Advances in Marine Biology, Volume 34

Advances in MARINE BIOLOGY

VOLUME 34

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Advances in MARINE BIOLOGY Edited by

J.H.S. BLAXTER Dunstaffnage Marine Research Laboratory, Oban, Scotland

A.J. SOUTHWARD Marine Biological Association, The Laboratory, Citadel Hill, Plymouth, England and

P.A. TYLER School of Ocean Science, University of Southampton, Southampton, England

ACADEMIC PRESS San Diego London Boston New York Sydney Tokyo Toronto

This book is printed on acid-free paper Copyright 0 1998 by ACADEMIC PRESS with the exception of Appendices 1,2 and 3 on pages 179 to 199 which are copyright 0Joyeux and Ward All Rights Reserved. No part of this publication may be reproduced or transmitted in any form or by any means electronic or mechanical, including photocopy, recording, or any information storage and retrieval system, without permission in writing from the publisher. Academic Press 24-28 Oval Road, London NWl 7DX, UK http://www. hbuk.co.uk/ap/ Academic Press 525 B Street, Suite 1900, San Diego, California 92101-4495, USA http://www.apnet.com ISBN 0-12-026134-0 Library of Congress Cataloging-in-PublicationData A catalogue record for this book is available from the British Library

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CONTRIBUTORS TO VOLUME 34 M.S. DAVIES, Ecology Centre, University of Sunderland, Sunderland SRI 3SD, UK S.J. HAWKINS, School of Biological Sciences, University of Southampton, Southampton SO17 lBJ, UK S . JENNINGS, School of Biological Sciences, University of East Anglia, Norwich NR4 7TJ, UK J.C. JOYEUX, North Carolina State University, Department of Zoology, box 7617, Raleigh, NC 27695, USA (Present address: Universite‘ Montpellier It Laboratoire d’Hydrobiologie Marine et Continentale, CNRS U M R 5556, case 093, Place E. Bataillon, 34095 Montpellier Cedex 5, France) M.J. KAISER, School of Ocean Sciences, University of Wales, Bangor, Menai Bridge, Anglesey, LL59 5EY, UK A.G. MCARTHUR, Josephine Bay Paul Centre for Comparative Molecular Biology and Evolution, Marine Biological Laboratory, Woods Hole, M A 02543-1015, USA D. MCHUGH, Department of Organismic and Evolutionary Biology, Museum of Comparative Zoology, Harvard University, Cambridge, M A 02138, USA V. TUNNICLIFFE, School of Earth and Ocean Sciences, University of Victoria, Victoria, B.C., Canada V8W 3N5 A.B. WARD,North Carolina State University, D.H. Hill Library, Box 7111, Raleigh, N C 27695, USA

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CONTENTS CONTRIBUTORS TO VOLUME 34

......................................

V

Mucus from Marine Molluscs M.S. Davies and S.J. Hawkins

1. 2. 3. 4. 5. 6. 7. 8.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Composition of Mucus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Properties of Mucus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mucus Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Functions of Mucus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mucus in Molluscan Energy Budgets . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecology. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ~

2 4 8 9 23 35 43 50 51 51

Constraints on Coastal Lagoon Fisheries J.-C. Joyeux and A . B. Ward

1 . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 . Material and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4. Geographical and Morphometrical Constraints . . . . . . . . . . . . . . . . . . . . . 5 . Environmental and Anthropogenic Constraints . . . . . . . . . . . . . . . . . . . . 6 . Final Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Appendix 1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Appendix 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Appendix 3 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

vii

74 77 86 117 134 151 153 154 179 184 194

viii

CONTENTS

The Effects of Fishing on Marine Ecosystems

.

S Jennings and M.J. Kaiser

1. 2. 3. 4. 5. 6. 7.

General Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Benthic Fauna and Habitat . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fish Community Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trophic Interactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Study of Fishing Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Management . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

203 208 236 257 292 303 311 313 314

A Biogeographical Perspective of the Deep-sea Hydrothermal Vent Fauna

.

V Tunnicliffe. A.G. McAtthur and D. McHugh

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. Hydrothermal Vents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3 . Other Related Faunas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4. The Biogeography of Faunas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5. Local to Regional-Scale Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 . Regional to Global-Scale Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7. The Distribution Patterns of Taxa . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8. Patterns in Diversity. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9. Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

355 358 379 385 394 398 402 420 426 426 426

TaxonomicIndex . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SubjectIndex . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cumulative Index of Titles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cumulative Index of Authors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

443 452 457 462

Mucus from Marine Molluscs Mark S. Davies’ and S.J. Hawkins’

.

Ecology Centre. University of Sunderland. Sunderland. SRI 3SD UK ’Biodiversity and Ecology Division. School of Biological Sciences. University of Southampton. Biomedical Sciences Building. Southampton. SO16 7PX. UK Introduction...................................................... Composition of Mucus ............................................. Properties ofMucus ............................................... Mucus Production ................................................. Functionsof Mucus ............................................... 5.1. Locomotion ................................................. 5.2. MucusTrails ................................................ 5.3. Feeding .................................................... 5.4. Protection. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.5. Other Functions .............................................. 6 Mucus in Molluscan Energy Budgets ................................. 7 . Ecology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1. Fateof Mucus ............................................... 7.2. Role in Biological Interactions ................................... 8. O v e ~ i e w ....................................................... Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References ........................................................ 1. 2. 3. 4 5.

.

.

2 4 8 9 23 23 25 27 32 34 35 43

43 45 50 51 51

ABSTRACT

Mucus functions in many invertebrate physiological processes and also influences structuring of the community and the ecosystem. Molluscan mucus is mostly water . The remaining components are protein. carbohydrate and lipid . The detailed structure of the protein-polysaccharide acidic glycosaminoglycan component is not yet known . Mucus is probably released in dehydrated form in distinct. membrane-bound packages. which then absorb water . A functioning mucus is probably formed by mixing of mucins from different types of gland . Under small deformities. hydrated mucus is a viscoADVANCES IN MARINE BIOLOGY VOL. 34 ISBN 0-12-0261340

Copyrighr 0 1998 Academic Press Limited All rights of reproduction in any form reserved

2

M.S. DAVIES AND S.J. HAWKINS

elastic solid, able to function as a rope. As stress increases, it yields to become a liquid which can return to the solid state once the stress is released. It is these properties that allow locomotion by molluscs on what is seemingly an adhesive. On dehydration, the strength and stiffness of mucus increase such that molluscs can suspend their body by a thread of it. Mucus production has been studied quantitatively by various methods, some gravimetric and some colorimentric using pedal, faecal, pseudofaecal and hypobranchial mucus: there is much spatial and temporal variation. In locomotion mucus is a coupling agent between foot and substratum; a medium in which propulsive cilia beat; and a drogue. Mucus deposited as a trail by gastropods is an important facet of their environment. Many species follow mucus trails possibly contributing to the observed patchy distributions of gastropods. The methods by which the presence and polarity of mucus trails are detected is poorly understood. Mucus plays a vital role in feeding. In filter-feeding bivalves, mucus aids the transport of food from gill to mouth and is employed to cleanse the mantle cavity of particles rejected by the labial palps. In gastropods mucus nets and bags are used to trap food prior to ingestion and some groups roll their prey in mucus to prevent its escape. Pedal mucus may be ingested after it has become studded with organic material and perhaps act as a fertilizer for microbial growth. A copious secretion of epithelial mucus is used to isolate molluscs from their environment and mucus may also serve as an ionoregulator. Mucus may also contain specific products to render the animal poisonous, distasteful or irritating. Agglutinin and lysozyme have been found in mucus from marine molluscs. Mucus secretion can present a considerable drain of energy (up to 70% of consumed energy). The fate of molluscan mucus is largely unknown and probably makes a considerable contribution to POM in inshore waters, although its is readily degradable by marine microbes. Given the persistence of mucus, densities of benthic gastropods and their motility patterns, much of the gastropod-inhabited benthos is likely to be covered for most of the time with a layer of pedal mucus.

1. INTRODUCTION

Mucus is essential to the function of marine molluscs. Mucus has a function in almost all physiological processes and plays a role beyond the use of its producers, at the levels of community and ecosystem. Nevertheless the literature is relatively sparse - particularly in comparison to the vast literature on mammalian mucus and its physiological functions. Hence comparatively

3

MUCUS FROM MARINE MOLLUSCS

little is known of the composition, structure, functions and fate of molluscan mucus. Recently there has been an increase in the awareness of the importance of mucus in the functioning of marine molluscs and its role in the ecosystem. Our review draws together the current knowledge concerning mucus from marine molluscs, synthesizes ideas and points the way towards future work. We first discuss the nature of mucus, its properties, production sites and functions. The importance of mucus in physiological and ecological energetics is then considered and finally its significance at the community and ecosystem level is assessed. Our discussions are confined in the most part to the secretions of marine molluscs, but we include relevant information on terrestrial and freshwater molluscs and other animal phyla. Table 1 summarizes key papers on mucus from other groups. We must first clearly define what we are discussing. The nomenclature of mucus is confused. We follow the nomenclature of Reid and Clamp (1978) who proposed a standard, simplified, terminology. Polymeric substances consisting of carbohydrate covalently linked to non-carbohydrates, usually lipid or protein, are glycoconjugates and exist in combination with protein as either proteoglycan (glycosaminoglycan) or glycoprotein depending on Table 1 Important references on mucus in non-molluscan groups.

Group/topic

Reference ~~

Invertebrate mucus: structure Invertebrate mucus: functions Invertebrate mucus: agglutinating and antibacterial activity Mammalian mucus: biochemistry Fish Fish: defence Fish skin: drag reduction Elasmobranch skin: composition Fish gills Tunicates: mucus filtration mechanism Nemerteans (terrestrial and freshwater) Coral: composition Coral: production and fate Sponges: antimicrobial activity Macroalgae Diatoms Foraminifera: bacteria “farming” Phytoplankton aggregates Microbes (marine) Microbes Plant roots

~~

~~

~~

Hunt (1970) Denny (1989) Astley and Ratcliffe (1989) Strous and Dekker (1992) Shephard (1994) Alexander and Ingram (1992) Bernadsky et al. (1993) Tromeur et al. (1992) Lumsden et al. (1994) Flood et ul. (1992) Moore and Gibson (1985) Meikle et ul. (1988) Coffroth (1990) Muricy et al. (1993) Evans (1989) Hoagland et al. (1 993) Langer and Gehring (1993) Kirarboe and Hansen (1993) Decho (1990) Sutherland (1989) Rougier and Chaboud (1989)

4

M.S. DAVIES AND S.J. HAWKINS

the protein-carbohydrate linkage. Mucus is used both as adjective and noun and the latter is regarded as the slimy secretion of any epithelial surface. The purified glycoconjugate from mucus is called mucin. Usage of vague terms, particularly those prefixed by “muco-”, such as “mucoprotein”, “mucopolysaccharide” and “mucosubstance” (excepting “mucocyte” which is any cell which produces mucus), is not encouraged. We advocate the abandonment of the adjectival “mucous” in favour of the shorter “mucus”, used as an adjectival noun.

2. COMPOSITION OF MUCUS

The major component of molluscan mucus is water (Wilson, 1968; Hunt, 1970; Grenon and Walker, 1980; Connor, 1986, Davies et al., 1990b) which forms from 81.4% (w/w) of wet weight in pedal mucus from Nucella emarginata (Connor, 1986) to 99.8% (w/w) in hypobranchial mucus from Busycon canaliculatum (Shashova and Kwart, 1959). Table 2 gives a summary of the constituents of mucus. There is mounting evidence that, at least for terrestrial species, the majority of this water is absorbed from the environment, and that the mucus is released in a membrane-bound dehydrated form (Kapeleta et al., 1996). This leaves 0.2% to 18.6% (w/w) as solid matter which can be divided into two groups: the high molecular weight protein-polysaccharide complexes and the inorganic salts. Grenon and Walker (1980) found the inorganic salts formed 3.1% (w/w) of the hydrated pedal mucus of Patella vulgata and consisted of sodium, magnesium and calcium ions, of which almost two-thirds were the divalent cations. Other analyses have been performed on the pedal mucuses of the terrestrial Helix pomatia (Burton, 1965) and the freshwater Lymnaea truncatula (Wilson, 1968); both these mucuses were shown to contain sodium and potassium ions. Inorganics in the hypobranchial mucus of Busycon canaliculatum amounted to up to 12% of dry weight and their main ionic components were sodium and calcium (Shashova and Kwart, 1959). There is a distinct seasonal variation in the composition of the pedal mucus of Patella vulgata (Davies et al., 1990b), which may be related to the reproductive cycle, reflecting similar variation in somatic tissues (Barry and Munday, 1959; Blackmore, 1969b). Ash content showed a maximum (-60% of dry mucus weight) in spring when the gonads were inactive. Protein (-36%) and carbohydrate (1620%) also reach maxima during the spring or early summer and show minima (-30% and 14-18%, respectively) towards the end of the spawning period in early winter, suggesting that energy might be diverted from mucus production into gonad development. Davies et al. (1990b) recorded a higher carbohydrate content in the

MUCUS FROM MARINE MOLLUSCS

5

mucus from small limpets than in that from larger limpets. The mucus produced by intertidal gastropods must function over a wide range of environmental conditions, for each of which a specific mucus composition may be most appropriate. Thus seasonal variation in, for example, temperature, may equally be responsible for variations in composition. Following CHN analysis of the pedal mucus of the Antarctic limpet, Nacella concinna, Peck et al. (1993) determined protein to be 31.1YO, carbohydrate to be 12.0% and lipid to be 2.5%. Similar values were recorded for Californian acmaeid limpets by Connor (1986) (Table 2). Iwasaki (1992) discovered that in the limpet Cellana grata, the protein content of trail mucus was some 16 times less than in the mucus produced while stationary; and the protein content of “stationary” mucus increased the longer the animals remained stationary (up to 24 h tested) (Table 2). According to Reid and Clamp (1978) and Denny (1983), the proteinpolysaccharide complexes (high molecular weight) components of mucus have attracted more work since it is these which confer on the mucus its particular and peculiar properties. In particular, specific complexes can have specific functions (see Section 3). These complexes have traditionally been split into two groups according to their structure: glycosaminoglycans (GAGs; often called mucopolysaccharides or proteoglycans) and glycoproteins (Reid and Clamp, 1978). In invertebrates there is a blurred distinction between these two groups since structurally they represent the endpoints of a continuum (Denny, 1983). The glycoproteins are composed of short, often branched, carbohydrate chains (oligosaccharides) bound to a large protein component which is important in determining the properties of the mucus (Gottschalk, 1972). In contrast, the GAGs are made up of high molecular weight, long and typically linear carbohydrate chains (Gottschalk, 1972) and the attached protein contributes little to the properties of the mucus (Denny, 1983). These complexes occur widely in invertebrates (see Hunt, 1970, for review). An account of the physical chemistry of the GAG present in the hypobranchial mucus of Busycon canaliculatum is given by Kwart and Shashova (1958). Although it has been established that glycoprotein and GAG do not occur in isolation from each other, for example, in the hypobranchial secretion of Buccinum undatum (Hunt, 1967, 1970), it is recognized that GAGS are far more common than glycoproteins in molluscan mucuses (Denny, 1983). Glycoproteins are conjugated proteins where the prosthetic group or groups are constituted by covalently bound heterosaccharide units (see Gottschalk, 1972 for details on glycoproteins). Each macromolecule has a molecular weight of between 10 x lo6 and 45 x lo6. The carbohydrate constitutes about 80% (w/w) of the mucin and occurs in clusters along a protein core. Each cluster has a molecular weight of 500 to 4000 and can be neutral or acidic, branched or linear and contains from one to about 20 monosac-

Table 2 Summary of biochemical content of mucus from aquatic gastropods. Values are mean percentages f standard errors. Sample sizes in parentheses.

Species (type of mucus) Patella vulgata (large) (pedal) (small) Patella vulgata (pedal)

91.9 f 0.2 (88) 91.6 f 0.3 (77) 90.1

Buccinum undatum

99.7

(hypobranchial) Buccinum undatum (hypobranchial)

ash

protein

49.3 f 1.3 (65) 31.7 f 0.4 (39) 47.6 f 2.5 (53) 32.5 f 0.3 (39) 30-40 32.8

carbohydrate

lipid

14.4 f 0.1 (39) 18.8 f 0.2 (39) 12.0

99.599.8

Lymnaea truncatula Wal) Lottia gigantea (pedal) Collisella scabra (pedal) Collisella digitalis (pedal) Nucella emarginata (pedal) Nacella concinna (pew Cellana toreuma (pedal)

91.8 (27)

Reference Davies et al. (1990b) Grenon and Walker (1980) Hunt (1970)

57.9

Busycon canaliculatum (hypobranchial)

v

Proportion of dry weight

Wet weight/ dry weight (water content)

Kideys and Hartnoll (1991) Shashova and Kwart (1959) Wilson (1968)

u p to 12

93.5 f 0.9 (3)

43.0 f 2.0 (3)

36.8 f 7.5 (3)

18.4 f 0.2 (3)

0.3 f 0.3 (3)

Connor (1986)

89.3 f 0.9 (3)

47.0 f 1.5 (3)

36.2 f 2.4 (3)

17.1 f 0.9 (3)

0.4 f 0.1 (3)

Connor (1986)

90.6 f 0.6 (3)

45.0 f 2.5 (3)

29.7 f 3.9 (3)

8.1 f 1.3 (3)

0.8 f 0.5 (3)

Connor (1986)

81.4 f 0.5 (3)

75.7 f 1.3 (3)

1.30 f 0.8 (3)

25.4 f 3.5 (3)

0.2 f 0.2 (3)

Connor (1986)

12.4 f 1.08 (10)

2.51 f 0.73 (10)

Peck et al. (1993) Iwasaki (1992)

"Values in pg cm-2 of deposited mucus.

50.3 f 2.7 (10) 31.1 f 2.05 (10)

0.374 f 0.02 (14) (trail>" 14.72 f 1.14 (8) (stationary, after 6h)'

MUCUS FROM MARINE MOLLUSCS

7

charide units, most commonly mannose, galactose and fucose. These tightly packed oligosaccharide units protect the core from proteolytic enzymes (Gottschalk, 1972). This type of mucus can be fragmented by the reduction of disulphide bonds (see Clamp et al., 1978, for review). The fine structure of the clusters is reviewed by Hunt (1970). The viscosity of vertebrate mucus is increased by the presence of sialic acid-substituted glycoproteins. In invertebrates this role is more commonly filled by acidic GAGs (Hunt, 1970) and sialic acid generally appears absent (Warren, 1963; Denny, 1983). Inoue (1965), however, identified sialic acid in hydrolysates from the digestive gland of the whelk, Charonia lampas as did Cottrell et a f . (1993) from the trail mucus of the terrestrial slug, Arion ater ( < 1 % of solids). Although less is known about the structure of GAG, it is related to that of glycoprotein (Gottschalk, 1972). Livingstone and de Zwaan (1983) give a limited account of GAG structure which they describe as heteropolysaccharides combined with specific proteins. The sugar groups are composed of two types of alternating monosaccharide units, at least one of which bears an acidic (carboxyl or sulphuric) group. The presence of uronic acid gives each GAG an anionic charge (sometimes increased by the presence of sulphate groups) which allows its separation from proteins and glycoproteins. The precise structure varies from mucin to mucin (Chandrasekaran and BeMiller, 1980) and anything other than a superficial structure has yet to be elucidated. The bonds linking protein and carbohydrate in gastropod GAGs have been shown to be covalent, alkali stable and 0-glycosidic (Hunt, 1970; Grenon and Walker, 1980), but this apart, little is known of their nature. Fountain (1982) suggested that agglutinating lectins might have a structural role (cross-linking carbohydrate and protein) in the mucuses from a wide variety of plants and animals, including a terrestrial pulmonate. Detailed accounts of mucus structure, giving information on, for example, the precise nature of some of the proteinxarbohydrate bonds are given by Clamp et a f . (1978) and Strous and Dekker (1992). These details are solely for mammalian mucus and so may not be consistent with molluscan findings. Shashova and Kwart (1959) proposed a molecular structure for the hypobranchial mucus of Busycon canaliculatum, based on the evidence available at the time, which suggested a polyhexoseammonium sulphate linked to an acidic protein moiety through calcium ions. Despite this, as far as the Mollusca are concerned, the exact biochemical composition and the linkage of the various components into a final mucus gel have yet to be understood, although Cottrell et al. (1993) present some valuable though limited information on the trail mucus of terrestrial slugs. Denny (1983) gives an account of what is known for the Mollusca and his comment that there are insufficient data to enable a model of molluscan mucus to be drawn is still true. It is clear then that much more biochemical and molecular work is needed to

8

M.S. DAVIES AND S.J. HAWKINS

understand how the mucus of molluscs is able to act in such a wide range of activities (see Section 5).

