Control Laws, Risk and Uncertainty and the Precautionary Approach in Ecosystem-based Management Douglas S. Butterworth1 and A.E. Punt2
1Department
of Mathematics and Applied Mathematics, University of Cape Town, Rondebosch, South Africa; 2School of Aquatic and Fishery Sciences, University of Washington, Seattle, USA
Abstract The traditional fisheries management approach involves scientists providing their best assessment of the status and productivity of a resource. They then use these results to recommend a control measure, such as a total allowable catch (TAC), based upon some harvest control law, which is usually associated with a biological reference point (e.g. F0.1). Superficially, the operational management procedure (OMP), or equally the management strategy evaluation (MSE), approach for providing TAC recommendations may appear identical, as this often also links the results from some form of assessment to a harvest control law. However, the key difference is that the OMP/MSE approach involves simulation testing of the whole process that gives rise to the TAC recommendation within an adaptive management framework. This testing includes checks that application of the control law adopted will not lead to major problems, even if key perceptions about the resource happen to be in error; in other words, explicit account is taken of scientific uncertainties, in the spirit of the precautionary approach. Furthermore, quantitative evaluations are provided of the levels of catch to be anticipated in the medium term, and how these trade off against levels of risk of unintended depletion of the resource, to provide managers with a readily interpretable basis to choose between different management options. However, the process involves some problems in defining risk, which have yet to be resolved. Examples where ecosystem considerations have been taken into account in extending this OMP/MSE approach beyond the single-species level can be divided conveniently into two broad categories, depending on whether they concentrate primarily on operational (e.g. by-catch) or biological (e.g. predator–prey) interactions between species, and examples are given of each. To date, actual practical applications of this approach are more readily found for cases of operational interactions, particularly in the area of marine mammal by-catch. For practical applications involving biological interactions, the key limiting factor thus far is the paucity of data to estimate the form and magnitude of predation and competition interactions, which precludes confident computation of the trade-offs between harvest policy options that differ in the extents to which they concentrate upon different species. Nevertheless there are approximate approaches for dealing with this problem. We recommend the use of such approaches, while recognizing their limitations, until the data needed to develop more reliable models of biological interactions become available.
© 2003 by FAO. Responsible Fisheries in the Marine Ecosystem (eds M. Sinclair and G. Valdimarsson)
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Introduction Even leaving aside ecosystem considerations, our title covers as many as four topics. While the relationships between some of these are rather obvious – for example, greater uncertainties relate to increasing risk – those between others perhaps are not as self-evident. Our first task, therefore, is to make the argument that what has been called the ‘operational management procedure’ (OMP) (Butterworth and Bergh, 1993; Butterworth et al., 1993, 1997; Cochrane et al., 1998; De Oliveira et al., 1998a,b; Butterworth and Punt, 1999) or, equivalently, the ‘management strategy evaluation’ (MSE) (Smith, 1994; Smith et al., 1999; Punt et al., 2001) approach to, fisheries management provides a framework within which these four topics are readily related and integrated. The customary process used by fishery scientists to provide advice on control measures for a resource involves two steps. First, the available data are integrated through a mathematical process called stock assessment to provide a ‘best’ assessment of the status and productivity of the resource – for example, its size, possibly expressed as a fraction of that before exploitation commenced, and the level of catch that is sustainable. Then a formula – a harvest control law – is used to translate this information into, say, a total allowable catch (TAC) recommendation for the coming year. For example, the rule could be to recommend a TAC that is a fixed fraction of the estimated resource size, where that fraction is chosen to move the resource towards some level considered to provide optimal exploitation. Often the fraction harvested, or the target abundance level, are chosen from ‘biological reference points’ in common use (e.g. F0.1). Superficially, the end product of the OMP/MSE approach (referred to hereafter as an OMP) often appears nearly identical: some method of estimating current resource abundance (whether in t or in terms of some relative index) and a harvest control law or formula to convert this to a TAC. Where the key difference lies is in the method used to select the law (or, more accurately, the