3. PROPERTIES OF MUCUS The properties of molluscan hypobranchial and pedal mucuses were reviewed by Denny and Gosline (1980) and Denny (1983, 1984). Even the simple physical properties remain to be fully explored, probably owing to the complex and heterogeneous nature of mucus. Solubility of the pedal mucus of Patella vulgata in a variety of solvents was described by Grenon and Walker (1980). Mucus was insoluble in seawater and distilled water, salt solutions, common organic solvents and in Nacetyl-L-cysteinewhich according to Davis et al. (1975) is a strong mucolytic agent. Grenon and Walker found mucus to dissolve in solutions of sodium hypochlorite and alkaline sodium sulphate. In our tests with mucus from the same source, however, these solutions did not dissolve the mucus, although hypochlorite did cause the mucus to disaggregate after about 30 min. Trials with l,l, 1-trichloroethane, which can break up vertebrate mucuses (pers. obs.), had no effect on pedal mucus from P . vulgata. Not surprisingly, 16M HN03 effectively broke up the mucus. The precise rheological properties of mucus are critical in enabling the mucus to function; the strength of mucus should not be underestimated. Denny (1989) reported the tensile stiffness of the mucus “ropes” of the terrestrial slug Limax maximus to be of the order of lo5 N m-*. The shear stiffness of gastropod pedal mucuses ranges from 200 to 400N m-2 (Denny, 1983). It is this stiffness that enables mucus to function as a tensile element in feeding strands, webs, nets, bags and curtains (see Denny, 1989); and a mucus secretion, loaded with protein, forms the crystalline style of bivalves and some suspension-feeding gastropods (Barnes, 1980). Under small deformities pedal mucus is a viscoelastic solid. As stress is increased the mucus yields to become a viscous liquid, the yield strength being proportional to the rate of deformation (see also Simkiss and Wilbur, 1977). Solidity, increasing with time, returns if the mucus is allowed to heal undeformed. Concordantly, the hypobranchial mucus of the whelk Buccinum undatum shows a Weisman effect - an elastic recoil - when stirred (Hunt and Jevons, 1963, 1966). It is these properties whch enable the familiar crawling locomotion of slugs and snails. Under the leading edge of a locomotory (muscular) wave travelling along the sole of a slug or snail, mucus is stressed and as a result flows, allowing the wave to progress. As the wave passes over the mucus the viscosity of the mucus increases. In many species this provides the adhesive coupling

MUCUS FROM MARINE MOLLUSCS

9

between substratum and mollusc which allows slugs and snails to adhere to walls and ceilings (see Denny, 1981, for a quantitative description of this for terrestrial slugs). It should be noted that mucus is not the sole tenacityenabling agent. Smith (1991, 1992) noted that suction can play a role in the adhesion of some limpet species, but presumably mucus helps to produce a good seal. Denny (1984) suggested that the rheological properties of mucus may place constraints on gastropod structure and performance. Denny measured the properties of Ariolimax columbianus mucus and derived a theoretical maximum speed of locomotion (0.6 mm s-I), comparable to speeds observed (0.8-2.3 mm s-'), by calculating the time taken for mucus to flow under the stresses imposed by the foot of a slug. Denny also calculated theoretical tenacities based on the properties of mucus and predicted the maximum size for molluscs in wave-swept environments (17.3 cm for a limpet-shaped organism), again with reasonable accuracy (typical maximal size, 25cm). Molluscan mucus also has a remarkable capacity for absorbing water when it is first secreted (Verdugo, 1990, see Section 4), although this ability in marine molluscs has not been assessed. When dehydrated both the strength and stiffness of pedal mucus increase substantially, such that littorinid snails can attach to rock walls (Wilson, 1929; Bingham, 1972; Denny, 1984) and hang from rock ceilings (Davies and Hawkins, pers. obs.) using a strand of dehydrated pedal mucus. Denny (1984) found the mean breaking strength for such mucus produced by Littorina aspera to be lo8N m-2. Patterns of dehydation of the pedal mucus of Patella vulgata with temperature, wind speed and relative humidity, point to > 50% dehydration - and often total dehydration - of exposed mucus on mid-upper shores during emersion (Davies et al., 1992~).In seawater, rehydration of the mucus (from a totally dehydrated state) is slow (from -10% to -20% of hydrated weight over a 6-h period) and although after this period the mucus regains its stickiness, such dehydration means that the mucus is unlikely to recover to a fully hydrated state. Davies et al. (1992~)also found in laboratory experiments that the pedal mucus of P. vulgata degrades (loses weight) in seawater at a rate of 10% 6 h-'. This degradation was ascribed to the mechanical action of moving seawater.

-

4. MUCUS PRODUCTION

The structure and the secretions of the molluscan epidermis were reviewed by Simkiss and Wilbur (1977). These authors also provided a comprehensive account of the functioning of the epidermis in the transfer of substances, including mucus, across the skin. They report that mucus as a functional

10

M.S. DAVIES AND S.J. HAWKINS

substance is often a product of various glands mixed with a general exudate from epithelial cells. Parker (191 1) suggested that during gastropod locomotion a “high viscosity” mucus was secreted by the suprapedal glands and a lower viscosity mucus was secreted by the sole, although there is no evidence to support this suggestion. Numerous authors (Deyrup-Olsen et al., 1983,1992; Martin and DeyrupOlsen, 1986; Verdugo, 1990; Luchtel et al., 1991; Deyrup-Olsen, 1996; Deyrup-Olsen and Jindrova, 1996; Kapeleta et al., 1996) have reported that the mucus of terrestrial slugs is released as membrane-bound “granules”. These are typically of size -5-10pm (Kapeleta et al., 1996). The bursting of these granules can be triggered by elevated pH (Deyrup-Olsen, 1996), certain carbohydrates (Deyrup-Olsen and Jindrova, 1996), mechanical force (e.g. during locomotion), cold shock and micromolar levels of ATP (Deyrup-Olsen et al., 1992). Deyrup-Olsen et al. (1992) obtained evidence that initiation of granule swelling may occur via an ATP receptor on the granule membrane which triggers calcium channels, allowing a large efflux which in turn triggers swelling. Hydration then occurs, in the presence of water, to a volume of 150-8OOx (Verdugo, 1991; Kapeleta et al., 1996) to produce a functioning mucus. In many cases such products from several different types of mucocyte (e.g. those on the sole of the foot of a gastropod, Shirbhate and Cook, 1987), will combine to produce a multifunctioning mucus (see Section 2). It seems likely that marine molluscs also release their mucus in discrete microscopic packets. Although this has yet to be demonstrated conclusively, pedal mucus from the limpet Patella vulgata can be seen emerging in packets under TEM (Figure 1). This has also been seen in the periwinkle Littorina litforea (Davies, in press). Mucus “filaments” ( < 1 pm wide, > l00pm long) and associated granules in the locomotory trails of Helix aspersa (Simkiss and Wilbur, 1977), Ilyanassa obsoleta (Bretz and Dimock, 1983) and littorinids (Tankersley, 1990; Davies and Hutchinson, 1995) have puzzled their observers, but may be related to the hydration of mucus. Histological examination of the gastropod epidermis has attracted much attention (see Simkiss and Wilbur, 1977; Grenon and Walker, 1978; Shirbhate and Cook, 1987 for reviews) as has the functions of the gastropod epidermis such as respiration (e.g. Jones, 1961), osmoregulation (e.g. Greenaway, 1970) and tenacity (Grenon and Walker, 1981). However, there has been little attempt to relate such function to the detailed structure of the epidermis. Mucus secretion is no exception in this respect even though it is a characteristic feature of the molluscan epidermis and is important in a wide range of physiological processes. Branch and Marsh (1978) described the structure of the foot in six Southern African Patella species in terms of interspecific differences in mucus secretion and tenacity. They found that high tenacity was associated with a low mucus secretion rate and a small

MUCUS FROM MARINE MOLLUSCS

11

Figure 1 Transmission electron micrograph showing the ultrastructure of the pedal sole of the limpet Putella vulgatu. The epithelial layer of columnellar cells is above (ventral to) the position of the basement membrane (bm). All epithelial cells are microvillose (mv) and some are ciliated (c). Mucus is contained in packets within the vacuoles (v) of the P9 (see Figure 2 ) mucocytes and is discharged (arrowed) onto the sole. Scale bar = 5pm.

number of pedal mucocytes, suggesting that limpet tenacity may be facilitated by Stefan adhesion, although calculations by Smith (199 1) suggest that limpets do not use Stefan adhesion. Grenon and Walker (1978) described histologically and biochemically the structure of the foot and pedal gland system of P . vulgara and proposed functions for each of nine gland types (P1to P9) identified (Figure 2). Six of the gland types release their secretions onto the foot sole and three onto the side wall. The sole epithelium consists of three cell types: non-ciliated cells, ciliated cells and P9 goblet cells (mucocytes) (Figure 1). These mucocytes are

12

M.S. DAVIES AND S.J. HAWKINS

Figure 2 Sterogram of the foot of the limpet Patella vulgata to show the pedal glandular system. PI = marginal gland (granular) composed of many cells discharging a proteinaceous secretion into the marginal groove. P2 = flaskshaped, subepithelial cells discharging acidic and neutral glycosaminoglycans on to the sole via necks opening between epithelial cells. P3 = isolated granular cells situated against the epithelium of the side wall that discharge a proteinaceous secretion via a short neck. P4 = flask-shaped cells found with P3 cells but situated deeper within the foot and discharging glycoprotein. P5 = granular flask-shaped gland cells with long necks opening on to the sole; occur throughout the foot, but increasing in density near the periphery, secrete weakly acidic glycosaminoglycan. P6 = scarce cells 120 ,um from the sole on to which they discharge a granular secretion via a thin neck. P7 = epipodial gland situated in the side wall, secreting weakly acidic glycosaminoglycan into the epipodial streak; present only in young animals. P8 = granular, club-shaped cells immediately below the sole epithelium discharging glycosaminoglycan on to the sole via a neck -60,um long. P9 = epithelial mucocytes occurring away from the periphery of the sole, secreting weakly acidic glycosaminoglycan. e.s. = epipodial streak. m.g. = marginal groove. p.r. = peripheral region. s. = sole. s.1. = subepithelial “space” layer. S.W. = side wall. (Redrawn after Grenon and Walker, 1978.)

-

randomly distributed throughout the foot, except in the peripheral region (Grenon and Walker, 1978), and are by far the commonest type of mucocyte present (pers. obs.). Grenon and Walker (1978) also examined the pedal glands of Acmaea (Tectura) tessulata and identified six gland types secreting on to the sole, homologous in function to the six of P . vulgata. Epithelial goblet cells are present in many prosobranchs (Fretter and Graham, 1994) and are not merely providing a surface lubricant or protective layer as their density is much greater in the sole of the foot than in areas where these functions are more important such as the side wall of the foot (Grenon and Walker, 1978). It is unlikely that they are responsible for producing locomotory mucus as this is probably secreted mostly by the marginal gland in

MUCUS FROM MARINE MOLLUSCS

13

the anterior marginal groove (Fretter and Graham, 1994; Grenon and Walker, 1978). The other, more likely function of P9 glands is adhesion. Grenon and Walker (1978) suggested some pedal glands secrete a highly viscous mucus for adhesive function, while others secrete a less viscous mucus for locomotory purposes. However, Denny and Gosline (1980) showed a single pedal mucus of Ariolimax columbianus was capable of altering its viscoelastic properties under different physical conditions. The acid GAG secretion of the P9 glands (Hunt, 1973; Grenon and Walker, 1978) is indicative of high viscosity in aqueous solution (Hunt, 1973) and suggests an adhesive function (see earlier). The density of P9 cells in the feet of three limpet species, Patella vulgata, P . depressa and P. ulyssiponensis ( P . aspera) was investigated by Davies (in press). He found that there was no significant change in P9 density in P. vulgata populations differing in shore height (at the one site investigated) but that significant differences were apparent (range of means: 63-96 mm-' of epithelium transversely-sectioned at 8 pm) between the species and between P. vulgata populations from shores of different exposure. Shirbhate and Cook (1987) identified five types of mucus-secreting cell in the foot of Littorina littorea (Figure 3). Most mucus is secreted by the two cell types in the anterior groove of the pedal sole which are differentiated on shape and texture and contain neutral GAGs. The remaining three types, two of which are subepithelial, secrete neutral, sulphated and carboxylated GAG. Although most other epithelial surfaces of most marine gastropods are likely to be secretory to some extent, there is a notable lack of information on gland types and discharges. An exception is Littorina littorea whose epithelia are described by Fretter and Graham (1994) and whose opercular gland system has been described by Shirbhate and Cook (1987). Goblet cells likely to be responsible for producing mucus are also apparent on the pallial margin of Patella spp. (Hackney et al., 1983; Hodgson et al., 1987) and are presumably present on most molluscan pallial surfaces. For bivalves, the most obvious site of mucus production is the ctenidia, although surprisingly little work has been done to determine mucocyte distribution and density, especially since such studies can reveal much about the feeding mechanism. Foster-Smith (1975) recorded many more mucocytes on the abfrontal surface of gill filaments than on frontal surfaces in Mytilus edulis, Cerastoderma edule and Venerupispullastra, and their uneven distribution is now established (see Owen and McCrae, 1976). Ahn et al. (1988) identified mucocytes in M . edulis containing neutral and acidic GAGs on the lateral gill surface and neutral or sulphated GAGs abfrontally. A useful and relatively new application to investigate the production sites of mucus and the subsequent movement of mucus is the endoscope (Ward et al., 1991, see Section 5 ) . Beninger et al. (1993) used endoscopy to provide functional correlates to more traditional histological determinations of

14

M.S. DAVIES AND S.J. HAWKINS

,11

- 12

c

0.5mm

I

Figure 3 A longitudinal section through the anterior foot of the periwinkle Littorina littorea showing the composition of the pedal gland, presence of the pedal groove and the different types of mucocytes present. L1 = large, reticular gland cells containing glycoprotein. L2 = cells with long necks that open to the ciliated epidermis within the pedal groove; granular contents. L3 = pyriform cells with necks opening between the epidermal cells of the sole; reticular contents of sulphated glycosaminoglycan. LA = cells generally distributed, but more common anteriorly; reticular or granular appearance; contain carboxylated glycosaminoglycan. L5 = flask-shaped cells with long necks; few in number; granular contents of neutral glycosaminoglycan. L9 = epidermal goblet cells; reticular; contain sulphated glycosaminoglycan. LIO = epidermal goblet cells; granular; contain protein and neutral glycosaminoglycan. (Redrawn after Shirbhate and Cook, 1987.)

mucocyte distribution. These authors used these techniques on Mytilus edulis to trace the origin of the ventral mucus strand, which transports trapped food particles to the mouth, as the mucocytes of the frontal surfaces of the gill filaments and/or the mucus glands within the filaments dorsal to the ventral grooves. Endoscopic observation by Beninger et al. (1992) suggests that in the scallop Placopecten magellanicus there are two types of mucus used in feeding and each has a different viscosity. Particles transported towards the dorsal margin are trapped in a low-viscosity mucus, while those transported ventrally are in a high-viscosity mucus. Beninger et al. (1993) further examined the gills of P. magellanicus and found that on the principal filaments mucocytes increase in number dorsalward and secrete a mixture of mucins. On the interprincipal (“ordinary”) filaments mucocytes

MUCUS FROM MARINE MOLLUSCS

15

mostly contain an acid GAG and increase in number towards the plical crests. Endoscopy revealed the principal filaments to be involved in feeding and the interprincipal filaments in cleaning activities. Beninger et al. (1993) also examined Mytilus edulis and found a more even distribution of mucocytes and a broader range of GAG secretions. Beninger and Dufor (1996) recorded both mucocytes secreting acidic GAG and mucocytes secreting neutral GAG on to the gill of Crassostrea virginica. The density of mucocytes decreased from plical crest to trough and the anterior plicae showed an increased proportion of acid GAG secretory cells. Beninger and Dufor suggested that this would increase the viscosity of mucus as it reaches the labial palps, facilitating the transport of filtered material to the palps (see Ward et al. 1994), and that conditions for the formation of a mucus cord were “a relatively large amount of acid-dominant mucus compressed into a tight space”. Sorting of food particles and a control on their volume entering the mouth is thought to be achieved by the labial palps (see, e.g. Bayne and Newell, 1983) where dense mucocytes occur (Beninger et al., 1995) (Figure 4). Subepithelial mucocytes with acidic secretions are concentrated on the anterior half of palp ridges in M . edulis and epithelial mucocytes with neutral secretions are concentrated on the anterior half of ridge crests. In the troughs subepithelial mucocytes dominate. On the smooth surface of the palps epithelial mucocytes with mostly neutral, but occasionally acidic, secretions are randomly distributed (Beninger et al., 1995) (Figure 4). The functioning of the palps in particle sorting is partially described by Beninger and St-Jean (1997a) and by Beninger et al. (1997a). Drawing on both published and unpublished evidence, Beninger and St-Jean (1997b) suggested that the mucus produced by various pallial organs differs according to the function of each organ. They postulated that: viscous acid-dominated GAGs are produced on, or from areas leading to, exposed surfaces and function in particle transport against the prevailing current flow, e.g. nonselective transport of pseudofaeces or in gill ventral grooves; lower-viscosity mixed GAGs are produced where transport is on, or from areas leading to, an enclosed surface and is with the current flow, e.g. transport of particles for ingestion; low-viscosity neutral GAGs occur where a reduced viscosity function is important, e.g. on labial palps where food particles are extracted from mucus prior to sorting. The mucus of cephalopods has received relatively little attention, perhaps because it is not regarded as functionally important as in other molluscan classes (see Packard, 1988, for information). Quantitative studies of mucus production have rarely been undertaken for their own sake (but see Denny, 1980a; Culley and Sherman, 1985; Davies et al., 1990a, 1992b; Davies, 1993; Davies and Williams, 1995, 1997; Kapeleta et al., 1996). They are more usually made incidentally in studies where the main aim is something else (e.g. biochemical study, Wilson, 1968;

16

M.S. DAVIES AND S.J. HAWKINS

Figure 4 Light micrographs of labial palp ridges of Mytifus edulis. (1) Anteroposterior section stained with alcian blue-PAS showing general organization and locations of acid-dominant secretion mucocytes (ADS). AF = anterior fold, P F = posterior fold, VT = vesicular connective tissue. (2) Antero-posterior section stained with modified Masson trichrome. C = cilia, RS = ridged surface, SS = smooth

MUCUS FROM MARINE MOLLUSCS

17

tenacity, Branch and Marsh, 1978) or in the construction of energy budgets for species in which mucus is thought to play an important role (e.g. Edwards and Welsh, 1982, Horn, 1986, Peck et al., 1987). The only mucuses whose production have been quantified are those secreted by the pedal sole and hypobranchial gland, those which emerge with the faeces and those involved in bivalve feeding. A large proportion of bivalve mucus is lost to the environment as pseudofaeces (Foster-Smith, 1975; Owen and McCrae, 1976; Bayne and Newell, 1983), most of which has the large bivalve gill as its ultimate origin (Morton, 1983, Beninger et al., 1993). Although attempts have been made to measure this loss, none has been totally satisfactory. Deslous-Paoli et al. (1983), Heral et al. (1983) and Deslous-Paoli and HCral (1984) used the technique of Sornin et al. (1983) to measure the “biodeposits” (faeces and pseudofaeces) produced by oysters. Sornin et al. used jars suspended under tables of oysters to collect their biodeposits. The biodeposits were then analysed for photopigments and organic C and N. Deslous-Paoli et al. (1992) placed Crassostrea gigas “on a division” such that pseudofaeces and faeces could be collected by pipette on either side. Organic and mineral contents of the biodeposits were determined. Walz (1978) estimated organic material in faeces and pseudofaeces by dispersing them with an airlift as they were produced by Dreissena polymorpha. He then took a sample of the medium (assuming its organic component to be uniformly suspended) and determined its C content using a C-H analyser. K i ~ r b o eet al. (1980) collected pseudofaeces in a sedimentation chamber attached via a tube to the inhalant siphon of Mytilus edulis. Pseudofaecal carbon content was determined by wet oxidation. Aiello et al. (1988) quantified mucus production of M . edulis gill by staining the collected mucus string using alcian blue and then reading in a spectrophotometer the concentration of free dye in a suspension of dissociated mucus formed by treatment with dioctyl ester of sulphosuccinic acid. Mean values ranging from 1.8-6.0mg mucin (g gill wet weight)-’ were Figure 4 (continued) surface, T = trough. (3) Detail of a single ridge (modified Masson trichrome stain). Note transition from pseudostratified to columnar ciliated epithelium in region of trough. BL = basal lamella, CE = ciliated epithelium, MF = muscle fibres, VC = vesicular cells. (4) Detail of (3) showing abundant nuclei (N) indicative of cell and ciliation density in crest of ridge epithelium. (5) Profile of ridges (alcian blue-PAS stain) showing epithelial location of neutral secretion mucocytes (NS and long arrows) and subepithelial location of acid-dominant mucocytes (arrowheads). Note ducts (D and short arrows) in overlying ciliated epithelium. AD = apical depression. (Reproduced from Beninger, P.G., St-Jean, S.D. and Poussart, Y. (1995). Labial palps of the blue mussel Mytilus edulis (Bivalvia: Mytilidae). Marine Biology 123, 293-303, Fig. 3, p. 296. Copyright Springer-Verlag, Heidelberg. By kind permission of the authors and Springer-Verlag.)

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M.S. DAVIES AND S.J. HAWKINS

recorded. Aiello et al. also recorded the velocity (means -0.5 mm s-’) and width (means from 24-56 pm) of the mucus string in the ventral food groove and concluded, by observing the effects of neurotransmitters, that at least the secretion from frontal mucocytes is under nervous control. The above techniques, particularly the latter, could be applied to estimate mucus loss by feeding an inorganic diet. Until this is done the results regarding mucus will remain difficult to interpret. None of the above studies was undertaken specifically to determine mucus production. Nevertheless, given the amount of pseudofaecal release (see, e.g. Bayne and Newell, 1983, Htral et al., 1983), the amount of mucus in pseudofaeces is likely to be energetically costly. Prins and Smaal (1989) reported net losses of carbon by M. edulis, > 40 mg I-’ of water pumped, attributed to loss of mucus bound in pseudofaeces, described by Hawkins and Bayne (1992) as “very significant indirect costs of filtration in turbid environments”. Hypobranchial mucus secretion was measured by Kideys and Hartnoll (1991). They filtered the seawater taken from tanks in which individual Buccinum undatum had been placed for 20-30 min. The residue (assumed to represent hypobranchial mucus) was produced by animals at the rate of 2-100 mg dry weight h-’, depending on animal size and had a calorific value of 8.76 kJ g-I. Lubricatory mucus may be secreted into the gut of molluscs to ease the passage of food (see Bayne and Hawkins, 1990) and will emerge with the faeces. Some vermetid molluscs use mucus produced in the foregut to transport food particles into the stomach (Morton, 1951). Beninger and Le Pennec (1993) observed mucus glands in the bucco-oesophageal region of Mytilus edulis, the mucus produced presumably aiding transport through the gut. They suggested the strong amylase activity of the digestive gland may be responsible for preventing accumulation of mucus in the gut, thus easing digestive processes. Quantification of intestinal mucus was achieved by Calow (1974a), Kofoed (1975) and Edwards and Welsh (1982). Calow and Kofoed estimated gut mucus using a radiotracer technique. Calow estimated mucus to represent up to 20% of the faecal organics of the freshwater snails Ancylus Juviatilis and Planorbis contortus. Kofoed reported & 19% depending on food for the mud snail Hydrobia ventrosa. Edwards and Welsh allowed the mud snail Zlyanassa obsoleta to feed on an organic-free sediment, then transferred the snails to a petri-dish containing seawater to void faeces. Faecal pellets were collected by pipette and transferred to precombusted filters for drying and ashing to determine the organic component of the faeces. Assuming all the ash-free matter was mucus (which is incorrect as other organics are secreted into the gut, see Goddard and Hoggett, 1982), Edwards and Welsh found mucus to account for 8% of the organics in the faeces from a “normal” diet. A similar technique was employed on ascidians by Goddard and Hoggett (1982). They fed their animals diatomaceous earth

MUCUS FROM MARINE MOLLUSCS

19

then dissected out each part of the gut and examined the contents for organic C and N. The quantification of pedal mucus from gastropods has been achieved in two ways. Either the mucus adhering to the foot is scraped off or the mucus is collected from the animal’s pedal trail. The former method was employed by Calow (1974b) who collected -5mg dry weight of mucus from 50 Lymnaea stugnalis. Calow’s technique was based on that of Wilson (1968) who removed the mouth of the shell of L. truncatula and “gently stroked” the head-foot with the bent end of a fine glass rod to remove 200-800 pg of hydrated mucus per snail. Calow also determined a calorific value for the mucus (24 kJ g-’). A similar collecting method was employed by Branch and Marsh (1978) who ran a “clean blade” over the foot of South African patellids to collect mean values of 0.50-1 1.47 mg hydrated mucus cm-2 of pedal sole. Horn (1986) used a glass rod to collect mucus from the foot of the chiton Chiton pelliserpentis after the animal had been adhering to a glass slide for 8 h. The collected mucus was then transferred to the slide for rinsing, drying and weighing. Kapeleta et al. (1996) collected the pedal (and dorsal) mucus from five species of terrestrial slug by moistening the epidermis in a high K + Ringer solution and then drawing a stainless steel spatula over the epidermis until mucus, in granule form, appeared as a “milky white fluid”. Kapeleta et al. calculated the mean number of granules contributing to the slugs’ trails as 0.58 x 106cm-2. Davies et ul. (1990a) and Davies (1993) measured mucus production of limpets by allowing animals to adhere to and move on glass plates for 6 h. At the end of this period, mucus was wiped from the sole using the rounded end of a pair of forceps and collected from the plate using a razor blade. A calorific value of 8.9 kJ g-’ was also determined. Mucus production rates should be regarded with caution as their magnitude can depend on when and from where animals were taken and on the environmental conditions at collection. Davies et al. (1990a) found pedal mucus production by Patella vulgata to be -230 pg dry weight h-’ for a limpet of 1g whole dry weight in saturated air. In seawater production was -70% of the rate in saturated air and -40% in air at 70% relative humidity (RH). Those animals in the low RH treatment did not move and so mucus production increased when animals were active. By assuming that limpets spend 12hd-’ in a low RH environment (stationary) and 12hd-’ in seawater (active), Davies et al. calculated that for a population (12.6 m-2) active whilst the tide is in during daylight (Hartnoll and Wright, 1977) at Derbyhaven, Isle of Man, pedal mucus production was 80 g m-2 year-’. Calculations such as these, based on many assumptions, should, of course, be regarded as very approximate. By allowing winkles to crawl over preweighed glass plates and estimating mucus production by subtraction, Davies et al. (1992b) similarly found that Littorinu littoreu produced

20

M.S. DAVIES AND S.J. HAWKINS

-35% less mucus in seawater than in saturated air (mucus production rate = 0.6 pg dry weight mm-' travelled for an animal of 0.5 g whole dry weight, energy content of mucus = 19kJg-') (Figure 5). Grimm-Jnrrgensen et ul. (1986) observed a similar phenomenon in the freshwater snail Fossuriu modicellu. Davies et al. (1990a, 1992b) suggested the phenomenon might be owing to the increased weight of emersed animals over immersed animals or to the lack of a diluent for mucus in air (the mucus may not be fully hydrated when secreted). For intertidal species, foraging when the tide is in will save energy as mucus production. This may not necessarily correlate with an increase in fitness, since foraging when the tide is in may have other energetic costs or be risky owing to the threat of dislodgement or predation. Nevertheless, mucus may still limit the distance over which it is profitable to search for food, especially when emersed. Smaller gastropods, which produce proportionately more mucus than larger ones (Davies, 1991; Navarro and Torrijos, 1995) may be particularly restricted, and this may possibly contribute to high juvenile mortality, particularly in limpets (Blackmore,

.s C

a

a

6

1

I

I

I

0.2

0.4

0.6

I

j 0.8

2

1.o

0

Log,, whole animal dry weight (g

Figure 5 Pedal mucus production during locomotion and associated energy loss for different-sized periwinkles, Littorina littorea, in saturated air (open circles) and in seawater (solid circles). (Redrawn after Davies et al., 1992b.)

MUCUS FROM MARINE MOLLUSCS

21

1969a; Bowman and Lewis, 1977). In one population of P . vulgata, pedal mucus production whilst immersed was found to vary by up to fivefold (from 50 to 221 pg dry weight h-' for a limpet of 1g whole dry weight), dependent on when, during a 22-month period, the animals were collected (Davies, 1993). Such variation is difficult to explain as no seasonal pattern was apparent. Mucus production also increased with shore height (a twofold increase) and with exposure to wave action (by 1 . 5 ~ )These . differences were attributed to differences in foraging behaviour since in situ foraging high on the shore and on exposed shores might be limited by exposure time (these populations feed when the tide covers them) and wave action, respectively, thus animals in the laboratory may have been moving, hence producing mucus, in opportune conditions. The plasticity of limpet behaviour, particularly that of P . vulgata, makes interpreting such information extremely difficult (see Hawkins and Hartnoll, 1983; Little, 1989). Peck et al. (1993) used a similar technique to that of Davies to quantify pedal mucus production by the Antarctic limpet Nacella concinna. They measured mucus production by allowing limpets (transported to the UK) to attach to weighed glass plates and reweighing the plates after the removal of the limpets and drying of the plates. Mucus production rates ranged from 0.49 to 1.87 mg dry mucus day-' for animals of 21.O-85.2 mg tissue AFDW; calorific value = 10.9kJg-'. Starvation of limpets for 5-7 weeks prior to experimentation reduced mucus production rates by 61%, despite locomotory activity being increased. Peck et al. also noted that mucus produced whilst attaching amounted to 80% of the mucus secreted in a 24-h period. For stationary P . vulgata in an aerial environment, the amount of mucus produced did not increase after the first lOmin of a 6-h period (Davies, 1993). These results suggest that mucus can retain its functional capability (adhesive properties) over periods of hours and that there are long periods during which the pedal secretory apparatus is inactive. Data on mucus production by tropical gastropods are notably lacking, although two studies have been performed. Davies and Williams (1995), using the technique of Davies et al. (1990a), measured pedal mucus production by stationary Cellana grata, a tropical limpet, as 3.8 mg dry weight over a 30-min period for an animal of 1g dry flesh weight (mucus calorific value = 10.8 kJ g-'). Locomotory production was also assessed (from 2.1 to 15.2pg dry weight mm-' for animals from 0.139-0.891 g dry flesh weight) by allowing animals to crawl up a glass plate inclined at 60" which was constantly sprayed with seawater. Mucus was scraped from the plates using a razor blade. Davies and Williams (1997) similarly recorded mucus production rates for the pulmonate limpets Siphonariajaponica, S. Sirius and S. atra. Siphonaria japonica (2.7pg dry mucus produced per millimetre moved for an animal of 0.1 g whole dry weight) produced more mucus when active than did S. atra (0.97pg dry mucus produced per milli-

22

M.S. DAVIES AND S.J. HAWKINS

metre moved for an animal of 0.1 g whole dry weight), although the mucus of S. japonica had a lower calorific value (5.6 kJ g-’) than did the pooled mucus of S. atra and S. Sirius (9.1 kJ g-I), perhaps indicating a trade-off between the quantity of mucus produced and its calorific value. Stationary S. atra and S. Sirius produced mucus at rates over a 10-min period which were not significantly different (mean = 55 pg dry weight for an animal of 0.1 g whole dry weight). Several workers have estimated production of mucus by examining trails. Calow (1972) scraped the mucus trails of Ancylus fluviatilis and Planorbis contortus from the sides of 2 1 vessels. He then isolated the mucus by filtration through a 0.8 pm pore filter and weighed the residue (mean production = 0.21 mg dry weight d-’ for 50 A.fluviatilis and 0.18 mg dry weight day-’ for 50 P. contortus). A similar method was employed by Gonzalez et al. (1990) in determining pedal mucus production by Concholepas concholepas. Navarro and Torrijos (1995) scraped pedal mucus from the walls of glass vessels in which C. concholepas had been crawling. Pedal mucus was produced at a rate of 0.7mg dry weight h-’ for an animal of log dry flesh weight and was converted to a calorific value (6.9 kJ g-’) using data from Gonzalez et al. (1990) for this species. Kideys (1991) used a similar technique, scraping the pedal mucus produced by the whelk Buccinum undatum from the sides of tanks. Mucus was produced at a rate of -2 mg dry weight h-’ for an animal of log dry whole weight at 8°C. Mucus production at higher temperatures (up to 15OC) was slightly greater, but generally pedal mucus was produced at only half the rate of hypobrancial mucus in this species. Iwasaki (1992) collected both trail and “stationary” mucus from the limpet Cellana toreuma by scraping mucus from glass slides to which the animals had attached or moved over in running seawater in order to determine protein content (see Section 2). Richardson (1975) allowed the terrestrial snail Cepaea nemoralis to crawl over a preweighed aluminium mesh and then dried, weighed and determined the calorific value of the resulting mucus film (14kJg-’) (mean production rate = 1.61mg dry mucus individual-’ d-’ ). Peduzzi and Herndl (1991) used a similar technique for Patella caerulea and Monodonta turbinata, replacing aluminium foil each hour, over 24 h This technique may have overestimated mucus production: for “average sized” P. caerulea, 12.27mg dry mucus individual-’ d-’; M. turbinata, 5.58 mg dry mucus individual-’ d-’. Denny (1980a) measured the length and the thickness (10-20pm) of the mucus trail of Ariolimax columbianus using the method of Lissman (1945) which involves freezing the mucus in absolute alcohol. Denny and Gosline (1980) collected A . columbianus pedal mucus by allowing the slug to crawl around a rotating glass rod. This technique produced 0.1-0.3 ml of mucus from each slug. Edwards and Welsh (1982) allowed Ilyanassa obsoleta to crawl across a combusted sediment in filtered seawater. The mucus trail produced was then dried and

MUCUS FROM MARINE MOLLUSCS

23

ashed before weighing (mean weight = 21.8 pgcm-2 of trail). Siddall(l984) collected the trail mucus of the conch Strombus gigas after groups at different densities were allowed to crawl for 7.3 h in a crystallizing dish filled with filtered seawater. The mucus was removed and dry weighed after being stabilized (using cetylpridimium chloride) to ease handling and stained with Evans blue. Siddall found that the rate of mucus production per centimetre travelled increased significantly at high population densities, although no explanation for this was proposed. Culley and Sherman (1985) and Peck et al. (1987) both used the same technique for estimating trail mucus production by Huliotis tuberculuta, the former to determine the effect of substratum particle size on mucus production (an exponential relationship of the form y = ax b ,where y = mucus production rate and x = substratum particle size) and the latter in the construction of an energy budget. In each case ormers (abalones) were placed in large preweighed crystallizing dishes filled with filtered seawater. After 24 h the animals were removed and the dishes rinsed in distilled water to remove seawater and faeces before drying to constant weight. Mucus production values were obtained by subtraction. Peck et al. (1987) removed some of the mucus and burned it in a calorimeter to yield an energy value (23 kJ g-'). The quantification of mucus production from gastropod epithelial surfaces other than the pedal sole, hypobranchial gland and gut has been neglected but these could prove to be sites of considerable mucus production. Much variation is evident in the quantities of mucus produced by different molluscs, although they are usually of a considerable amount. However, comparison is hindered by the non-uniformity in the units in which mucus production is expressed.

5. FUNCTIONS OF MUCUS

The uses to which mucus is put are varied. Mucus is a highly complex product whose uses far outstrip that of a simple friction-reducer. Nevertheless it is an energy-rich product (see Section 6) and so is likely to be a product which is conserved. Here we list the broad functions of use to the producer of the mucus. 5.1. Locomotion

The importance of mucus in molluscan locomotion has been recognized since Barr (1926) found Milax sowerbii was incapable of movement once its pedal gland had been cauterized. Denny (1980b) reported that owing to

24

M.S. DAVIES AND S.J. HAWKINS

mucus production the cost of locomotion in crawling molluscs is greater than locomotion in any other animal. In gastropods, mucus, usually from the anterior pedal gland (see Gainey, 1976; Grenon and Walker, 1978; Shirbhate and Cook, 1987), secreted as a thin layer (10-20pm, Denny, 1980a) between gastropod foot and the substratum can function in locomotion in two ways. It can act as a medium in which pedal cilia beat to propel the animal (Miller, 1974), for example, in the sediment dwellers Hydrobia ulvae and Cassis ruberosa (Hughes, 1986). The mucus can alternatively act to couple the force exerted by the pedal muscles to the substratum for those groups (e.g. patellids, littorinids) which require a firm grip on the substratum (Miller, 1974; Branch and Marsh, 1978). The mucus acts as a sticky adhesive which extends the range of these groups to vertical faces and overhangs (see Branch and Marsh, 1978; Grenon and Walker, 1978) and is important in reducing predation by seabirds which attempt to prise marine gastropods and polyplacophorans from their substrata (Hahn and Denny, 1989). The way in which this coupling facilitates both locomotion and adhesion is described by various authors (e.g. Jones and Trueman, 1970; Miller, 1974; Branch, 1981; Denny, 1981, 1989). The properties of mucus which allow this form of locomotion are described by Denny (1983) (see Section 3). Whatever the mechanics of molluscan crawling, mucus is lost from its producer and is laid down as a characteristic trail. It may be thought that the work of Barr (1926) demonstrated the necessity of mucus for gastropod locomotion. However, lack of locomotion in Milax sowerbii may have been a stress response and although the mechanism of locomotion and the production of mucus have undoubtedly co-evolved, many creeping organisms (e.g. annelids) can propel themselves using muscular waves without the need for the production of a mucus. It is our contention therefore that the employment of mucus in locomotion has been successful, since it allows organisms to occupy verticals and overhangs which hitherto were inaccessible habitats. Thus the adhesive properties of mucus are an essential part of the locomotory mechanisms of many molluscs, and the mechanisms may have adapted to accommodate those adhesive properties. The freshwater bivalve Corbiculafluminae has been shown to use mucus in locomotion by extruding mucus from its inhalant siphon until the hydrodynamic drag generated is sufficient to “balloon” the animal downstream (Prezant and Chalermwat, 1984). Reports of similar behaviour in marine molluscs are common. Microscopic mucus threads are exuded, usually in combination with byssus threads, to enable a brief planktonic existence for numerous post-larval bivalves including Mytilus edulis (Sigurdsson et al., 1976; Lane et al., 1985; Beukema and de Was, 1989) and Cerastoderma edule (de Montaudouin, 1997). Martel and Chia (1991a) confirmed empirically that the gastropods Barleeia spp., Tricoliapulloides and juveniles of Nucella emarginata and Littorina sitkana all exhibit drifting using a mucus thread or

MUCUS FROM MARINE MOLLUSCS

25

threads. The phenomenon has been studied in more detail in gastropods of the genus Lacuna which secrete pedal mucus and raise the sole of the foot to release the mucus produced which then extends to between 50 and 160 x the shell length, depending on animal size. The mucus then acts as a drogue to disperse the animals and has been observed to aid settlement by becoming entangled in algae. The mucus thread reduces the sinking rate of the gastropods by up to four times (Martel and Chia, 1991b). Such post-metamorphic drifting may aid dispersal and enable individuals to escape from an unfavourable habitat: Abelson et al. (1994) give a good account of the hydrodynamics involved in mucus-thread transport and its implications for settlement. Indeed Martel and Diefenbach (1993) observed higher rates of drifting behaviour (3-5x) in juvenile Lacuna vincta and L. variegata maintained in an unfavourable environment (without macroalgae) as compared to a favourable environment (with macroalgae). Octopod larvae of several species from Hawaii are thought to use a similar mucus drogue, secreted from integumental pores, for dispersal (Young et al., 1989).

5.2. Mucus Trails

The complex topography of gastropod habitats, where visual tracking may be difficult, has been suggested to explain the use of pedal mucus trails (Figure 6) as an aid to navigation (Denny, 1989). Trail-following behaviour

Figure 6 Mucus trails of the dogwhelk Nucella lapillus, visible owing to adhesion of fine sedimentary particles.

26

M.S. DAVIES AND S.J. HAWKINS

may also have evolved at least in part as a means of recycling energy-rich mucus (Connor, 1986, Davies et al., 1992a) and a mechanism of producing efficient foraging by speeding aggregation on patchy resources (see Hawkins and Hartnoll, 1983; Deneubourg et al., 1988). Individuals may follow their own mucus trails to home (Funke, 1968; Cook et al., 1969; Cook, 1971; Cook and Cook, 1975; Cook, 1977; McFarlane, 1980, 1981; Chelazzi et al., 1983, 1985, 1987; Chelazzi, 1990; Della Santina, 1994; see Cook, 1979 for review of earlier work), those of conspecifics to aggregate (perhaps for protection), and often to mate (Moulton, 1962; Breen, 1973; Townsend, 1974; Lowe and Turner, 1976; Trott, 1978; Trott and Dimock, 1978; Hirano and Inaba, 1980; Bretz and Dimock, 1983; Cook, 1985; Wareing, 1986; Chelazzi et al., 1985, 1988; Branch and Barkai, 1987; Chelazzi, 1990; Cook, 1992; Erlandsson and Kostylev, 1995) and those of other species to find prey (Paine, 1963; Gonor, 1965; Blair and Seapy, 1972; Murray and Lewis, 1974; Cook, 1985; Carte and Faulkner, 1986; Cimino and Sodano, 1989). The locomotory force applied is reduced in trail-following (“tracker”) snails in comparison to “marker” snails (Tankersley, 1989). Thus trail following may reduce the metabolic energy required for locomotion (Hall, 1973), although the energy cost of mucus production in locomotion has been found to be 35x that of the metabolic energy cost of locomotion (Davies et al., 1992b) and Cook (1992) found that the quantity of mucus produced by Limax pseudojlavus (terrestrial slug) in following its own mucus trail was the same as for the original deposition of mucus. The speed of Zlyanassa obsoleta is unchanged when it traverses conspecific mucus trails (up to four trails thick) in comparison to a substratum of sand (Dimock, 1985) but Littorina irrorata increases its speed over mucus trails in comparison to sand (Hall, 1973). L. littorea also increases its speed over trails in comparison to glass in both mating and non-mating seasons (Erlandsson and Kostylev, 1995), although this is not owing to the presence of mucus directly facilitating a faster locomotion. Since mucus trails may both stabilize the substratum and produce a smoother surface over which to move, some species may reduce the amount of mucus produced whilst trail-following. Surface topography has been shown to determine the amount of mucus required for locomotion (Culley and Sherman, 1985). Further studies are necessary to determine whether gastropods can save energy in this way. The mechanism by which gastropods can detect the presence and polarity of a mucus trail is, however, poorly understood. The presence of a physical cue in the mucus has been suggested by Simkiss and Wilbur (1977), Bretz and Dimock (1983) and Stirling and Hamilton (1986) and the filaments found in mucus trails (see Section 4) may provide this cue. Both Simkiss and Wilbur (1977) and Bretz and Dimock (1983) noted that the filaments were frayed at their ends furthest from the animal laying the trail. The presence of a chemical cue in the mucus, perhaps as a volatile chemical

MUCUS FROM MARINE MOLLUSCS

27

whose concentration can indicate the age - and perhaps direction - of the trail, has been suggested by Hall (1973), Gilley and Swenson (1978), Bousfield et al. (1981) and Raftery (1983). Stirling and Hamilton (1986) showed that a polarity cue persisted in the mucus trails of Littorina irrorata for at least 60 min in air. Calculations by Denny (1989) showed that for a volatile substance to be involved in the detection mechanism, the snails must detect concentration gradients less than M mm-l, which seems unlikely. However, an initial testing of the trail over distances of several centimetres (as is done by L. irrorata, Hall, 1973) would greatly reduce the need for such sensitivity. Robbins and Hamilton (1996) have shown that trailfollowing and detection of trail polarity by L. irrorata is dependent on the presence of intact cephalic tentacles. 5.3. Feeding

Probably the most extensively studied molluscan feeding system is the filtering mechanism in bivalves. Here food particles are trapped by the gills, bound in mucus and transported along ciliated tracts to the labial palps for sorting and eventually to the mouth (Yonge, 1949; Bayne et al., 1976; Morton, 1983, see Purchon, 1977 for review). The role of mucus in this process was only inferred until endoscopy revealed that mucus was important in intact, as opposed to opened and thus stressed animals (Beninger et al., 1991,1992 and Tankersley and Dimock, 1993). Ward et al. (1993) and Beninger et al. (1993) examined particle transport in situ by endoscopy and confirmed that mucociliary action (the propulsion of particles bound in mucus by cilia) transports filtered material to the ventral ctenidial margin and thence along ciliated grooves. Although mucus plays a critical role in this process, the finer points of its function are still in the process of being understood, and are likely to be clarified using the endoscope as a tool combined with histology or confocal microscopy. Some of this mucus is ingested with the food, indeed Beninger and Le Pennec (1993) found evidence that particles suspended in water alone are not ingested. Mucus is also employed by bivalves to cleanse the mantle cavity of uningested particles. These pseudofaeces are bound in mucus and ejected through the inhalant aperture (Yonge, 1926,1949; Barnes, 1980; Bayne and Newell, 1983, Ward et al., 1993). Beninger et al. (1995) concluded that since pseudofaeces are not produced from the ventral particle groove of the gill (Beninger et al., 1993; Ward et al., 1993), pseudofaecal mucus must originate on the labial palps, which certainly contain enough mucocytes for this to be true (Beninger et al., 1995; Beninger and St-Jean, 1997a; Beninger et al., 1997a). Beninger et al. (1995) noted in Mytilus edulis that while the palps were in an inclined position mainly acid-dominant secretions will be discharged but while in an erect

28

M.S. DAVIES AND S.J. HAWKINS

position mainly neutral secretions will be discharged, and this was presented as evidence for the function of the palps in particle selection. Using Crassostrea virginica fed on Tetraselmis suecica, Newell and Jordan (1983) showed that energy, carbon and nitrogen in the pseudofaeces were all reduced compared to that in the food and their results suggested that this oyster can select particles of differing types for ingestion and rejection. Newell and Jordan hypothesized that this is achieved by a chemosensory mechanism acting on particles which have been freed from mucus. This may be achieved by the lowering of the viscosity of the mucus through the beating of labial cilia (Ward et al., 1994). This hypothesis has yet to be confirmed, but future endoscopy may provide the necessary information. Beninger et al. (1997b) give evidence that food particles are transported in mucus “rafts” (10-25 pm thick) which are propelled by the tips of cilia and which sit atop the periciliary space that contains a medium less viscous than the mucus but more viscous than the surrounding water, and suggest a twolayer model of mucociliary transport. For an earlier and fuller discussion of the propulsion of mucus by cilia see Sleigh (1989). Given the amount of pseudofaecal release (see, e.g. Bayne and Newell, 1983; Hbral et al., 1983), the amount of mucus in pseudofaeces is likely to be energetically costly. Bivalves such as Macoma balthica (Taghon, 1982) may also use mucus whilst deposit feeding. It is presumably secreted by the inhalant siphon, and preferentially adheres to organic, rather than inorganic particles in the sediment, whereby the former are selected for ingestion. Connor and Quinn (1984) and Connor (1986) suggested that pedal mucus might serve a nutritional role for territorial grazing gastropods and is ingested after acting as a trap and possible fertilizer for food. These authors showed that the growth of microalgae (a component of the diet) was enhanced in the presence of the pedal mucus of the homing limpets Lottia gigantea and Collisella (Macclintockia) scabra, but not in the presence of mucus from the non-homing limpet Collisella (Tectura) digitalis, nor that of the carnivore Nucella emarginata. They also showed that trail mucus persisted from 4 to 15d, depending on species and position on the shore, and suggested that this time was sufficient to allow for ingestion of mucus. Davies et al. (1992a) showed that the mucus produced by Patella vulgatu persisted for up to 80d (Figure 7) and could similarly trap microalgae, especially diatoms (Figure 8), particularly in the first 24 h of exposure. Davies and Williams (1995), however, found that the pedal mucus of the tropical limpet Cellana grata persisted for only 6 d and the mucus trapped microalgae (with a peak at 4 d of exposure), but only marginally better than bare surfaces. Santelices and Bobadilla (1996) demonstrated that the pedal mucus of Chilean gastropods trapped microalgal particles in the laboratory and microalgal and macroalgal particles in the field, again showing that more algal material was collected after a shorter exposure period (1 h)

29

MUCUS FROM MARINE MOLLUSCS

100

Littorina littorea

80

60 40 n

E 20 at c, c en

. I

f0 Very sheltered Moderately sheltered Semi-exposed

oa

++

E E

. I

-Ei

t

-*-

Very sheltered Moderately sheltered Semiexposed

-+-

..-.."