combination of the data inputs, the basis for estimating abundance and the harvest control law that comprises the complete OMP). The OMP used to recommend TACs for the directed sardine fishery in South Africa (De Oliveira et al., 1998b) provides a particularly simple example to illustrate this difference. This is because the data input specified – the annual hydro-acoustic spawning biomass survey of this resource – provides the requisite abundance index directly. The harvest control law is simply to set the TAC recommendation as 13.75% of this hydro-acoustic abundance estimate, subject to the constraints that the TAC should not reduce by more than 25% from that for the previous year. Note the adaptive nature of this approach: the TAC is raised or lowered in response to monitoring results that indicate respectively an increase or decrease in resource size; but on what basis were the 13.75 and 25% figures selected? The selection process involves projecting the resource status forward for a period of typically 10–20 years in computer simulations, where the future catches are set by applying candidate OMPs with their harvest control laws. For example, in the sardine case above, the candidates reflect alternative values for the harvest percentage (the 13.75%). The choice between these alternatives is then made on the basis of anticipated performance as indicated by the statistics that summarize the results of the simulations. These ‘performance statistics’ typically include measures relating to the size and variability of catches, and also to the risk of depleting the resource below a level at which future recruitment success might be seriously impaired. Conducting such simulations requires a basis (or ‘model’) to predict how the resource will respond to different future levels of catch. If only the ‘best assessment’ of the resource was used for this purpose, again there would be little difference from the traditional approach. The major novel feature of the OMP selection process is the formal recognition that this ‘best assessment’ may well be wrong, so that simulations are conducted for several other plausible appraisals of the resource status and
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Role of Harvest Control Laws
dynamics, which reflect the range of uncertainties in present knowledge of the resource and the extent to which management actions will be implemented. The aim is to seek an OMP that is as ‘robust’ as possible to these uncertainties. What this means is that the anticipated performance, as summarized by the performance statistics, should not change appreciably over the range of uncertainties. In other words, we are looking for an OMP that is adaptive in an even stronger sense than described above for the sardine example – we want it to self-correct over time even if some of the assumptions made in developing our ‘best assessment’ were wrong. In the sardine case, for example, checks were made that the OMP (particularly here its harvest control law) would still perform appropriately even if assumptions for hydro-acoustic target strength made in deriving absolute abundance estimates from the surveys were in error by quite substantial amounts. Thus the requirements of the precautionary approach are being met: the OMP selection process provides a formal structure to take account of scientific uncertainties. The uncertainties or potential errors that need to be considered in the OMP evaluation process fall into three categories. They are model errors (in assumptions about the dynamics of the real resource), data errors (regarding how what is measured and used for input to the TAC computations relates to the actual status of the resource) and implementation errors (relating to plausible differences between the management recommendations and what actually occurs in the fishery, such as between the TAC authorized and the actual subsequent catch). The process of selecting an OMP also requires that management objectives be operationalized, so that they can form the basis for the performance statistics used to compare candidate OMPs. Single-species management objectives generally fall into three categories: maximizing catch, minimizing risk to the resource and minimizing catch variability over time. These objectives cannot be satisfied simultaneously; for example, in the case of the sardine fishery mentioned above, increasing the proportion
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of the survey biomass harvested gives greater catches but increases the risk of depleting the resource more than desired. Reducing the maximum percentage drop allowed in the TAC from year to year necessitates a lower level of utilization overall, so that catch reduction at times of downward fluctuations in resource size remain sufficient to control risks. The final selection between candidate OMPs involves decision makers seeking their desired trade-off between these conflicting objectives. This process is facilitated by choosing measures of performance that are readily meaningful to non-scientists. This is straightforward as far as anticipated catch is concerned, and not too difficult for catch variability. The problem lies with risk. This is often expressed as the chance of depleting the resource below some reference level (often the criterion used is 20% of the abun