3E

P.

s 100

v)

I

6 80 60

Patella vulgata

40 20 0

0

4

8

12

16

20

24

Time (days) Figure 7 Onshore (Isle of Man) persistence of pedal mucus from the periwinkle Littorina littorea and the limpet Patella vulgata at three sites over two dates. Persistence is recorded by allowing animals to crawl over a Perspex (Plexiglass) plate that is marked into cells. The plate is then dipped into a suspension of graphite and the graphite adheres to the mucus. The plate is then placed onshore and monitored as graphite is removed from the cells (presumably along with mucus) by wave action. Mucus persists for the order of weeks: limpet mucus for about twice as long as periwinkle mucus. The rapid loss of mucus from one plate for each species was attributed to its position in a microhabitat prone to siltation. (Redrawn after Davies et al., 1992a.)

30

M.S. DAVIES AND S.J. HAWKINS

(A) JANUARY, 24h 86 -

= Control 0 F! vulgatamucus

4-

(B) APRIL, 24h

a E

4

1210-

(C) APRIL, 7 days

86-

SHORE TYPE

Figure 8 The adhesive nature of gastropod mucus. Mean number (+SE) of diatoms observed under the SEM at 300x on cellulose nitrate filter discs coated with pedal mucus from the limpet Patella vulgata and on control discs. Dates and times refer to the periods the discs were left onshore prior to SEM observation. Mucus has a clear effect in accumulating diatoms. (Redrawn after Davies et al., 1992a.)

than after longer periods (6 and 12h). Interestingly, these authors also suggested that pedal mucus might be useful as a device to trap algal spores in intertidal environments with the aim of documenting changes of spore abundance in the water. No direct evidence for the ingestion of mucus trails

MUCUS FROM MARINE MOLLUSCS

31

has, however, been shown, although individual Hydrobia ulvae have been observed to ingest their own pedal mucus which is used to trap organic material as the animals float in the water between ripple marks on the beach (Fenchel et al., 1975; Newell, 1979). Given the density of molluscan grazers on many shores, the production rates of trail mucus, the persistence of the mucus and the trail-following habit of many species, it is likely that trail mucus does fulfil a provendering role, albeit perhaps serendipitously. Pedal mucus may be a preferential food for, for example, littorinids or amphipods, in habitats where levels of nutrient-rich detritus are low. It would pay grazers to know when, after laying a trial, is the optimum time in terms of the nutritional content of the trail to ingest it. In addition, evolutionarily stable strategy theory predicts that if grazers are adding a factor to the trail to promote attachment of, or growth of, food particles then such grazers would be territorial (Davies et al., 1992a). The evidence that grazing species have evolved to exploit their mucus trails for feeding is circumstantial and the area would benefit from further research, particularly in species which maintain “gardens” (see Branch et al., 1992). The ingestion of mucus has also been shown in the terrestrial slugs Ariolimax columbianus (Richter, 1980) and Limax maximus (Denny, 1989). Suspension feeding using a mesh of mucus (to filter) or curtain of mucus (to adhere) small food particles from the water column is common in adult (see Yonge, 1928; Morton, 1951; Walsby, 1975; Graham, 1985; Fretter and Graham, 1994) and larval gastropods (Hamner et al., 1975). Gilmer (1972) observed unsupported mucus webs of up to -2m diameter produced by planktonic opisthobranchs Gleba cordata and Corolla spectabilis, which are -50mm in length. The mucus is produced from mucus glands along the periphery of the wing plate and when entangled with food is moved, using ciliary action, into the mouth. Several sessile species use mucus nets or bags (Jerrgensen, 1966) where porous sheets of mucus are held extended into flowing water before being hauled into the mouth (e.g. Bithynia tentaculata, Schafer, 1952; Olivella columellaris, von Seilacher, 1959; Gadinalia nivea, Walsby et al., 1973; Dendropoma maximum Hughes and Lewis, 1974; Barnes, 1980; Serpulorbis squamigerus, Nelson, 1980). Other species, such as Crepidula fornicata (Orton, 1912, 1914; Barnes, 1980) and Trimusculus reticulatus (Walsby, 1975) secrete a mucus net across their inhalant aperture to catch food particles (Hughes, 1986). Umbonium species, turritellids, vermetids and some capulids (Orton, 1912; Yonge, 1938) capture food particles on their gills (see Hughes, 1986) before the food is transported to the mouth in mucus strings. Many naticid gastropods also use mucus for feeding purposes, coating their prey in mucus to prevent its escape (Kohn, 1983). During the capture of Umbonium vestiarium by Natica gualteriana the latter secretes (presumably from the foot) mucus into the former’s aperture (Savazzi and Reyment,

32

M.S. DAVIES AND S.J. HAWKINS

1989). This action appears to have no detrimental effect on U.vestiarium save that it prevents it from emerging from its shell and thus is quiescent while the naticid drills into it. Savazzi and Reyment concluded that that the mucus acts in a chemical, rather than physical manner in preventing the emergence of the prey. Mucus is involved in the drilling activity of numerous gastropods (Carriker, 1981), but its role is uncertain. Mucus is also used by deposit-feeding scaphopods. The secretions of gland cells located at the head of captacula were thought to be involved in prey capture as a toxin (Morton, 1959) and adhesion (Shimek, 1988), but are now thought to aid the passage of food particles along ciliated tracts (Byrum and Ruppert, 1994) in a fashion similar to that in bivalves. 5.4. Protection

The use of mucus to isolate an animal from its environment, or to actively counter some facet of environment is common in marine Mollusca and it is perhaps here that mucus is most diverse in function. Mucus is produced from most molluscan epithelia (Simkiss and Wilbur, 1977), acts as a barrier to diffusion (Grimm-Jnrrgensen et al., 1986) and may function in selective ion transport (Ahn et al., 1988). Gastropod epithelial mucus is often a first line of defence and has been shown to reduce exposure to physical stress and predation. Wilson (1929), Bingham (1972), Morris et al. (1980), Denny (1984), McMahon and Britton (1985), Britton (1995) and Davies and Hawkins (pers obs.) have noted the habit of littorinids and amphissids of attaching themselves to vertical rock using a strand of pedal mucus as a glue between substratum and shell. The animal’s head is then retracted behind the operculum. This behaviour is thought to render the animal less susceptible to desiccation and overheating. By secreting a veil of mucus (which then dries to form a wall) between shell and substratum, the limpets Acmaea (Tectura) digitalis, A. (Macclintockia) scabra and A. persona can reduce desiccation stress (Wolcott, 1973). A similar phenomenon occurs in terrestrial snails which secrete a CaC03/ mucus matrix across the shell aperture which dries to form a water-tight seal (e.g. Helix aspersa, Otala lactea, Sphincterochila boisseri, Machin, 1967; Schmidt-Nielsen et al., 1971). Mucus has also been implicated in protecting antarctic limpets from extreme cold (Hargens and Shabica, 1973). The role of mucus in fish as a barrier to pollution has been extensively studied (Shephard, 1994), but in molluscs this has not been directly assessed. However, excess mucus production by bivalves after exposure to heavy metals (Lakshmanan and Nambisan, 1985; Moraes and Silva, 1995; Sunila, 1987; Hietanen et al., 1988; Sze and Lee, 1995) and hydrocarbons (Axiak and George, 1987) has been reported; and excess pedal (Mills et al.,

MUCUS FROM MARINE MOLLUSCS

33

1990) and intestinal (Triebskorn, 1989; Triebskorn and Ebert, 1989) mucus production by slugs in response to metaldehyde (a molluscicide) has been observed. Davies (1992) described a reduction in pedal mucus production in limpets, Patella vulgata, exposed to single heavy metals, although this reduction was probably owing to an accompanying lack of activity. Mucus can function in predator avoidance by rendering the gastropod distasteful and/ or toxic (e.g. the dorsal secretions of Doriopsilla albopunctata, Reel and Fuhrman, 1981 and Phyllidia varricosa, Johannes, 1963; the hypobranchial mucus of Calliostoma canaliculatum (pers. comm., N. Smaby); the secretions of the mantle edge in Cellana spp., Branch and Branch, 1980); by anaesthetizing the predator (Trimusculus reticulatus, Rice, 1985); by fouling the predator’s feeding apparatus (e.g. Ariolimax columbianus, Richter, 1980); or by making the animal too slippery to handle (e.g. Calliostoma species, Sellers, 1977; Harrold, 1982). Handling of molluscs can also induce copious mucus secretion (e.g. Buccinum undatum, Strombus gigas pers. obs.), as can tissue disruption upon dissection (e.g. Milax sowerbii, Barr, 1926; Archidoris pseudoargus, McCance and Masters, 1937). Interestingly, some nudibranchs are able to sequester poisons from their food which subsequently emerge in their mucus, providing a defence for these molluscs which cannot retreat into a shell. Indeed the evolution of loss of shell in this group may well have coincided with the ability to use mucus in a defensive capacity. Poisons or deterrents will probably all emerge with mucus, but few studies have specifically recognized this. Examples include: Avila et al. (1991) who observed that Hypselodoris webbi secretes the allomone longifolin (an ichthyodeterrent) in its dorsal mucus, the allomone originating in the sponge Dysidea fragilis; Paul et al. (1990) who observed that Nembrotha spp. secrete tambjamines (ichthyodeterrents from the ascidian Atapozoa sp.) in their mucus; Gustafson and Andersen (1985) who discovered terpenoids from sponges, bryozoans and coelenterates in the mucus of Archidoris montereyensis and Anisodoris nobilis. The African land snail Achatinu fulica produces an agglutinin (lectin) in its mucus (Iguchi et al., 1985) which is a 70 000 MW glycoprotein (Mitra et al., 1988). Lectin activity has also been reported for the mucus of the terrestrial gastropods Arion empiricorium (Habets et al., 1979), Helix aspersa (Fountain and Campbell, 1984; Fountain, 1985) and Archachatina marginuta (Okotore and Nwakanma, 1986). Astley and Ratcliffe (1989) examined the mucus of some species of marine mollusc but could find no lectins, although they were present in the epithelial mucus of Loligo vulgaris (Marthy, 1974). Whilst these discoveries present numerous potential roles for mucus (slug mucus is apparently used in some human therapy, Habets et al., 1979), it may be that the lectin has no function other than structural within the mucus matrix (Fountain, 1982). McDade and Tripp (1967) recorded the presence of lysozyme in oyster mucus and hypothesized that

34

M.S. DAVIES AND S.J. HAWKINS

this formed an antimicrobial defence, although lysozyme may merely prolong the functional life of the mucus by slowing down bacterial breakdown. Kubota et al. (1985) purified a glycoprotein (“achatin”, 140 000 MW) from the pedal mucus of Achatina fulica which showed no lysozyme activity but did kill both gram-positive and gram-negative bacteria by acting on cytoplasmic membranes (Otsuka-Fuchino et al., 1992). The use of mucus as a carrier for these compounds provides an unstirred layer on the surface of the animal in which the compounds can be held and prevents them from dispersing in an aquatic environment (Denny, 1989). Bakus et al. (1986) reviewed the chemical ecology of marine organisms and whilst they rarely mentioned mucus it is likely that mucus is employed as a carrier of secretable chemicals in most of the taxa they describe. The functioning of limpet pedal mucus in limpet tenacity (to prevent dislodgement by, for example, predation) is a subject of debate. Smith (1991, 1992) concludes that a glue-like adhesion is responsible for the great tenacities observed in limpets by Grenon and Walker (1981) and Denny (1984), although whether this glue is related to, or is part of, mucus is not clear. It may be that these tenacities are is not a product of a glue or mucus at all, but are owing to a very flexible muscular foot which can effectively mirror, and hence grip, the microscale contours of substrata. This would explain the way in which limpets can increase their tenacity when disturbed. Davies and Case (1997) who studied the tenacity of two littorinid species, concluded that “muscular grip” does not play a role in adhesion. They suggested that the mechanism of adhesion in these animals involves mucus. Grenon and Walker (1982) measured the thickness (3 pm) of the mucus layer under the foot of Patella vulgata after it had been attaching to an alga for 2d. Grenon and Walker suggested that adhesion was afforded by the thinness of the layer (cf. Branch and Marsh, 1978) caused by the slow uptake of water and mucus from the sole of the foot by epithelial cells, a mechanism proposed by Zylstra (1972) and Machin (1975). Cephalopods also use mucus in escape from predators. Their “ink”, squirted at predators to confuse them is bound with mucus to prevent its rapid dispersion in water (Denny, 1989). 5.5. Other Functions

Mucus, by virtue of its properties, is often thought of as a lubricant (Simkiss and Wilbur, 1977; Simkiss, 1988), but a more careful examination of its viscoelastic properties (e.g. Denny, 1983; see Section 3) would suggest that this is not the case and if anything is sticky, acting as a glue to retard phenomena rather than promoting their rate of activity. However, mucus may act protectively to ease the passage of particles through the gut of

35

MUCUS FROM MARINE MOLLUSCS

gastropod (Edwards and Welsh, 1982) and bivalve (Bayne and Hawkins, 1990) molluscs. Some bullomorph opisthobranchs move through sediments by ciliary action in a tube of mucus, but the mucus can also act both as a lanyard should the slug fall from its tube and as a camouflage as it collects small particles owing to its adhesive nature (Rudman, 1971). The Silurian burrowing bivalve Nionia prisca was thought to use mucus to line an inhalant tube in the sediment (Liljedahl, 1992). Pedal mucus can also act as a buoyancy device for aquatic transportation. In pelagic prosobranchs, such as Janthina, a sticky raft of bubbles is produced which also serves as protection for eggs (Barnes, 1980). When the limpet Helcion pellucidurn is dislodged, a mucus “sail” is produced which gives the animal sufficient lift to recolonize its habitat (Vahl, 1983), a function overlapping with dispersal (see above). Lindberg and Dwyer (1983) reported patellid mucus to be important in home scar formation where it etches the substratum, forming a template for radular erosion. Theisen (1972) and Jones (1984) observed a shell-cleaning action by the foot of the bivalve Mytilus edulis and the top-shell Calliostoma zizyphinum respectively, and suggested pedal mucus is important in keeping the shell free from epibionts. Indeed, there is some evidence to suggest that this mucus - including the epibionts which its traps - is ingested (pers. obs.). This area needs further investigation since the mucus might have useful antifouling attributes. Mucus is, of course, also present in the larvae of molluscs where complex arrangements of mucocytes have been shown (e.g. Cranfield, 1973; Lane and Nott, 1975). The pedal mucus produced by bivalve larvae may play a role in locomotion and substratum selection (Cranfield, 1973; Hermans, 1983).In pteropod larvae and prosobranch veligers mucus feeding nets have been observed (see Hamner et al., 1975).

6. MUCUS IN MOLLUSCAN ENERGY BUDGETS

Physiological energetics is concerned with the study of the gains and losses of energy at the level of the individual. This can then be scaled up to the population level. Such studies are geared towards producing values for the terms in, and then balancing, the equation:

+

C = Pg iP,. R iF iU

where C = consumption, Pg = somatic growth, P, = reproductive investment, R = respiration, F = faeces and U = excretion. This is a standard International Biological Programme (IBP) equation and is a transformation of that offered by Petrusewicz (1967). The equation describes the net energy

36

M.S. DAVIES AND S.J. HAWKINS

exchange in the individual organism following the laws of thermodynamics and assuming steady-state conditions (Bayne and Newell, 1983). Carefully constructed energy budgets for populations can help to produce an understanding of energy transfer in ecosystems (ecological energetics). Such budgets constructed for intertidal grazer or carnivore species (see Wright and Hartnoll, 1981; Bayne and Newell, 1983; Hawkins and Hartnoll, 1983; Hawkins et al., 1992 for reviews) can show the importance of these species in terms of energy flow within the littoral community, although energy budgets must be interpreted with caution, since specific budgets rely on a specific set of assumptions used in compiling them. The position of mucus in the equation is a matter of disagreement. Although defining U as excreta, Bayne and Newell (1983) then included the secretion of mucus in this term. Hawkins and Hartnoll (1983) noted this error and mentioned an alternative position of mucus in the equation, as part of production. Richardson (1975), Branch (1981) and Horn (1986) argued that secreted mucus is derived from assimilated (physiologically useful) energy and is therefore a part of production (as P, or P,,,,). Hawkins and Hartnoll (1983) suggested an “9’(secretions) term to embrace urine (Su),mucus (S,), dissolved organic matter (SdJ and exuviae (S,) secretions. Peck (1983) used S to represent mucus production, preferring to separate it from both P and (I. Johannes and Satomi (1967), Calow and Fletcher (1972) and Edwards and Welsh (1982) noted that mucus production may also contribute to the F term, along with other metabolic products such as digestive enzymes. We propose that mucus be separated to its own term, M , both for correctness and to emphasize its importance in the energy budget. In doing so, mucus is freed from the other terms and so can be incorporated independently into terms such as assimilation (A = P R M). Unless stated otherwise, M should include all mucus secretions be they pedal, faecal, or from any other epithelial surface. This was the approach adopted in various studies (Peck et al., 1987; Davies et al., 1990a, 1992b; Kideys and Hartnoll, 1991), although Bayne and Hawkins (1990) and Hawkins and Bayne (1992) preferred to further divide mucus by its origins (faecal, pseudofaecal) in their analyses of bivalve energy balance. However, in most molluscan energy budgets mucus has been either ignored (e.g. Hughes, 1970, 1971; Baluyut, 1977; Bayne and Widdows, 1978; Barkai and Griffiths, 1988; Carefoot, 1989; Wilbur and Hilbish, 1989) or relegated to the U term which is then regarded as negligible or given an arbitrarily low figure (e.g. Huebner and Edwards, 1981; Wright and Hartnoll, 1981; Lucas, 1982; Hartnoll, 1983; Davis and Wilson, 1985). (See Branch, 1981; Hawkins and Hartnoll, 1983 for reviews.) The omission of mucus can arise because many workers measure C, P and R and balance the energy equation by assuming that C - (P R) represents the unassimilated ration (Mann, 1969). The lack of the incorporation of mucus is probably

+ +

+

MUCUS FROM MARINE MOLLUSCS

37

owing to an underestimation of the role of mucus in energy flow coupled with the paucity of data on its production as a result of inherent technical difficulties in its measurement; although an appreciation of its importance is not new and was commented on by Paine (1971). Numerous workers have made estimates of mucus production by difference, i.e., by measuring all other components of the energy budget and obtaining a value of mucus production by subtraction from the value of C (e.g. Mann, 1965; Paine, 1965; Kofoed, 1975). This estimate of M is prone to the experimental errors of all the other budget component estimates and deprives the budget of an internal check on its balance. Although many quantitative estimates of mucus production have been made (see Section 4), few have been incorporated into full energy budgets, and this has only been done for four molluscan (gastropod) species (Zlyanassa obsoleta, Edwards and Welsh, 1982; Chitonpelliserpentis, Horn, 1986;Haliotis tuberculata, Peck et al., 1987; Patella vulgata, Davies et al., 1990a) (Figure 9). In other cases mucus has often been expressed as a proportion of C or of assimilated energy. Richardson (1975) measured the locomotory mucus production of Cepaea nemoralis and found it to represent -12% of assimilated energy. Richardson then incorporated the value into an estimate of total somatic production (P,) and the value of mucus in the energy budget was lost. Kofoed (1975) measured locomotory mucus “excretion” in Hydrobia ventrosa, found this mucus to represent -9% of assimilated carbon, but then incorporated it in the U term, again losing the individual energy value of mucus. Calow (1972, 1974b) used the calorific value of Lymnaea stagnalis locomotory mucus to assess the energy loss of the locomotory mucus produced by Ancylus Jluviatilis and Planorbis contortus. Mucus was found to represent -9% of absorbed energy in the former species and -26% in the latter. Calow did not calculate a full energy budget. Kideys and Hartnoll (1991) calculated mucus production as a percentage of consumed energy for the whelk Buccinum undatum and found pedal mucus to represent 1 1% and hypobranchial mucus 17%. Blandenier and Perrin (1989) derived a combined value for excretion and mucus production by subtraction for the freshwater pulmonates Lymnaea peregra (47% of C) and Physa acuta (34% of C), although the calorific values are lost in calculation. Navarro and Torrijos (1995) calculated pedal mucus production, depending on animal size, to represent between 6 and 20% of absorbed energy in Concholepas concholepas, smaller animals using proportionately more energy for mucus production. However, work in which a mucus production term has been inserted into full energy budgets has demonstrated the importance of mucus in molluscan energy balance (see Fig. 9). Edwards and Welsh (1982) compiled a budget for a population of I. obsoleta and found mucus production accounted for 13 607 kJ mP2year-’. These workers measured both trail and faecal envelope mucus and calculated values of 10 083 kJ m-* year-’ and

38

M.S. DAVIES AND S.J. HAWKINS

llyanassa obsoleta

C

kJ rn' year.'

10 816

-Pg

31

pr 85

-pc

3

R 294

Chiton pelliserpentis (low shore) kJ m' year.'

C 1521

F 132

1 1 635 23

2

R 103

t

34

363

7

1

Figure 9 Energy budgets, including a mucus term, for the marine molluscs Zlyunussu obsoletu (Edwards and Welsh, 1982), Chiton pelliserpentis (Horn, 1986) Patella vulgutu (Wright and Hartnoll, 198 1; Davies, 1991) and Huliotis tuberculuta (Peck el ul., 1987). The thickness of each arrow indicates the magnitude of each

39

MUCUS FROM MARINE MOLLUSCS R 498

C

4 Patella wlgata k~ m" year.'

2227

2 808 F

Pg 8 8

P, 9 6

U 2

1MP 624

R 2

C 8

I"

F 1

(0.01 g whole dry weight)

U 0 -1

MP 2

R 1200

(50 g whole dry weight)

F 795

MP 1121

Pg 497

P, 205

U 37

term, but note change of scale for H . tuberculata. C = energy consumed, F = energy of faeces, R = metabolic energy (heat), Pg = growth production, P, = reproductive production, P, = shell production, U = energy of excretory products, M p = pedal mucus production, Mf = faecal mucus production. Except for Ilyanassa obsoleta, Mr is included in F.

40

M.S. DAVIES AND S.J. HAWKINS

3524 kJ m-2 year-' for these terms, respectively. The total mucus component represented 31% of C and 60% of assimilated energy. Horn (1986) compiled population energy budgets for the chiton C. pelliserpentis and found the energy lost as both locomotory and faecal mucus in high-shore chitons was 37OkJm-*year-' (70% of C; 77% of assimilated energy) and 658 kJm-2 year-' (58% of C; 66% of assimilated energy) in low-shore chitons. Peck et al. (1987) compiled a laboratory budget for H . tuberculata and found the energy lost as locomotory mucus from animals of 0.01 g whole dry weight was 1.9 J d-' (23% of C ) and from animals weighing 50 g, 1120J d-' (29% of C). Davies et al. (1990a) measured mucus production in Patella vulgata and revised an earlier energy budget compiled by Wright and Hartnoll (1981) for this species by inserting their mucus (M) term. In the recalculated energy budget, mucus represented the largest sink for absorbed energy (at least 52%) and amounted to at least 23% of C. Further recalculations (Davies, 1991), basing mucus production values on the size of animals used in the original budget calculations (Wright, 1977) gives mucus as 7 1YOof absorbed energy and 3 1YOof C. Thus where it has been measured, mucus production has been shown to represent a consistently costly energy drain. This prompted Lamotte and Stern (1987) to comment "the measurement of mucus remains the most important gap in the knowledge of energy budgets in molluscs". Moreover, it is clear that those energy budgets, at least for marine gastropod species, which have been compiled without a mucus production term will be severely flawed and conclusions concerning function and constraints on lifestyle will be in error. For example, prior to the recalculation of Davies et al. (1990a), P . vulgata was thought to expend much (75%) of its assimilated energy on maintenance (R); post-recalculation this value was lowered to 38% suggesting this limpet operates much more economically than was previously thought. Table 3 provides a summary of the importance of mucus in molluscan energetics. The fate of mucus deposited during locomotion is considered in Section 7, but it is relevant here to mention energy export from molluscs. Most of the mucus produced by molluscs is available as a source of nutrition and energy to other organisms. The magnitude of such export shows remarkable consistencies between species, given that each calculated value is based on many different assumptions: Zlyanassa obsoleta, 2049 kJ m-2 year-' (Edwards and Welsh, 1982; Connecticut); Chiton pelliserpentis, 363 and 635kJm-2year-' (Horn, 1986; New Zealand); P . vulgata, 720 and 1624kJm-2year-' (Davies et al., 1990a, Davies, 1991; Isle of Man); Cellana grata, 829-6000 kJ m-2 year-' (Davies and Williams, 1995); Siphonaria japonica, 27.9-8026 kJ m-2 year-' (Davies and Williams, 1997); Siphonaria atra, 6.6-1541 kJ m-2 year-' (Davies and Williams, 1997). In one respect, these might be considered as maxima since much

41

MUCUS FROM MARINE MOLLUSCS

Table 3 Importance of mucus in molluscan energy budgets.

Species

Mucus type

Total Faecal Pedal Pedal Pedal + faecal Pedal Pedal Faecal Chiton pelliserpentis Pedal (high shore) Faecal Chiton pelliserpentis Pedal (low shore) Faecal Haliotis tuberculata Pedal Lymnaea peregra Total Physa acuta Total Patella vulgata Pedal Buccinum undatum Pedal Hypobranchial Pedal Concholepas concholepas Navanax inermis Ancylusjuviatilis Ancylus juviatilis Planorbis contortus Hydrobia ventrosa Cepaea nemoralis Ilyanassa obsoleta

Importance

Reference

7% of C , 4 6 % of Ab 9% of Ab 26% of Ab 9% of A 12% of A 23% of C, 8% of C, 68% of C, 1% of c, 56% of C, 2% of c, 23-29% of C , 47%a,bof C , 34%’’b O f c, 31% of C, 1 1 % of C , 17% of C , &20% of A

Paine (1965) Calow (1972) Calow (1974b) Calow (1974b) Kofoed (1975) Richardson (1975) Edwards and Welsh (1982) Horn (1986)

Peck et al. (1987) Blandenier and Perrin (1989) Davies (1991) Kideys and Hartnoll (1991) Navarro and Torrijos (1995)

C,, measured energy consumed; C,, energy consumed summed from other components of energy budget; Ab, energy absorbed; A, assimilated energy. ‘Estimated by subtraction. bIncluding excretion.

of the mucus is likely to be recycled by re-ingestion on subsequent grazing excursions. Mucus production is also an important consideration when calculating the “scope for growth” (SFG), as demonstrated by Navarro and Torrijos (1995) who included mucus in calculations of SFG in the gastropod Concholepas concholepas. Fry’s (1947) concept of the “scope for activity” as the difference between active and standard metabolic rates, was adapted by Warren and Davis (1967) in their definition of SFG as “the difference between the energy of the food an animal consumes and all other energy utilisations and losses”. Bayne ef al. (1987) defined SFG as “the energy available to the individual for growth and reproduction after all metabolic demands have been met from the absorbed ration”. This is slightly different from Warren and Davis’s definition but its method of calculation remains the same: SFG = C - ( R + F

+U)

42

M.S. DAVIES AND S.J. HAWKINS

SFG, like other budget terms, is expressed as energy per unit time, or per unit area per unit time. Gabbott (1976) defined SFG as the difference between the assimilated ration and the energy lost in respiration (A - R). However, this is incorrect because not all of this energy is available for growth, since mucus is a component of assimilated energy. By definition SFG should not include mucus production. However, if mucus is expressed in the budget as part of production (as P, or Pmuc)then, following the above equation, mucus production will be included in the SFG and the SFG value artificially raised. A similar result will ensue if mucus production is ignored in budgets. However, if mucus production is included in the U term then the SFG value will not be biased as it will not include mucus production. If mucus production is expressed separately as S or M then a modification of the equation ensures the SFG is not biased, for example, using M: SFG = C - ( R + F + U + M ) Although the SFG provides a descriptor of growth availability over a wide range of environmental variables and has been used widely in studies of physiological adaptation in molluscs, particularly bivalves (see Bayne, 1976; Newell and Branch, 1980; Bayne and Newell, 1983; Bayne et al., 1987), many authors, for example, Dame (1972), Worrall et al. (1983), Garton (1986), Bayne et al. (1987), Magnusson et al. (1988), ignored mucus in their SFG calculations. They all defined SFG correctly, but then did not fulfil the definition in their calculations. This had the effect of artificially increasing the SFG value. Unfortunately, the amount by which the SFG was inflated in this way is unknown, since the quantity of mucus released with faeces and pseudofaeces was not determined. The observed discrepancies between the SFG and actual growth (e.g. Dame, 1972; Bayne and Worrall, 1980; Hummel, 1985; see Bayne and Newell, 1983, for review) may be owing at least in part to a lack of consideration of mucus production, and is an area ripe for further study, given the likely importance of pseudofaecal mucus in bivalve energetics. The study of mucus in bivalve energetics assumes particular importance as SFG is used as an index of physiological condition, since it is quick to determine and allows repeat measurements on the same animal (Bayne and Newell, 1983). However, lowered SFG is used as an indicator of pollution, often in turbid estuarine areas where pseudofaecal - and hence mucus production is likely to be great (see Prins and Smaal, 1989). Thus SFG may be most inaccurate where its value is most relied upon. Measurements of absorbed ration are routinely made by subtracting the energy content of the faeces from that of the consumed food. This widely used approach (see Bayne and Newell, 1983) assumes that organic material is only taken from, and not added to, the contents of the gut. However,

MUCUS FROM MARINE MOLLUSCS

43

lubricatory mucus (see Section 4), digestive juices, bacteria and sloughed-off intestinal cells will also emerge with the faeces and are more often than not left unaccounted. Such “metabolic faecal loss” (MFL) accounts for the negative absorption efficiencies recorded by many workers from bivalves in low food concentrations (see Bayne and Hawkins, 1990). The magnitude of MFL has rarely been measured (although it is thought to represent up to 20% of assimilated energy in bivalves, see Bayne et al., 1989) and its composition is unknown. MFL varies with season and animal condition (Hawkins and Bayne, 1992) and can comprise 25-89% of absorbed N (Hawkins and Bayne, 1985).

7. ECOLOGY

Since mucus is functionally important within marine mollusc species and, where measured, is an important component of animal energy balance, it is hardly surprising that mucus can also shape interactions between species. Here we address its wider roles at the community and ecosystem levels in terms of the fate of the mucus produced by marine molluscs and its involvement in biological interactions. 7.1. Fate of Mucus

There are a few good estimates of mucus production and its export from particular organisms to the wider marine environment (see Section 6). Rocky shore molluscs are perhaps the best studied group (e.g. Connor, 1986; Herndl and Peduzzi, 1989; Peduzzi and Herndl, 1991; Davies et al., 1992a,b,c; Davies and Williams, 1995). They lay mucus down as trails when foraging as well as covering their epithelial surfaces with mucus. From rates of production per individual, population level estimates of total mucus deposited on the shore can be made by simple multiplication. It is, however, important to know how long this mucus persists in order to evaluate its role. Field studies (Davies et al., 1992b) have shown that Patella vulgata mucus persists for up to 80d; somewhat longer than the non-homing Littorina littorea (40d) (Figure 7). Davies et al. (1992b) showed that about 8-10% (w/w) was lost over a 6-h period of immersion. They attributed this largely to the mechanical action of seawater rather than microbial breakdown (but see Herndl and Velimrov, 1986), such that higher up the shore the persistence of mucus would be longer. In the intertidal zone, rehydration and dehydration of mucus will contribute to its physical breakdown. In areas of greater wave action mucus would be expected to breakdown faster (Connor, 1986) than

44

M.S. DAVIES AND S.J. HAWKINS

in shelter. On a moderately exposed rocky shore in the Isle of Man a population of P. vulgata produced 220mgm-’ d-’ of mucus. Energy exported from P . vulgata alone on that shore would be 21 J m-2 for the first 6 h and of the order of 42 Jm-2d-’ assuming 12h in water (Davies et al., 1990a, 1992~).These are probably upper limits as much of the mucus could be recycled by re-ingestion on subsequent grazing excursions. Notwithstanding re-ingestion and attack by other macroscopic feeders, it is still likely that mucus from intertidal molluscs makes a considerable direct contribution to both particulate and dissolved organic matter (DOM) in coastal waters. Many plants and lower animals (including 11 invertebrate phyla, see Stephens, 1967, for review) absorb certain dissolved compounds from seawater (Khailov and Burlakova, 1969). Other organisms such as bacteria (Khailov and Burlakova, 1969), uni- and multicellular algae (Provasoli, 1963) and some invertebrates (see Allendorff, 1981; Stephens, 1983; Wright and Secomb, 1986) rely on DOM as their main source of nutrition, although the processes of uptake are poorly understood (Khailov and Burlakova, 1969). Mucus trails will also form a substrate for bacterial utilization (Herndl and Peduzzi, 1989; Peduzzi and Herndl, 1991) which will form another pathway for mucus degradation. Autotrophic organisms, such as diatoms, may also derive nutritional benefit from the mucus (Connor and Quinn, 1984; Connor, 1986). Thus simple export to the water column may be only one of several potential fates for the mucus laid down by intertidal molluscs. On soft shores, mucus is produced by epibenthic diatoms (Hoagland et al., 1993) as well as burrowing animals including bivalves and polychaetes (Jones and Jago, 1993). Tube-forming amphipods (Corophium) and polychaetes (Pectinaria, Owenia) also use mucus to bind particles together. The main ecological implication of these exopolymers is the stabilization of sediments (e.g. Eckman, 1985; Madsen et al., 1993) and thereby a reduction in erosion and transport (e.g. Paterson, 1989). Burrows in the sediment are consolidated by mucus (Grant et al., 1986; Nehring et al., 1990) and enable greater irrigation of the sediment allowing oxygen penetration to greater depths. Much of this oxygenation is confined to the burrow walls themselves, and does not penetrate laterally into the sediment (Wetzel et al., 1995). Mucus is the essential binding component of bivalve pseudofaeces (see Section 5.3). Considerable amounts of particulate organic material (POM) can be exported to the water column by the pseudofaecal production (Haven and Morales-Alamo, 1966; Foster-Smith, 1975; Hildreth, 1980; Kautsky and Evans, 1987) of dense beds of filter-feeding molluscs, particularly if the sediment load is high in the seston. Intertidal (Tsuchiya, 1980) and subtidal mussel (Kautsky and Evans, 1987) and oyster beds are natural exporters of such pseudofaecal material (Deslous-Paoli et al., 1992) which

MUCUS FROM MARINE MOLLUSCS

45

can make a considerable contribution to POM in inshore waters. The effects of biodeposits are increased in the enclosed systems used for intensive or extensive culture of bivalves (Sornin et al., 1983). Biodeposits from rope, raft or raised platform culture can smother underlying benthos as well as causing increased oxygen demand, leading to anoxic sediments (Grant et al., 1995; McIntyre, 1994; Hevia et al., 1996; Mazouni et al., 1996) such that benthic populations decline in number (Kusuki, 1977). By aggregating particles, production of pseudo faeces also increases the flux of suspended material to the benthos (by up to 200x; Deslous-Paoli et al., 1992), particularly in rope- or raft-cultured bivalves (Mariojouls and Kusuki, 1987; Mariojouls and Sornin, 1987). The long-term fates of mucus remain speculative. Radiolabelling of mucus would be a useful technique to accurately determine its persistence. Additional work aimed at determining the fate of mucus once released from the substratum and its possible uptake by marine organisms is required, especially modelling work involving laboratory microcosm experiments where the effect of a mucus input could be monitored. 7.2. Role in Biological Interactions

Mucus is involved in various intra- and interspecific interactions which in turn shape population and community structure and dynamics. Given the persistence of mucus (Figure 7), the density of gastropods and their motility patterns, at least on British shores, most of the substratum is likely to be covered for most of the time with a layer of mucus; a layer that is bound to be important at the ecosystem level. Conspecific trail following behaviours often result in aggregated distributions. These in turn can lead to patchiness in communities owing to uneven distribution of grazing or predation intensity (Hawkins et al., 1992). Wahl and Sonnichsen (1992) noted that the shells of Littorina littorea were less fouled where snails were more aggregated and suggested that this might be owing to antifouling properties of pedal mucus deposited as snails crawled over one another. Trematodes parasitic on marine snails may also use mucus trails as an aid to the dispersion of their cercariae (Curtis, 1993). Predatory gastropods often follow the mucus trails of their prey. Some prey use obnoxious mucus as their first line of defence when attacked by predators (Rice, 1985; Avila et al., 1991). Siphonaria mucus, for example, can be particularly noxious and is avoided by predatory species (Branch, 1981, pers. comm.; but see Iwasaki, 1993). Mucus of potential predators may also elicit escape responses, even in the absence of the predator itself. The well-defined escape response of Littorina irrorata (Hamilton, 1977; Warren, 1985; Dix and Hamilton, 1993), which seeks refuge on the stalks of salt-

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marsh vegetation, can be stimulated by the mucus of one of its major predators (Dix and Hamilton, 1993) plus that of other predatory whelks which it does not usually encounter (Dix and Hamilton, 1993). Mucus may be involved in enhancing food resources for grazing molluscs. Mucus deposited by gastropods has been shown to be adhesive and concentrates both organic and inorganic food particles (Connor, 1986; Peduzzi and Herndl, 1991; Davies et al., 1992a) (Figures 8 and 10). Presumably mucus produced by other molluscs will behave in the same way. It is known that more algal propagules and microalgae attach to mucus-covered surfaces than to clean surfaces (Santelices and Bobadilla, 1996; pers. obs.), although this work was done on smooth glass slides which may exaggerate the effect. The stimulation of both bacterial and microalgal growth by the pedal mucus of acmaeids has been shown (Connor and Quinn, 1984; Connor, 1986): in general microalgae follow bacteria in colonization after the bacteria have “conditioned” the mucus trail. This conditioning may involve bacterial release of substances from the mucus (breakdown products) or the accumulation of bacterial metabolic residues. Both could stimulate microalgal growth. Herndl and Peduzzi (1989) and Peduzzi and Herndl (1991) have shown increased microbial activity in the trails of limpets and trochids which would enhance their food value for subsequent grazers. Similar increases in microbial activity have been observed in the deposited mucus of certain Foraminifera which “farm” the bacteria and fungi present in the mucus (Langer and Gehring, 1993). However, for such mucus which promotes growth to be a crucial element in food gathering, some kind of expectation that an individual animal will recoup its investment is essential for an evolutionarily stable strategy to develop (Calow, 1979; Davies et al., 1992a): the producer of the mucus must also be its consumer (see Section 5). Thus those gastropods which mate (and follow a mucus trail to do so) do not expend energy on producing mucus which serves to enhance their diet, but may nevertheless reduce the net cost of mucus production (through its partial recycling) and reduce the cost of foraging (see Section 5). Using mucus serendipitously in this way, as a “bonus” to nutrition, seems to be a likely strategy for nonterritorial gastropods. For those species which apparently deliberately use mucus as a provendering agent (e.g. L. gigantea, C. (Macclintockia) scabra, Connor, 1986), the enhancement of nutrition in this way may lead to a reduction in foraging distances and a net saving in energy, as secreted mucus, which could be used to increase gonadal output. In addition, the mucus assemblage may be organically richer (Figure 10) than the microalgal assemblage on rock surfaces and so these animals may have an increased absorption efficiency. Thus differences in energy budgets between grazers may partly depend on their mode of foraging. Use of trails for nutrition where gastropods move more than a few body-lengths also implies either a

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Figure I0 The adhesive nature of gastropod mucus. Scanning electron micrograph of a sheet of cellulose nitrate left in the intertidal (Isle of Man) for 24 h, the left-hand half of which was covered in a layer of pedal mucus from the limpet Patella vulgata. Adhering to the mucus is a wide range of inorganic and organic matter, including diatoms (arrowed). Scale bar = 100pm.

kinaesthetic memory or an aid to navigation within the trail (see Section 5). Gastropod pedal mucus can absorb metals from solution (by up to 10,000 x; Davies, unpubl.) and grazing on mucus might not be beneficial. Mucus which is not ingested might enhance succession by accelerating the recruitment to, and the colonization of, shores by macroalgae, taking into account the chances of spores and sporelings being dislodged or eaten by a grazer, and surviving (see Lubchenco and Gaines, 1981; Hawkins and Hartnoll, 1983). Microbiota within a mucus matrix may experience a concentration of nutrients and thus grow faster than those outside the mucus. For gardening to be proven, there is a need to measure the degree of reingestion of mucus by limpets and chitons making individual excursions from home scars: what is needed is the amount of the trail which is retraced on initial and subsequent foraging excursions and whether this is likely to be higher than that of neighbours. The incidence of “cheating” by other individuals by consuming the mucus laid down by others needs to be measured too. This could well be a strategy pursued by non-homing juvenile limpets. The degree of ingestion by other members of the assemblage such as nonterritorial limpets, trochids and littorinids needs to be assessed. The ener-

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getic value of mucus trails with or without microbial or algal enhancement needs to be compared with the background value of the microbial film. It would also be beneficial to know when re-ingestion of trails takes place. Presumably it occurs after a period of trail enrichment: do gastropods re-ingest when the energy value of the deposited trail aggregation is at a maximum? Mucus has been implicated in gardening behaviour (e.g. Connor, 1986; see above) but this has not been demonstrated for the spectacular gardening species of South Africa (Branch, 1981; see Branch et al., 1992 for review). For the gardening species Patella longicosta, McQuaid and Froneman (1993) tested the effect of applying mucus to Ralfsia (the “gardened” alga). No elevation of production was apparent, but this experiment was undertaken in the rather artificial conditions of the laboratory and the mucus was applied with a paintbrush. South African limpets tend to garden a particular species. In contrast, Lottia maintains an enhanced microbial film lawn which it defends. Perhaps mucus is more important when microbial gardening is occurring; this is certainly the case in many sediment communities where gardening has been implicated (Woodin, 1977; see Branch et al., 1992 for review) as well as with Foraminifera growing on seagrasses (Langer and Gehring, 1993). Mucus may also influence the distribution of species, particularly when larvae use it as a cue during settlement. Experimentation to detect the presence of such cues should use naturally laid mucus trails, since Seki and Kan-No (198 1) found that in comparison mucus collected from the pedal sole did not illicit a response of conspecific settlement in Haliotis discus hannai. Mucus may be a positive cue, as has been demonstrated for several species of abalone (e.g. Haliotis, Seki and Kan-No, 1981; Searcy-Bernal et al., 1992). Larvae preferentially settle on mucus tracks laid down by conspecifics (Seki and Kan-No, 1981). This behaviour was used extensively in the aquaculture industry (Hooker and Morse, 1985; Hahn, 1989), although more recently chemical stimulants such as GABA have been widely used (Morse et al., 1979). This behaviour results in gregarious settlement, with the larva presumably using the presence of adults as an indicator of favourable conditions for survival in the future. Settling larvae can also exhibit avoidance behaviour. This has been demonstrated for barnacles in response to the mucus of predatory gastropods. Johnson and Strathmann (1989) showed that Balanus glandula avoided tiles previously occupied by Nucella lamellosa; Semibalanus cariosus also settled less on the tiles but did not show such a strong response and the results were not significant at the 0.05 level. The response of B. glandula was also induced by rubbing mucus from the foot of the whelk over the tiles. A similar but less pronounced effect was also induced by mucus from the limpet Tectura (Collisella) scutum, which is known to bulldoze settling bar-

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nacles from the surface (Dayton, 1971). Mucus from the brown alga Fucus distichus was also shown to have an inhibitory effect. Johnson and Strathmann (1 989) suggested that Fucus distichus is not a demonstrated hazard to barnacles (Dayton, 1971; Farrell, 1987); although in some instances sweeping of the substratum by other species of Fucus (Menge, 1976; Grant, 1977; Hawkins, 1983) has been shown to reduce barnacle settlement and thus it may represent a hazard. Johnson and Strathmann (1989) offer avoidance responses as an alternative explanation to biological disturbance in generating different localized patterns of recruitment. However, no studies have been made which separate the putative sweeping effect from a possible inhibition of settlement by mucus. A note of caution is required in interpreting these data. In their discussion Johnson and Strathmann pointed out that settlement was also enhanced on plates covered by nudibranch mucus (Archidoris montereyensis - a sponge-eating species). They attributed this to fertilization of the microflora (e.g. Connor and Quinn, 1984), but we also consider that a direct mechanical effect of the mucus itself cannot be ruled out. After weighing up the various evidence Johnson and Strathmann discount the direct barrier effects of the mucus itself and suggest that it is being used as a cue by the larvae of B. glandula to avoid the future risk of predation by whelks. The almost negligible effect on S . cariosus is attributed to their lower vulnerability to predation. They predicted that the response to limpet mucus would be stronger than that of the whelk as bulldozing was predicted to be a greater risk early on, to settling and recently metamorphosed barnacles. This was not the case. They pointed out that bulldozing may not be such a large risk as previously thought (cf. Miller, 1986; Dayton, 1971; Hawkins, 1983). Some species of barnacles can, paradoxically, use mucus from their whelk predators as a positive cue to settle. Raimondi (1988) showed that cyprislarvae of Chthamalus anisopoma were attracted to surfaces covered by the mucus of their predators, Acanthina angelica. C. anisopoma occurs in the harsh environment of Baja, California, but only occurs in the same zone as A . angelica. B. glandula in contrast has a refuge above the Nucella spp. which prey upon it. Johnson and Strathmann (1989) suggested that the risk of death to C . anisopoma owing to A . angelica may be less than that from desiccation. C . anisopoma also has a neat morphological defence to predation by Acanthina. Two morphs occur (Lively, 1986a,b); the bent morph is difficult for whelks to eat, whilst the upright morph is more vulnerable but more competitive as it grows faster and is more fecund. Thus C . anisopoma settling in the same zone as the whelks have a defence. This morphological response is induced by contact with the whelks (presumably via mucus) during growth; it is not owing to differential mortality of two genetically different polymorphisms.

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Some more recent work (Proud, 1994) has also shown that cyprids of Semibalanus balanoides avoid areas on which Nucella lapillus have been placed and presumably left mucus or some other chemical signature. Interestingly, settlement seemed to be enhanced in areas where the grazer Littorina littorea had been placed and subsequently crawled away. Barnacles can be overgrown by ephemeral algae and will persist longer in areas with many grazers. Thus a clue that the area is well grazed could indicate enhanced probability of survival in the future. Whether the cue is positive or negative, settling larvae can clearly detect the presence of chemicals laid down in mucus and use them to stimulate or inhibit settlement. This shows adaptation to the widespread presence and often persistent nature of mucus on hard surfaces in the sea.

8. OVERVIEW

Back in 1979, Calow made the following plea for further work on mucus: There can be little doubt also that much mucus is produced by animals and that it is likely to have a non-trivial influence on the physical and trophic properties of the ecosystems into which it is released. Hence, both from the point of view of evolutionary biology and ecosystem ecology, mucus deserves more serious attention than it has been given (Calow, 1979). Since 1979 much more attention has been given to the importance of molluscan mucus at the individual level, its role at the population and community levels and its contribution to ecosystem function. Much remains to be done, however, in linking biochemical composition with function - an area almost devoid of work in the invertebrates - particularly where an organism manufactures many types of mucus from different tissues or glands. Nevertheless, it is clear that mucus is a key component in most physiological functions of molluscs including locomotion, respiration, feeding and digestion. It is also the major interface with the environment, being used as protection, lubrication, food gatherer and even a conduit for defensive chemicals (see Section 5). The cost of mucus production for chitons, gastropods and bivalves has begun to be appreciated. In chitons and gastropods it can be up to 80% of ingested energy and is usually at least 30%. It is very likely to be more expensive than the respiratory costs of locomotion in many animals. It may also be costlier to produce in air than water for intertidal species, which will be a constraint on foraging (Santini et al., 1995). The cost of mucus begs the question whether some of its production is recouped by reingestion. Its value could be further enhanced by additional energy sources

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as microorganisms utilizing the film or microalgae or microscopic propagules of seaweeds settling and growing more rapidly in the mucus matrix. For this to be an evolutionarily stable strategy an individual animal must benefit from its initial investment; this may well be a crucial element in the evolution of gardening by some gastropods - particularly limpets - where individuals defend a territory. The maintenance of grazed microbial lawns by dense populations of gastropods or chitons means that their mucus is likely to be a very important part of the biofilm coating most shores. In addition to gastropod mucus, there will be that exuded by diatoms and the other components of the film. Methods to quantify the source and frequency of these different exopolymers and their role in the biofilm are required. Confocal microscopy coupled with lectin-based staining is one way forward. It is very likely that the floral composition of the biofilm is not only influenced by grazing activities but by the mucus deposition of the grazers. Predatory molluscs crawling over the rock surface also make their contribution and leave their signature. This mucus can be used as a cue, both enhancing and inhibiting larval settlement. For bivalves, there is much work to be done on determining rates of mucus production. Finally all this mucus must go somewhere. No doubt some is ingested, but some is exported to the water column where in the words of Calow it will make a “non-trivial” contribution, as yet unquantified except in a few preliminary estimates. Mucus does deserve “more serious attention”. We hope this review stimulates some.

ACKNOWLEDGEMENTS We are grateful to the following: Dr G. Walker for permission to redraw Figure 2; Dr A. Cook for permission to redraw Figure 3; to Dr P. Beninger and Springer-Verlag GmbH for permission to reproduce Figure 4.

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Constraints on Coastal Lagoon Fisheries Jean-Christophe Joyeux’ and Ann Baker Ward2

’North Carolina State University. Department of Zoology. box 7617. Raleigh NC 27695. USA (Present address: Universitk Montpellier II. Laboratoire d’Hydrobiologie Marine et Continentale. CNRS UMR 5556. case 093. Place E . Bataillon. 34095 Montpellier Cedex 5 . France) 2North Carolina State University. D.H. Hill Library. Box 7111. Raleigh NC 27695. USA

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1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. Overview . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. What is a Lagoon? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Some observations on the Yield ................................. 2.3. Some Observations on the Fisheries .............................. 2.4. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3. Material and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Data Description and Collection .................................. 3.2. Statistical Treatment ......................................... 4 Geographical and Morphometrical Constraints ......................... 4.1. Descriptive Statistics ......................................... 4.2. Geographical Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Morphometrical Data ......................................... 4.4. Multivariate Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5. Discussion ................................................. 4.6. Conclusions ................................................ 5. Environmental and Anthropogenic Constraints ......................... 5.1. Descriptive Statistics ......................................... 5.2. Water Exchange Data ......................................... 5.3. Physico-chemical Data ........................................ 5.4. Biological Data .............................................. 5.5. Fishing Effort and Catch per Unit Effort ........................... 5.6. Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.7. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6. Final Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References .................................................... Appendix1 ....................................................

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Appendix2 Appendix3

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ABSTRACT We have estimated the individual and synergetic effects of geographical, morphometrical, environmental and anthopogenic parameters on the fishery yield of coastal lagoons throughout the world. For this analysis we have relied on published literature and other information obtained directly from researchers and fishery officials, since special global scale sampling was not possible. Lagoons that are exploited through alternative fishery practices had significantly higher yields than the others. The fishery yield was dependent upon a lagoon’s geographical location and morphometry . The significant environmental and anthropogenic factors determining the yield were the exchanges of water between lagoons and the ocean, the physico-chemical properties of the water, the extent of aquatic vegetation, and the fishing pressure. Overall, the results show the single influence of freshwater input to be negligible while the influence of oceanic tide exercised through the inlets was significant. Fishery yield appeared to be correlated with the mean annual concentration of nitrites. Submerged and emerged vegetation were good indicators for fishery productivity. The single most influential factor of all was the fishing pressure. Data collected for 292 fished lagoons are given in appendix tables and include annual fishery yield; water area; mean and maximum depth; watershed area; annual freshwater inputs; inlet(s) width; height of ocean tide; minimum and maximum temperatures and salinities; mean and maximum concentrations of nitrite, nitrate, and orthophosphate; areas of immersed and emergent vegetation; and number of fishermen.

1. INTRODUCTION

Coastal lagoons are generally small water bodies (Kapetsky, 1984), widely and profusely distributed along the world’s coasts. Brackish-water lagoons and similar bodies of water, such as backwaters and saline lagoons, are estimated to occupy 13% of the world’s coastline (Cromwell, 1971; Lasserre, 1979a). Like other estuarine areas, lagoons are widely used for fishing and aquaculture. Their fishery yields, i.e. annual landings per unit

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area, can be very high (Kapetsky, 1984), to the point that Amanieu (1973) stated that on a regional basis the economic value associated with fishing and aquaculture in Mediterranean lagoons often equalled or was more important than the value from adjacent sea shelf areas. The reasons for t h s high production are not known, but the accessibility and protection offered by these areas compared with those of the open sea are probably instrumental in their intensive use. However, the economic influence of bodies of coastal water extends well outside their own geographic boundaries; estuaries and lagoons are of great importance to shelf fisheries in numerous regions of the world (Newell and Barber, 1975; McHugh, 1976; Pollard, 1981; Lenanton and Potter, 1987). More precisely the yield of offshore fisheries seems correlated to the estuarine area (Barret and Gillespie, 1973; Amanieu and Lasserre, 1981; Yaiiez-Arancibia et al., 1985; Deegan et al., 1986; Soberon-Chavez et al., 1986; YBiiez-Arancibia and Aguirre Leon, 1988) or to the immersed or emerged vegetation associated with coastal waters (Turner, 1977, 1979 and 1986; Martosubroto and Naamin, 1977; Nixon, 1980; Pauly and Ingles, 1986). The fishery productivity of estuaries and coastal marine ecosystems is considered dependent upon primary production (Nixon, 1981, 1982; Nixon et al., 1986), which is itself dependent upon nutrient concentration (Houde and Rutherford, 1993). Authors have speculated that nutrient loading is responsible for biological and fishery richness of lagoons (Nixon, 1982; Kapetsky, 1984). Also, increases in agricultural and urban waste run-off during the last decades are presumed to be at the origin of increasing fishery yields in some lagoons (Kapetsky, 1984). Assuming that this fundamental role of nutrients has a measurable biological impact in estuaries, fishery production depends on nutrient uptake, assimilation efficiency and fishing efficacy (Kapetsky, 1984; Bayley, 1988). The first two factors are expected to modulate the standing crop of fish and crustaceans according to the physical and chemical characteristics of the lagoon. Connections between lagoon and the hinterland appear to be of utmost, but somewhat remote and unclear, importance in respect to fisheries. Indeed, fishery landings have been correlated with freshwater inputs in, or salinity of, coastal waters (e.g., Texas Department of Water Resources, 1979, 1981a,b). The continental domain is also the source of physical and chemical inputs; for example, excess of nutrients is a common and welldocumented occurrence (e.g., McComb et al., 1981; Hodgkin and Birch, 1982; McComb and Humphries, 1992 for Peel-Harvey estuary). Loss of seagrass beds and serious mortality of mangrove seedlings have also been blamed on high levels of eutrophication (Brodie, 1995). Attacks on aquatic vegetation, including chemical (oils, pollutants) and mechanical (dredging, filling) factors, have resulted in lost biological productivity and fishery production (Boesch and Turner, 1984; Pollard, 1984; MacDonald and

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Hancock, 1992), even for some species that are not directly dependent upon vegetation. Lagoons also connect, either permanently or intermittently, to the ocean which is often the main source of commercial species caught in estuaries (Jhingran and Gopalakrishnan, 1973; Miller et al., 1984; Hettler and Barber, 1993). Specifically, inlets and passes are an obligatory thoroughfare for the young of offshore spawning species that utilize estuaries as nursery areas to complete their development, Therefore, capabilities, or possibilities, to reach and enter estuaries may greatly affect subsequent catches at an older age, both inshore and offshore. Astronomical tides and non-local, wind-driven processes have been proposed as the principal mechanisms responsible for larvae and early juvenile penetration into estuaries (Lawler et al., 1988: Pietrafesa and Janowitz, 1988). These factors will also influence the physico-chemical regime of the lagoon according to its morphometry and morphology. Since the local economic value of their fisheries is important (Amanieu and Lasserre, 1981; Pollard, 1981; Houde and Rutherford, 1993), lagoons regionally generate a large number of studies. Various authors have attempted to predict fishery yields in coastal waters (Fogarty, 1989), especially in US estuaries. Most of them reported significant relationships between the yield of a species or of a group of species and the freshwater inflow previously entering the system (Copeland 1966; Meeter et al., 1979; Texas Department of Water Resources, 1979, 1981a,b; Browder and Moore, 1981). Less attention, however, has been given to broader approaches to estuarine productivity which are the most likely, in our opinion, to demonstrate functional links between fishery production and causal parameters. Few studies have addressed, for example, the regional or global influence of the environmental variables upon the yield, although this approach has often been successful for managing freshwater ecosystems (Rawson, 1951, 1952; Ryder, 1965; Ryder et al., 1974; Schlesinger and Regier, 1982; Hanson and Leggett, 1982; Jenkins, 1982; Marshall, 1984; Jackson and Ssentongo, 1988; Regier et al., 1988; Ranta and Lindstrom, 1989, 1990; Moreau and De Silva, 1991; Ranta et al., 1992). In lakes, reservoirs, rivers, and floodplains, relationships between fishery yields and environmental variables are primarily based on the Morphoedaphic Index. This index is derived from the ratio of total-dissolved-solid concentration to the mean depth. Relationships may also integrate temperature, latitude, benthic biomass, water area, area of drainage basin, length of shoreline, water transparency and fishing effort. Syntheses for lagoon environments were locally initiated by Copeland (1966) for the estuaries of Texas, regionally by Amanieu and Lasserre (1981) for the lagoons of the Mediterranean, and globally by Kapetsky (1984). These studies offered a direct, but limited, view into the parameters

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responsible for, or related to, the fishery production into a variety of systems. Thus, Amanieu and Lasserre (1981), Kapetsky (1983, 1984) and Chauvet (1988) have successfully shown that fishery yield in lagoons is related to water area and fishing pressure. Marten and Polovina (1982) and Bayley (1988) failed to show any relationship between yield and fishing effort in, respectively, intertropical and African lagoons. At the present time it is not known if the fishery yield from lagoons is linked to any environmental (sensu luto) characteristic other than water area. However, numerous environmental parameters, possibly acting in positive or negative synergy, could be implicated in the mechanisms responsible for natural productivity (Nixon, 1981, 1982, 1988). Previous syntheses on lagoon fishery yields therefore lacked a multidimensional approach in the sense that only a few parameters, such as latitude, water area and fishing effort, were included and tested with respect to fishery yields. To remedy this deficiency we searched for data on fishery production and variables relevant to fishery production in lagoons, with the ultimate aim of providing a basis for their management. For this reason, the first objective was to identify the parameters affecting the fishery yield and to characterize their individual influence. The second objective was to build a comprehensive model, or models, integrating interactions among these parameters. Once the general principles of productivity are known, monitoring or control of the key parameters, or links among parameters, will introduce greater efficiency in managing lagoon fisheries. Because lagoons and their associated fisheries are extraordinarily diverse and because the relations between one and the others are likely to be complex, this review has been divided into several sections. Section 2 provides a specific, and selective, context to subsequent analyses. Section 3 describes the data that were compiled and the statistical methods used for the analysis. Section 4 reports our results on the geographical and morphometrical constraints on the fisheries, on which Man has no, or extremely limited, influence. Section 5 expands the results from the preceding section on the environmental constraints, often directly or indirectly affected by human activities, and the anthropogenic constraints. Section 6 concludes this work in proposing some points of action.

2. OVERVIEW 2.1. What is a Lagoon7

Given the confusion associated with the respectivedescriptionsof a lagoon and an estuary and the necessity to define precisely the object of our research, we

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need to provide some background definitions. Because definitions for lagoon and estuary are neither inclusive nor exclusive, these two words are often alternatively used as synonyms and antonyms. There is, generally, an implicit assumption to accept most lagoons as a kind of estuary, along with drowned river valleys and fjords. Pritchard (1967) who considered coastal lagoons to be estuaries, defined an estuary as “a semi-enclosed coastal body of water which has a free connection with the open sea and within which sea water is measurably diluted with fresh water derived from land drainage”. According to this description, bodies of coastal waters with ephemeral inlets or hypersaline characteristics (e.g. Sivash, 133*, Ukraine; Caimanero-Huitzache, 232, Mexico) are not estuaries, although they are indubitably lagoons. Furthermore, it is acceptable to recognize estuaries within lagoons (e.g. Pamlico and Neuse River estuaries in Pamlico Sound, 285, USA) or fjord-like structures (e.g. Sacramento River estuary in San Francisco Bay). Emery and Stevenson (1957) distinguished estuaries from lagoons by the origin of the barrier that separates them from adjacent coastal waters. More specifically, these authors characterized estuaries by a barrier that is shaped by “non-marine agencies”, while the barrier is of marine origin for lagoons. Later, Colombo (1977) proposed that “Lagoons are shallow bodies of brackish or sea water partially separated from an adjacent coastal sea by barriers of sand or shingle, which only leave narrow openings through which seawater can flow”. Again, hypersaline bodies were excluded. Moreover, untypical barriers disqualified some lagoons (Berre, 3 1, France; Fondi, 45, Italy; Maracaibo, 255, Venezuela; among others). Similarly, Lasserre (1979a) described coastal lagoons as shallow depressions located between the shore and a bar that generally allows some kind of communication with the ocean. According to this author, “the feature distinguishing lagoons from estuaries is the presence of the offshore bar”. Lasserre (1979a) distinguished four types of lagoons: estuarine, open, partially closed and closed lagoons. Lankford (1976) broadly defined a lagoon as “a coastal zone depression below MHHW [mean higher high water], having permanent or ephemeral communication with the sea, but protected from the sea by some type of barrier”. He recognized five types of lagoons, each divided into several subtypes. The classification was based on various mechanisms at the origin of the depression and on barrier characteristics. Lankford’s (1976) types of lagoons were: Type I - lagoon created through differential erosion during the periods of lower sea level (e.g. during glaciations), with or without a barrier; *Numbers refer to the list given in Table 1 (see Section 3).

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Type I1 - lagoons mainly produced by irregular sedimentation at the mouth of a freshwater system; Type I11 - submerged margin of the continental shelf with sand barriers developing from the action of the sea; .Type IV - lagoons produced by living organisms such as corals and mangroves; Type V - depressions or barriers produced by faulting, folding or volcanism. Interestingly, drowned river valleys, commonly called estuaries, and fjordlike bodies were included in the definition of Type I. Yaiiez-Arancibia (1977) noted that only 74, out of the 123 lagoons surveyed by Lankford (1976) in Mexico, were typical coastal lagoons with sandy bars. Kjerfve (1986) hypothesized that “physical lagoons characteristics and variabilities depend primarily on the nature of the channel(s) connecting the lagoon to the adjacent coastal ocean”. Discussing Lankford’s Types I, I1 and 111 without regard to the origin of the depression and the barrier, Kjerfve recognized three categories of lagoons. Choked lagoons have a small single entrance and are most common on coasts with high wave energy and low tidal range (although this may not be characteristic for the Mediterranean region; cf. Uyguner and Gozenalp, 1959; De Angelis, 1960; Kerambrun, 1986; Ardizzone et al., 1988). Leaky lagoons have multiple or proportionally large entrances, and are located along coasts with variable tidal and wave characteristics. Restricted lagoons are in the middle of the spectrum and are usually located on low/medium wave energy coasts with a low tidal range. Finally, Day and Yaiiez-Arancibia (1982) concluded that “from an ecological point of view, however, coastal lagoons and estuaries constitute a similar type of ecosystem and we can speak of a lagoon-estuarine environment”. This observation is consistent with the findings of Boynton et al. (1982) who demonstrated that fjords, embayments, lagoons and river-dominated estuaries can be significantly classified according to salinity, light extinction coefficient, latitude and flushing rate. By contrast, other parameters such as phytoplankton production, chlorophyll a and nutrient concentrations are not of significant importance. Our search for data on lagoons focused on Lankford’s (1976) Types I, I1 and 111, closest to Colombo’s (1977) and Lasserre’s (1979a) definitions, without avoiding most other types of lagoons/estuaries. We consciously rejected some lagoons of Type I (open drowned river valleys), and of Type IV (coralgal lagoons). The former subtypes have no bar, while the latter define a particular environment that is clearly differentiated from coastal lagoons and estuaries as they were defined previously. We have included a few lagoons of unclear origin.

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2.2. Some Observations on the Yield

Available landing data rarely reflect what was actually harvested from a lagoon, and this was a common problem during the collection of the data. In fact, data represent those that were either surveyed, reported or sold. The part of the catch that is regularly discarded seems highly variable and essentially dependent upon the fishing techniques, the targeted species and the purpose of the fishery. For example, surveys indicated that the commercial yield in Mauguio (30, France) varied between 74.6 and 120.3kgha-’yr-I, without eel, for the years 1986-1989 (Figure 1). This was the by-catch and eel, the targeted species, represented an additional 24-54% (Bouchereau, pers. comm.). The discards, which include undersized fish, shrimp or crab, and non-valuable species, varied between 14.0 and 31.3 kgha-’ yr-’ (Quignard et al., 1989; Bouchereau et al., 1990). Considering the landings of eel, the discards represented 8.7-19.4% of the commercial catch, which is not negligible (Figure 1). Unfortunately, data on discards are rare so this aspect of the lagoons’ non-utilization or misuse could not be studied. Wide differences in the ratio of discards or by-catch to the target species can be expected. In contrast to the relatively low quantity of by-catch in Mauguio (between one to three times the volume of the targeted species), Yaiiez-Arancibia and Aguirre-LBon (1988) reported that in the Gulf of Mexico by-catch was 4.2 to 15.9 times higher by weight than catches of the targeted shrimp. The definition of what is included in discards and by-catch also changes according to the availability of the target species.

Figure I Yearly catches from Mauguio, France, during the years 19861989. See text for explanations. “Total” includes official landings of eel and the landings of other species (by-catch) estimated by survey.

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Bayley (1988) noted that "as effort increases, which is normally accompanied by changes in gear, such as mesh-size reduction in gillnets, valuable species (often larger ones) become depressed, and a larger number of species become profitable". It is not clear how such evolution would affect the yield and the discards. There is also some evidence that rather large variations among yields reported from similar and geographically close lagoons may specifically result from differences in marketing possibilities. For example, Bouchereau et al. (1990) estimated that the commercial catches of crab (Curcinus mediterraneus) in Mauguio were 37 kg ha-' yr-' in 1989. Although this species is of low commercial value, a small market developed for the food industry, so that statistics concerning this species were therefore included in the estimate of the yield. In contrast, Ardizzone (1984) mentioned that 100 kg ha-' yr-' of C. mediterraneus were annually discarded in Burano (40) owing to a lack of market in Central Italy. In Marano/Grado (43, Italy), northern Adriatic, the yield of crab C. mediterraneus is reported 1 kg ha-' yr-' (Brambati et al., 1988), as commercialization is limited to molting individuals (softshell). Accuracy and precision of landing estimates are largely dependent upon sampling procedures because, in most systems but especially in large ones, coexisting fisheries have different purposes, use different gear and rely on different landing sites and commercial methods (Ewald, 1964; PereiraBarros, 1969; Durand et al., 1982; Balakrishnan Nair et al., 1983; Samarakoon, 1986; Madhusoodana et al., 1992; Maria Siluvai Raj et al., 1992; North Carolina Division of Marine Fisheries, 1996). For example, in Pamlico Sound (285, USA), the industrial seine fishery for Atlantic menhaden (Brevoortiu tyrannus) coexists with the shrimp trawling fishery, the gillnet fish fishery and the crab pot fishery. In these conditions, few procedures that are not mandatory will furnish accurate landing estimates. As a result, studies and surveys often focus on part of the fisheries, generally the commercial fish or shrimp landings, or both, depending primarily on their respective commercial or sport values. For the Cuban lagoons (262-267), the Texas lagoons (269-275, USA), and Barataria Bay (277, USA), this data-partitioning forced us to merge observations and statistics from different studies, that were often related to different years, with unknown consequences. Methods for estimating the yield in lagoons vary enormously; comparisons are therefore difficult. Two independent studies on the same entity are necessary to illustrate the variability affecting the estimations among studies. For the year 1989, Bouchereau et d. (1990) and Ruiz (1994) respectively estimated the fishery yield from the lagoon of Mauguio (30, France) to be 97 and 42 kg ha-' yr-', eel excluded. The estimations by Bouchereau et al. (1990) were derived from a year of survey at one landing harbor extrapolated to the whole lagoon, while Ruiz (1994) based his computations on

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confidential book-keeping records for tax purpose from a few fishermen. The same year 1989, landings of eel from the lagoon complex (Mauguio (30) plus Mkjean (not listed) plus Pkrols (not listed) plus Pierre Blanche/ Mourres/A’idolle (28) plus Arne1 (not listed)) were estimated to be 176 t by Ruiz (1994), and 231 t from voluntary declarations to the authorities (Bouchereau et al., 1990 and pers. comm.). It is easy to conclude from these observations that a (large?) part of the variability of the estimates of the fisheries yield between lagoons is based on the sampling/surveyingprotocols of the investigators. Official fishery landings are often available by administrative unit at the lowest geographic level, which may or may not accurately cover the system studied. Data may therefore combine information for both offshore and inshore commercial fisheries (Irby, 1974, for Choctawhatchee Bay, Florida, USA; Harris et al., 1983, for Charlotte Harbor and Lake Worth, Florida, USA). Landing statistics may also exclude industrial fisheries which land their captures in other harbors (Pamlico and Albemarle Sounds, 285 and 286, USA; Epperly and Ross, 1986). Finally, recreational activities and illegal fishing are rarely included in landing estimates. Likewise, underreporting, which seems common, and personal consumption, which is highly developed in some countries, are generally unaccounted for. In Mauguio (30, France), in contrast to other species, eel is exploited through a cooperative. In 1989, eel landings totalled 54% of the commercial catches estimated from voluntary declarations to the authorities, but accounted for only 24% of the commercial catches estimated by survey (Bouchereau et al., 1990 and pers. comm.). We suspect that the reported yield is generally underestimated, since the combined impact of the unreported activities listed above is clearly significant.

2.3. Some Observations on the Fisheries The use of different kinds of devices and practices may influence the volume of the catch. In fact, besides the use of “usual” active or passive gear such as trawl, seine, nets, and traps, production can be based on a variety of “alternative” methods. The interesting part about these methods is that they involve a limited technology and, most often, a high degree of cooperation among fishermen is required for a maximum return (see e.g. Ardizzone et al., 1988; Chauvet, 1988). When known, their existence is reported in Table 1 (see Section 3). Valliculture sensu luto, i.e. including fish reservoirs,janos and hoshas, is a technique intensively employed in some regions of the world (Jones and Sujansingani, 1954; De Angelis, 1960; Rowntree et ul., 1984; Toews and Ishak, 1984; Ardizzone et al., 1988; Chauvet, 1988). Valliculturing is the

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practice of extensive aquaculture from natural colonization, without external feeding or fertilizer. Target species are allowed to enter the lagoon area, generally as larvae or early juveniles, but prevented from leaving until having reached a commercial size. Communication between lagoon and ocean are limited to the period of recruitment. Yield enhancement is sometimes assured by importing larvae or juveniles from other areas, practice which is called fry stocking (D’Ancona, 1954; Ardizzone et al., 1988). Although we are not aware of any specific study on the biological effects of capturing fry in the natural environment, we are concerned this practice may be extremely damaging to the areas where the fry are caught. Moreover, because fry are generally protected, fry stocking has been associated with illegal fishing and smuggling through international borders. The presence of fish barrages, called bordigues in Mediterranean France, Algeria and Tunisia, pantenas in Spain, and lavorieri in Italy, is reported in Table 1 (see Section 3). Chauvet (1988) showed their presence was correlated with high fishery yields. These devices do not impede water movements between lagoon and sea and, thus, allow a natural colonization of the estuarine area. They are generally set to catch fish in their seaward migration. The use of fish barrages is generally seasonal, according to migration patterns. In La Spiaggia, Is Brebeis and Maestrale (50-52, Italy), their function is to bar fish from leaving the estuarine system (Rossi and Cannas, 1984), instead of capturing them. Brushparks furnish well protected areas which are fished from time to time (Buffle, 1958; Food and Agriculture Organization/United Nations Development Programme, 1971; Welcomme, 1972, Kapetsky, 1981). Two mechanisms of action have been recognized. In the short term, a few days to a few weeks, brushparks act as traps for fish seeking shelter. Using this principle, small versions of brushparks have been traditionally used in Madagascar (Lasserre, 1979b) and Sri Lanka (Senanayake, 1981; Samarakoon, 1986; Wijeyratne and Costa, 1987). In the longer term, larger brushparks evolve to provide a more productive environment by supporting increased primary production and providing an adequate substratum and protection for the eggs and the young individuals. The larger and more complex brushparks are BCnin’s acadjas (Welcomme, 1972). The effects of the introduction of exotic fish species into estuarine environments has, to our knowledge, never been generally assessed. There are two obvious similarities among the four described cases the authors are aware of. All instances concerned an euryhaline non-migratory, essentially herbivorous, cichlid species, that was accidentally or intentionally introduced into the fresh or brackish waters of tropical islands, namely Madagascar, Sri Lanka, Puerto Rico, and Cuba. Scientific names are kept as they appeared in the original works: Tilapia mossambica, Oreochromis mossambicus, and Sarotherodon mossambicus refer to the same species. This

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fish has been introduced in 67 countries, and nowadays the aquatic fauna of each of these four countries comprises between 22 and 24 exotic species (Welcomme, 1992). Coincidentally, in Egypt (Manzalah, Borullus, Edku and Mariut, 135-137, 139), Ivory Coast (Grand Lahou, EbriC, and Aby/ Ehy/Tendo, 152-154) and Benin (AhemC, Nokoue/Porto Novo, and Ouidah/Grand Popo/Aho, 156158), a sometimes large part of the yield depends upon the catches of endemic tilapia. With the exception of Edku, hypersaline during the dry season, these lagoons have above average to very high fishery yields. Note the coincident use of alternative fishery techniques (hosha in Manzalah, and acadjas in Benin’s estuaries) in several of these lagoons. This observation suggests that voluntary or involuntary introduction of tilapia species in lagoon environments can only be successful, in terms of yield, if the exploitation of the new resource is accompanied by profound changes in management. However, none of the four countries named above has a tradition of cooperation among fishermen (prerequisite for the use of acadjas) or of extensive aquaculture after manipulation of the environment (hosha). In absence of management, some evidence for increased yields of the local fisheries conflicts with the sharp decline or collapse of the catches of other species sometimes reported in years following the introduction. At least two cichlid species, Tilapia rendalli and Oreochromis mossambicus, were successfully introduced in Madagascar’s Pangalanes (162; Moreau, 1987). Both species now total 12% of the catches in the lower Pangalanes South of Andevoranto. Moreau (1987) noted that present-day catches in the Pangalanes (8kgha-’yr-’) are much lower than those in 1967 (43 kg ha-’ yr-’). Tilapia mossambica was successfully introduced into the freshwaters of Sri Lanka in 1952 (Fernando, 1965). Colombo Lake (173) was stocked that year with 200 T. mossambica fingerlings, then restocked in 1953 with 2250 T. mossambica and 500 Cyprinus carpio (Mendis, 1964). This highly artificial lagoon was isolated from the sea by a gate to Colombo’s harbor, had concrete walls, and was rain-fed (Mendis, 1964). Colombo Lake’s yield of fish reportedly increased to 2200 kg ha-’ yr-’ by 1957, with tilapia as the dominant species in the catches (Mendis, 1964). No mention was made of C. carpio. We were unable to gather further information on the changes affecting this fishery, although Mendis (1964) indicated that the yield remained at this level through the year 1963. There is some evidence that this exceptional situation did not continue. The lagoon is now extremely polluted and silted (Scott, 1989). More recently, De Silva and De Silva (1984) listed Sarotherodon mossambicus in the fish fauna of Sri Lanka’s lagoons and estuaries. This species was restricted to freshwater areas and was reported as a minor species. Similarly, neither Wijeyaratne and Costa (1987) nor Jayasuriya (1985) reported a significant presence of tilapia in the fishery landings from httalam and Negombo (174 and 175, Sri Lanka). Small S.

CONSTRAINTS ON COASTAL LAGOON FISHERIES

85

mossambicus (10.5 cm mean length) were reported in the by-catch from kralls in Rekawa (176, Sri Lanka; Parakum Maitipe and De Silva, 1986). Tilapia mossambica was introduced in Puerto Rico during the late 1950s (Erdman, 1967, in Burger et al., 1992). This species now composes 55-79% of the total number of fish caught in three Puerto Rican estuaries (Burger et al., 1992). In 1980, tilapias were already extremely abundant in one of these three lagoons (Puerto Rico Department of Natural Resources, 1981, in Burger et al., 1992). In the Humacao lagoons (not listed), tilapias are small and highly diseased (A. Zerbi, pers. comm.). Introduction of tilapias in the freshwaters of Cuba began during the 1960s, and, in 1973, several species were seeded in the region of the Tunas de Zaza (Garcia et al., 1984). Subsequently, coastal lagoons were colonized from the freshwater environments. Although no exotic cichlid species was reported present in the lagoons of the Tunas de Zaza in 1975-76 (Gonzalez Sanson et a/., 1978), the commercial fishery for Oreochromis aureus began three years later in 1979 (Garcia et al., 1984). A 100% increase in the annual yield of the fisheries for the Tunas de Zaza lagoons was reported in the years following (Gonzales Sanson et al., 1985). This increase was a result of large catches of 0. aureus. These authors reported a concomitant decrease in the catches of the native mugilids. 2.4. Conclusions

We will not attempt to prove that our data set is representative of all lagoons of the types we selected for this study. We have found examples of saline lagoons where the only apparent fishery was of brine shrimp, sometimes purposely introduced (Nash, 1978). Those are not included in the compilation. Neither are anchialine lagoons which communicate with the ocean through subterranean openings or fissures (Thomas et al., 1991), and which are generally too small to sustain commercial fisheries. In spite of such restrictions to the compilation, or because of them, we acknowledge that it has been impossible to find one lagoon that was specifically not fished because of its absolute lack of fish or macroinvertebrate productivity. This result is not surprising because, for example, in the few cases negative yields were reported (e.g. West et al., 1985), the lagoons mentioned were too small, less than 1.5 km2, for fisheries to operate economically. As a matter of fact, because null yields are rarely reported and because most reported yields might concern relatively good yielders, the assumption that the data here collected accurately represent all the lagoons can not be asserted. Therefore, our sample represents more accurately a subset of fished lagoons. For this subset of fished lagoons, the overall poor quality of the landing records questions the necessity and the possibility of completing a

86

J.-C. JOYEUX AND A. B. WARD

meaningful analysis. The reason why such an analysis should be performed is that we would not learn anything in waiting for a global improvement of the landing statistics . . .. Thus, the actual question is to know if such an analysis can be performed. The answer is, without doubt, positive. Imprecision of the records and unreported catches are negligible in regard to the variation in yield among lagoons. Including such yield-values certainly increases statistical noise, but does not invalidate positive results. Moreover, when such a situation is known, remedial statistical actions are available (see next section). Before closing this section, a final point needs to be covered. A low fishery yield of fish and crustaceans does not necessarily mean that a lagoon does not produce human food. Appendix 1 gives the yields of bivalve and gastropod fisheries and aquaculture in a subset of 109 lagoons. Comparisons with the data for the fisheries yield displayed in Appendix 2 show that a low yield of fish and crustaceans can be largely compensated for by an important mollusc fishery or aquaculture yield. The best example, almost a caricature, is Barnegat Bay (287, USA) where the fish yield (4 kg ha-' yr-') is insignificant compared with the clam yield (6740 kg ha-' yr-', shells on; Hillman and Kennish, 1984). This stresses the fact that our compilation does not concern the human-consumable biological yield of lagoons per se.

3. MATERIAL AND METHODS 3.1. Data Description and Collection

Statistical approaches using time series for selected lagoons only indicate proximate causes for production (Copeland, 1966; Meeter el al., 1979; Texas Department of Water Resources, 1979, 1981; Browder and Moore, 1981). Therefore, in order to determine the ultimate, global factors affecting fishery production, each lagoon was considered as one entity, independently of the number of years of record available. Data were taken from published works, unpublished reports and personal communications by diverse authors and officials. Name, location of lagoons, years of records, short comments on the fisheries, and bibliographic references for the data are given in Table 1. The complete data set details information on 292 lagoons, and is given in Appendices 2 and 3. The data set analyzed contained 274 lagoons. Matching the records from different periods was a challenge, and it was impossible to restrict the compilation to perfect matches. Hence, we often merged data not only from different years but also from several studies, under the unverifiable assumption that each of them would be representative. When data from different

Table I Names and locations of lagoons, bibliographic references, years of study and comments for the yield of lagoon fisheries around the world. The categories listed under ‘sea area’ refer to the fishing areas of the world defined by the F A 0 for statistical purposes. References between parentheses were used for variables other than the yield of the fishery. The years of study refer to the yield. In “Comments”, “Bordigue” indicates the presence of a fish barrage. The appellation “Reservoir” was attributed when water movements between sea and lagoon are regulated by gates allowing a total closure. The practices of valliculture and fry stocking are recorded. The annotation “Closure” indicates that part or all of the lagoon is closed to commercial fishing, either permanently or occasionally. The designation “Complex” was attributed when the lagoon is composed of morphologically well individualized subunits. The position of subunits that do not directly communicate with the sea is indicated as “second line” and “third line”.

SEA AREA country

[State/Region] No.

Lagoon

References

Years of study

ATLANTIC NORTH EAST Poland 1 Zalew Szczecinski = Gr. Oderhaff Winkler, 1990 Germany 2 Peenestrom Winkler, 1990; Rechlin, pers. comm. 3 Kleiner Jasmunder Bodden Noack, 1978 in Winkler, 1990; (Rechlin, pers. comm.) 4 DarD-Zingster-Bodden West Winkler, 1990; (Nausch and Schlungbaum, 1991; Rechlin, pers. comm.) 5 DarD-Zingster-Bodden East Winkler, 1990; (Nausch and Schlungbaum, 1991; Rechlin, pers. comm.) 6 Greifswalder Bodden Winkler, 1990; Rechlin, pers. comm. Winkler, 1990; Rechlin, pers. comm. 7 Kleines Oderhaff 8* Stelasund/Kiibitzer Bodden Rechlin, pers. comm. Rechlin, pers. comm. 9“ Wismar-Bucht 10“ Unterwarnow/Breitling Rechlin, pers. comm.

Comments

75-86 77-86, recent 65-76 77-86

Second line

77-86 76-85, recent 75-86, recent 91 91 91

Complex Complex

Table 1 (continued)

00 00

SEA AREA country

[State/Region] No.

Lagoon

References

11 12

Schlei Fjord Neustaderbinnenwasser

Nauen, 1984 Nauen, pers. comm. to Mr. Kapetsky

71-80 73

FranCe 13

Etangs de Graveyron

Labourg, 1976; Stequert, 1972 in Lasserre et ul., 1976 Labourg, 1976; Labourg, 1976;

74

Reservoir

74 74

Reservoir Reservoir

14 Etangs de Certes 15 Etangs de 1'Escalopier MEDITERRANEAN AND BLACK SEA Spain 16 Albufeira de Valencia 17 Encayissada 18

Canal Vell

19

Tancada

20

Les Olles

21 22

Mar Menor Soleta

=

Goleta

Years of study

San Feliu, 1973 67/68, 69/70 Demestre et ul., 1977, 1989; 66-86 (Comin, 1982; De Sostoa and De Sostoa, 1985) 6686 Demestre et al., 1977, 1989; (Comin, 1982; De Sostoa and De Sostoa, 1985) Demestre et al., 1977, 1989; 6686 (Comin, 1982; De Sostoa and De Sostoa, 1985) Demestre et al., 1977, 1989; 66-86 (De Sostoa and De Sostoa, 1985) Arnal and Guevara, 1975 67-73 Amanieu and Lasserre, 1981 Unknown

Comments

Bordigue Bordigue Only the most valuable Bordigue species were recorded Bordigue Bordigue

France 23 24 25 26 27

Lapalme Ayrolle-Campignol Prkvost Gruissan Sakes-Leucate

28

Pierre Blanche/Mourres/A'idolle

29

Thau

30

Mauguio

31

Berre

=

Etang de I'Or

Bourquard, 1985 Bourquard, 1985 Bourquard, 1985 Bourquard, 1985 Bourquard, 1982; (Arnaud and Raimbault, 1969; Boutikre e f al., 1982) Le Corre and Autem, 1982; (Amaud and Raimbault, 1969) Jouffre and Amanieu, 1991; (Amaud and Raimbault, 1969) Quignard et al., 1989; Bouchereau et al., 1990; Bouchereau, pers. comm.; (Amaud and Raimbault, 1969; CEMAGREF, 1989) Amanieu, 1973; (Arnaud and Raimbault, 1969)

84 83-84 Unknown 84 78

Frisoni, 1982; (Colombo, 1977) Frisoni, 1981 Frisoni, 1981

78

8 1/82

Complex

Recent. <91 86-89

71

[Corsica]

32

Biguglia

33 34 IMY 35

Diana Urbino

36

Orbetello

Sacca di Scardovari

Rossi, 1981; Rossi, pers. comm. to Mr. Kapetsky Ardizzone et al., 1988

Bordigue

78 78 68-77

Fry fishing

62-82

Fry stocking, bordigue

(D

Table I (continued)

0

SEA AREA country

[State/Region]

No.

Lagoon

37

Paola

38

Lesina

39

Varano

40 41

Burano Valli di Comacchio

42 43

Aquatina GradoIMarano

44 45 46 47 48

=

References Sabaudia

Costa and Minervini, 1982; (Karvounaris, 1963) Rossi and Villani, 1980; Ardizzone et al., 1988; (Rossi and Colombo, 1976; Lumare, 1984) Lumare, pers. comm. to Mr. Kapetsky; (Marolla,

Years of study

Comments

63-79

Bordigue

5 1-84

Valliculture

81-83

1980)

Recent, <84

Monaci

Ardizzone, 1984 Ardizzone et al., 1988; (Colombo, 1972; Rossi and Colombo, 1976; Barnes, 1980) Rossi and Corbari, 1982 Zerbinato, 1981 in Ardizzone et al., 1988 Ardizzone and Corsi, 1985

Fondi Fogliano Valle Nuova Valli di Polesine

Ardizzone and Rossi, 1983-84 Ardizzone, 1984 Rossi and Pappas, 1979 Rinaldi, 1960

56-80

1781-1982 7679 < 81 83-84

Recent, < 84 50-78 54-56

Reservoir, valliculture Valliculture, complex Fry stocking, bordigue Bordigue Valliculture Valliculture, 30 valli

[Sardinia]

49 50 51 52 53 54 55 56 57 58 59 60 61 62 63

64 65 Greece 66 67 68 69 70 71 72 73 74 75 76 77

Is Brebeis

....

Keramoti Eratino = Karassu Agoulinitsu Mesolon@-Etoliko Vulkaria Mouria Vasova Pappas Prokopos Kutichi = Kotihi Monolimni = Limni Paloukion Drana = Drakontos

Ardizzone et al., 1988 Rossi and Cannas, 1984 Rossi and Cannas, 1984 Rossi and Cannas, 1984 Cottiglia, 1981 Cottiglia, 1981 Cottigha, 1970 Farris et al., 1978 Cottiglia, 1970 Cottiglia, 1970 Cottiglia, 1970 Cottiglia, 1970 Cottiglia, 1970 Cottigha, 1970 Cottiglia, 1970 Cottigha, 1970 Cottiglia, 1970

57-78 80 80 80 67-77 68-77 68 72, 75-77 68 68 68 68 68 68 68 68 68

Tsiokas, pers. comm. Tsiokas, pers. comm. Pillay, 1967 Tsiokas, pers. comm. De Angelis, 1960 Pillay, 1967 Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Pillay, 1967 Tsiokas, pers. comm. Tsiokas, &rs. comm.

Recent, <92 Recent, <92 63/64 Recent, <92 <60 63/64 Recent, <92 Recent, <92 Recent, <92 61/63 Recent, <92 Recent, <92

Fry stocking Bordigue Second line Common Third line bordigue

Bordigue Bordigue

Drained now Bordigue Bordigue Drained now Bordigue

Table I (continued)

SEA AREA country [State/Regioa]

No. 78 79 80

81 82 83 84 85 86 87 88 89

90 91 92 93 94 95 96 97 98 99 100 101 102 103

Lagoon

References

Years of study

Lafra/Lafmda Limni Mavrolimni xyrolimni Sajada (3 lagoons) Lefcas Chalkiopoulou Gavojiani Pogonitza Mazoma Tsopeli Tsukalio Logam Kofta-Paliobouca Solinari Thermisia Vivari-Drepano Sabariza Agrilos Voda Lutsa-Papadia Bastia Agiasma = Kuburnu Kutavos Paleopotamos Ghialova

Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Ananiadis, 1984 Ananiadis, 1984 Ananiadis, 1984 Ananiadis, 1984 Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm.

Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <84 Recent, <84 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92

Comments

Complex

Bordigue

104 Vistonida-Lagos 105 Avlax Papaliu 106 Vathi 107 Aliki 108 Mana 109 Ptelea 110 Elos Turkey (Aegean Sea) 111 Bafa 112 Karine 113 Pamukulu/Gala/Dalyan/ Biiciirmene 114 Homa

Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm. Tsiokas, pers. comm.

Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92 Recent, <92

Uyguner and Gozenalp, 1959 Uyguner and Gozenalp, 1959 Cataudella, 1983a

< 59 < 59

Recent, <83

Cataudella, 1983a; (Cataudella and Ferlin, 1984)

Recent, < 83

Tuzlu Armak/Hoyrat/Tahir Hersek Poyraz Biiyuk Cekmece Kiiciik Cekmece Tuzla

Cataudella, 1983a Uyguner and Gozenalp, Uyguner and Gozenalp, Uyguner and Gozenalp, Uyguner and Gozenalp, Uyguner and Gozenalp,

Recent, <83 < 59 < 59 < 59 < 59 < 59

Karabogaz Tuzlu/Liman/Balik

Uyguner and Gozenalp, 1959; (Barnes, 1980) Uyguner and Gozenalp, 1959

< 59

Uyguner and Gozenalp, 1959

< 59

Uyguner and Gozenalp, 1959 Cataudella, 1983a Uyguner and Gozenalp, 1959

< 59 < 59

Bordigue

Turkey (Marmara Sea)

115 116 117 118 119 120 Turkey (Black Sea) 121

122 Koca = Semenlik Dumanli Turkey (Mediterranean Sea) 123 Yumurtalik/Yelcoma/ Hurma-Bogazi 124 Karatas/Akyatan 125 Paradeniz 126 Ova

1959 1959 1959 1959 1959

< 59

< 59

Bordigue Bordigue

(0

Table I (continued)

P

SEA AREA country

[State/Regionl

No.

Lagoon

References

Years of study

Comments

127

Kiiycegiz

Cataudella, 1983a

Recent, <83

Bordigue

128 129

Dniestrovski liman Dneprovsko-Bugsky Liman

Recent, <92 Recent, <89

130

Berezansky Liman

131

Tiligulsky Liman

132

Khadzhibeisky Liman

Zaitsev, pers. comm. Zaitsev, 1989 (Zaitsev, pers. comm.) Mironov 1990 (Zaitsev, pers. comm.) Mironov, 1990 (Zaitsev, pers. comm.) Mironov, 1990 (Zaitsev, pers. comm.)

Ukraine (Black Seal

Recent, <90 Recent, <90 Recent, <90

Ukraine (Caspian *) 133 EgYPt 134

Bardawil

135

Manzalah

Sivash =

Sirbonis

Zaitsev, pers. comm.

Recent, <92

Pisanty, 1981; (Knungalz, 1980)

72-78

Bishai and Yosef, 1977; MacLaren, Inc., 1981; Reid and Rowntree, 1982; (Halim and Guerguess, 1981 and 1982; Rowntree et al., 1984; Lemoalle and Saad, 1987)

63-68, 80, 82

Tilapia Hosha

Endemic

... ...

...

136

137

Borullus

=

Brollus

Edku

138

Port-Fouad = El-Mallaha

139"

Mariut

MacLaren, Inc., 1981; Reid and Rowntree, 1982; (Halim and Guerguess, 1981 and 1982; Rowntree et al., 1984; Lemoalle and Saad, 1987) Reid and Rowntree, 1982; Shaheen et al. (MS); (Halim and Guerguess, 1981, 1982; Banoub, 1983; Rowntree et al., 1984) Kerambrun, 1986; Vanden Bossche and Bernacsek, 1991 Vanden Bossche and Bernacsek, 1991

... ...

... I

Recent, < 82

1 76, 77, 82 72, 75

Endemic Tilapia

81

Bordigue

Algeria

140

Mellah

141" Tunisia 142

Oubeira Ischkeul

Cataudella, 1983b; (Chassany-de Casabianca et al., 1981) Cataudella, 1983b =

Tindja

144

Bizerte = Menzel = Abderahmen Ghar el Melah = Porto Farina

145

Tunis

146

Bou Grara

143

=

Goulette

81

Pillai, 1975; Abdelmajid, 1979; Vidy, 1983; (Zaouli, 1975) Pillai, 1975; Abdelmajid, 1982

61-78

Pillai, 1975; Abdelmajid, 1982; (Savourk, 1977) Pillai, 1975; Abdelmajid, 1982; (Belkhir and Hadj Ali Salem, 1982; Carbonel and Pujos, 1982; Zaouli, 1987) Zaouali, pers. comm. to Mr. Kapetsky

61-63, 67-76, 78 61-78

Bordigue

Recent

Bordigue

Second line, bordigue

73, 77

Tilapia

Table I (continued)

z

SEA AREA coontry

[State/Region] No.

Lagoon

References

147

Kelbia Sebkha

Pillai, 1975; Abdelmajid, 1982

148 149 Moroeco 150

El Biban Khniss = Lac de Monastir

Lemoalle and Vidy, 1984 Abdelmajid, 1979; (Vallet, 1977)

Nador = Mar Chica de Menilla Brunel, 1985 = Sebkha bou Areg ATLANTIC CENTRAL EAST Morocco Merga Zerga = Moulay Bou Bayed et ul., 1987 151 Selham Ivory coast 152 Grand Lahou = Tagba/Make/ Berron, 1980; (Dufour and Tadio Durand, 1982; Durand et ul., 1982; Dufour, 1987a) 153 Ebrit Gerlotto and Briet, 1976; (Dufour and Durand, 1982; Durand et ul., 1982; Dufour, 1987a) Charles-Dominque et al., 1980; 154 Aby/Ehy/Tendo Durand et ul., 1978; Durand, pers. comm. to Mr. Kapetksy; (Dufour and Durand, 1982; Durand et ul., 1982; Dufour, 1987a)

Years of study

Comments

61, 66, 68-70, 77-78 62-82 71-78

Bordigue Bordigue

Recent. <85

Recent, <87

only eel

69/70

Onlyhfish, Endemic complex ...

7577

Some brush parks

... ... ... 1..

79

Only finfish, complex

... ... ... ... ... Tilapia

Ghana

155

Sakumo-Tema

Bhin 156

Ahemt

157

NokouelPorto Novo

158

=

=

Sakumo I1

Ayeme

Ouidah/Grand Popo/Aho

Pauly, 1976; (Kwei, 1977; Biney, 1982 and 1986; Gordon, 1987)

71

Only !infish

FAO/UNDP, 1971; (Maslin and Bouvet, 1986) Centre Technique Forestier Tropical, 1969; (Maslin and Bouvet, 1986; Colleuil and Texier, 1987) FAO/UNDP, 1971

69/70

Acadjas

59

Acadjas, complex

69/70

Acadjas, complex

Togo 159

Lac Togo

Vanden Bossche and Bernacsek, 1991; (Millet, 1987)

83

Some brushparks

Nigeria 160

Lagos/Lekki

FAO, 1969; (Balogun, 1987; Dufour, 1987b)

65

Only finfish, brushpark, complex

161" INDIAN WEST Madagascar 162

Ogun

Kapetsky, 1981

80

Pangalanes

Lasserre, 1979b

61-78

163 164 165

Anony Masianaka Taolanaro Lagoons

Lamarque, 1957 in Lasserre, 1979b 53 Recent, < 87 Moreau, 1987 Recent, <87 Moreau, 1987

Emt 166"

Um El Rish

Vanden Bossche and Bernacseck, 1991

79

Some brushparks, complex Only finfish

Complex

Endemic

... ... ... ...

... ... Tilapia

Table I (continued) (D

SEA AREA

a0

coontry [State/Region]

No.

References

Lagoon

Years of study

Comments

Complex

India [BeralaI 167

Cochin

George and Sebastian, 1972; Madhupratap et al., 1977; (Gopinathan et al., 1984; Sosamma and Mathew, 1989)

?

Mandapam EMore Backwater Kovalam Backwater Kazhuveli Backwater Vellar/Coleroon Estuaries/ Killai Backwaters = Port0 Novo

Tampi, 1959; (Tampi, 1969) Padalkar, pers. comm. Padalkar, pers. mmm. Padalkar, pers. comm. Venkatesan, 1969 in Jhingran and Gopalakrishnan, 1973; (Krishnamurthy, 1971; Krishnamurthy and Sundaraj,

58

Kerala Backwaters

=

[TamilNadu] 168 169 170" 171" 172

Recent, <93 Recent, <93 Recent, <93

Only prawn

67/68

Complex

Mendis, 1964; (Costa, 1968)

57

Tdapia introduced

Jayasuriya, 1985; (Durairatnam,

82

1973

Sri Lanka 173

Colombo Lake

174

Puttalam

=

Beira

1963) 175

Negombo Lake

Wijeyaratne and Costa, 1987; (De Silva and Silva, 1979)

Recent, <87

176

Rekawa

Parakum Maitipe and De Silva,

Recent. < 89

1986

Only finfish, Some brushparks only shrimp

Jayasekara and Jayakody, pers. comm. Jayasekara and Jayakody, pers. corn.

91/92

Pulicat

Jhingran and Gopalakrishnan, 1973; (Chacko et al., 1953)

67-69

Godavary Estuary

Laltitha Devi, 1988

79-84

Complex

Chilka

Jhingran and Natarajan, 1969; (Banerjee and Roychoudhury, 1966; Kowtal, 1967)

57-65

Janos

Dept. Fish. Wildl. West. Aust., 1981; (Black et ul., 1981; Birch, 1982; Hodgkin and Birch, 1982) Dept. Fish Wildl. West Aust., 1981; (Hodgkin and Lenanton, 1981) Dept. Fish. Wildl. West. Aust., i981; (Hearn et ul., 1984) Dept. Fish. Wildl. West. Aust., 1981; (Lenanton and Hodgkm, 1985) Dept. Fish. Wildl. West. Aust., 1981; (Hodgkin and Lenanton, 1981; Lenanton and Hodgkin, 1985)

76/77, 79/80

Closure, complex

76/77, 79/80

Closure

76/77, 79/80

Closure

76/77, 79/80

Closure

76/77, 79/80

Closure

177"

Malala

178"

Mawalla

INDIAN EAST India mamil Nadu and Andhra Pradesh] 179 [Andhra Pradesh] 180

I0-l

181

Australia western Australia] 182 Peel-Harvey

183

Oyster Harbour

184

Leschenault

185

Broke Inlet

186

Wilson Inlet

=

Brook's Inlet

91/92

2

Table 1 (continued)

0

SEA AREA coontry (State/Regioo]

No.

Lagoon

References

Years of study

Comments

187

Irwin Inlet

R p t . Fish. Wildl. West. Aust.,

76/77, 79/80

Closure

1981

wictoria] 188 189

Gippsland Lakes Port Phillip Bay

190

Western Port

Beinssen, 1978; (Kjerve, 1986) 73/74 Beinssen, 1978; (Poore and 73/74 Kudenov, 1978; Smith and Maher, 1984) Beinssen, 1978; (Hams et al., 1979) 73/74

Complex

West et al., 1985 West et al., 1985 West et al., 1985

Closure Closure Closure

p e w Sooth Wales] Wallagoot Lake Merimbula Lake Pambula Lake PACIFIC CENTRAL WEST Thailand 194 Lake Songkhla 191 192" 193

72-82 72-82 72-82

Recent, <85 Sinclair, 1985 in Scott, 1989; (Limpadanai, 1980; Kjerfve, 1986)

With small aquaculture

Vietnam Tam Giang Lagoon Nai Swamp PNW (PACIFIC NORTH WEST) Japan 197 Kitaura/Kasumigaura 195 196

Scott, 1989 Scott, 1989

Recent, <88 Recent, <88

Seki et ul., 1981; (Yamamoto, 1981) 54-63

Only finfish Only finfish

Complex

PACIFIC SOUTH WEST Australia [New South Wales] 198 Clarence River 199 Woolgooda Lake 200 Innes/Cathie Lakes 20 1 202 203

Wallis Lake Smiths Lake Myall Lakes/Myall River

204

Lake MacQuarie

205

Tuggerah Lakes

206" 207

Botany Bay/Georges River Lake Illawarra

208 209 210 21 1 212 213 214

St George Basin Swan Lake Lake Conjola Bum1 Lake Durras Lake Clyde River Coila Lake

215

Tuross Lake

216 217 218

Lake Brou Lake Dalmeny Wagonda Inlet

West et al., 1985 West et al., 1985 West et al.. 1985

72-82 72-82 72-82

West et al., 1985 West et al., 1985 West et al., 1985; (Roy aud Peat, 1976) West et al., 1985; (Baas Becking, 1959; Baas Becking et al., 1959; Spencer, 1959; Batley, 1987) West et al., 1985; (Powis and Robinson, 1980; Collett et al., 1981) West et al., 1985 West et al., 1985; (Roy and Peat, 1976) West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985; (Roy and Peat, 1976) West et al., 1985; (Roy and Peat, 1976) West et al., 1985 West et al., 1985 West et al., 1985

72-82 72-82 72-82

Closure Closure Closure, complex Closure Closure Closure

72-82

Closure

72-82

Closure

72-82 72-82

Closure Closure

72-82 72-82 72-82 72-82 72-82 72-82 72-82

Closure Closure Closure Closure Closure Closure Closure

72-82

Closure

72-82 72-82 72-82

No closure No closure Closure

Table 1 (continued)

SEA AREA COMW [State/Region] No.

Lagoon

References

219 220 22 1 222 223 224 PACIFIC NORTH

Corunna Lake Tilba Tilba Lake Wallaga Lake Murrah Lagoon Wapengo Lagoon Nelson Lagoon EAST

West et West et West et West et West et West et

lo~egonl 225

Tillamook Bay

226

Netarts Bay

227

Coos Bay

Years of study

Comments

72-82 72-82 72-82 72-82 72-82 72-82

No closure No closure Closure Closure Closure No closure

Percy et al., 1974; (Burt and McAllister, 1959; NOAA, 1990) Percy et al., 1974; (NOAA, 1990)

7&7 1

Percy et al., 1974; (Burt and McAllister, 1959; Proctor et al., 1980; NOAA, 1990)

70-7 1

OdY commercial crab OdY commercial crab Commercial

Guzman Arroyo, 1987

76, 81

al., 1985 al., 1985 al., 1985 al., 1985 al., 1985 al., 1985

USA

PACIFIC CENTRAL EAST Mexico [Goerrem) 228" Tres Palos

70-7 1

OdY

Macrobrachium OdY

[Sonoral 229

Bahia de Lobos

68-73

Only shrimp

68-73

Only shrimp

230 [Sialoa] 23 1

Bahia de Guasimas

Garduiio Argueta, 1974; (Gilmartin and Revelante, 1978) Garduiio Argueta, 1974

Laguna Ceuta

Cervantes Castro, 1984

Recent, <84

232

Ca'imanero-Huitzache

Edwards, 1978; (Barnes, 1980)

73/76

233

Castro-Ortiz and Sanchez Rojas, 1976 Castro-Ortiz and Sanchez Rojas, 1976; (Gilmartin and Revelante, 1978)

65-76

234

Bahia de Altata-Ensenada del Pabellon Bahia Santa Maria La Reforma

Oyster included only shrimp, complex Only shrimp

65-76

Only shrimp

lrvrnayitl 235

Agua Brava

Cervantes Castro, 1984; (Flores-Verdugo et al., 1990)

Recent, < 84

Oyster included, complex

[Chiapasl 236

Toro/Joya/Buena Vista

73-76

Complex

237 238 239 240

Colorado Rio San Pedro San Blas Chila/Penitas

Huerta Maldonado, 1980; (Mee, 1975) Chapa Saldana et al., 1977 Chapa Saldana et al., 1977 Chapa Saldana et al., 1977 Chapa Saldana et al., 1977

72/73 72/73 72/73 72/73

only shrimp Only shrimp Only shrimp Only shrimp, complex

[Salvador] 24 1

Jiquilisco

78, 90

242"

Barra de Santiago

243a

Estero de Jaltepeque

El Salvador, 1979; CENDEPESCA, 1991 (Ibarra, pers. comm.); (Gierloff-Emden, 1982) CENDEPESCA, 1991 (Ibarra, pers. comm.) CENDEPESCA, 1991 (Ibarra, pers. comm.)

90 90

-L

0

w

Table I (continued)

a

0 P

SEA AREA country

IState/Regionl No.

Lagoon

ATLANTIC SOUTH WEST UWPY 244 Rocha Castillos 245 Umguay/Brd 246 Mirim

Brd [Rio Grande do Sd] Dos Patos 241

References

Years of study

Comments

Mistakidis, 1964 Mistakidis, 1964

61-63 61-63

o n l y shrimp only shrimp

Castello, pers. comm.to Mr. Kapetsky

75, 77-79

Castello, pers. comm. to Mr. Kapetsky; (Castello, 1985; Chao et al., 1985; Kjerve, 1986)

70-79

[Saota Catarina]

248 249

Mirim/Imarui/Santo Antonio Concei@o Lagoon

Tremel et ul., 1975 Sierra de Ledo et al., 1985; (Assum&o et al., 1981)

70-73 70-81

Complex

[Siio Paulo] 250

Iguape-Cananeia

Sant'Ana Diegues, 1976; (Miyao et al., 1986; Schaemer-Novelli et al., 1990)

50s

only shrimp

Araruama

Slack-Smith et al., 1977

73-74

only shrimp

Mundau

Pereira-Barros, 1967, 1969

67, 68

[Rio de Janeiro]

251

ATLANTIC CENTRAL WEST Venezuela 253 Unare

Khandker and Lares, 1968; Padron, 1982; (Posewitz, 1969; Okuda, 1981) Padron, 1982; (Posewitz, 1969) Nemoto, 1971; (Ewald, 1964; Rodriguez, 1969 and 1973) Gamboa ef al., 1971; (Okuda, 1981)

62-78

Cienaga Grande de Santa Marta

Colombia, 197@-74;(kvarez Leon and Blanco Racedo, 1985)

7s74

Mexico (Gulf of Mexico) [Campe&e] 258

TCrminos

Yaiiez-Arancibia and AguirreLebn, 1988; (Signoret, 1974; Yiiiez-Arancibia and Day, 1982; GurtiCrrez-Estradas and Castro del Ric, 1988)

82-85

weracruz] 259

Tamiahua

68-7 1

260

Alvarado

Garcia, 1975; Anon., 1980; (Contreras, 1984; Deegan et al., 1986) Cervantes Castro, 1984; Villalobos et al., 1969; Deegan et al., 1986; Britton and Morton, 1989)

pamadipas] 26 1

Laguna Madre de Tamaulipas

Martinez Mata, 1980; (Deegan et al., 1986)

77

254 255

Piritu Maracaibo

256

Taracarigua

Colombia 257

67-77 70 64-66

Recent, <84

Oyster included, complex

--L

Table I (continued)

0 b)

SEA AREA counbry

Years of study

[State/Region] No.

Lagoon

References

Cuba 262 263 264 265 266 267"

El Basto Lugones Toletes Caney Las Caobas Ciego

Fictive year during the period 1972-75

Puerto R i a 268

Godinez, 1974; Alvarez Lajonchere, 1978; Bdez and Alvarez Lajonchere, 1980; Baez Hidalgo and Alvarez Lajonchere, 1982; Baez Hidalgo, pers. comm.; (Lalana Rueda et al., 1980; G o d l e z Sanson, 1983; G o h l e z Sansdn and Aguilar Betancourt, 1984; Lalana Rueda and G o h l e z Sanson, 1985)

Joyuda

Lopez et al., 1988; (Negrbn et al., 1982; Deegan et af., 1986; Stoner, 1986)

Recent, < 88

USA (Gulf of Mexico) lTex=l 269

Laguna MadreIBaffin Bay

Hefferman and Green, 1977; (Texas 68-73, Dept. Water Res., 1983; Deegan 75/76 et al., 1986; Britton and Morton, 1989)

Comments

Maximum, oyster included

Commercial + sport

270

27 1

272

273

274

68-73, 5/76

Commercial + sport

68-73, 74/75

Commercial + sport

68-73, 74/75

Commercial + sport

65-74, 75/76

Commercial + sport

68-73, 75/76

Commercial + sport

Breuer et al., 1977; (Louisiana S. Univ., 1970; Armstrong, 1982; Deegan et al., 1986; Britton and Morton, 1989)

66-75, 75/76

Commercial + sport

62-76

OdY commercial shrimp OdY commercial

Breuer et al., 1977; (Texas Dept. Water Res., 1981a; Armstrong, 1982; Deegan et al., 1986; Britton and Morton, 1989) Aransas/Copano/Espiritu Santo Hefferman and Green, 1977; (Texas Dept. Water Res., 1981a; Bays/Green Lake = MissionArmstrong, 1982; Deegan et al., Aransas and Nueces 1986) Hefferman and Green 1977; (Texas San Antonio/Mesquite Bays = Dept. Water Res., 1980b; Guada1upe Armstrong, 1982; Deegan et al., 1986) Breuer et al., 1977; (Texas Dept. MatagordalEast Matagordal Water Res., 1980a; Armstrong, Lavaca Bays = Lavanca-Tres 1982; Deegan et al., 1986; Britton Palacio and Morton, 1989) Breuer et al., 1977; (Texas Dept. Galveston Bay = Trinity-San Water Resources, 1981b; Jacinto Armstrong, 1982; Deegan et al., 1986)

Corpus Crisiti/Nueces Bays

ITexsl

Louisiana] 275

Sabine Lake

[Louisiana] 276

Calassieu Lake

Texas Dept. Water Res., 1979; (NOAA, 1985, 1990)

277

Barataria/Caminada Bays

59-72, c. 70 Barrett and Gillespie, 1973; Day et al., 1973 in Nixon, 1981; (Conner and Day, 1987; Gibson, 1991)

=

Sabine-Neches

2

Table 1 (continued)

0 Q)

SEA AREA country [State/Region]

Years of study

Comments

No.

Lagoon

References

(Alabama] 278

Mobile Bay

279

Little Lagoon

Swingle, 1976; (Deegan el af., 1986; 64-72 NOAA, 1990) Swingle, 1976 64-72

OdY commercial OdY commercial

US Dept. Interior, 1970 in Nixon, 66 1981; (Myers and Iverson, 1981; Livingston and Kitchens, 1984) Tampa/Old Tampa/Hilsborough Lewis and Estevez, 1985; (NOAA, 52-78 Bays 1985)

OdY commerical fish

Worida (Gulf of

Mexico)] 280 28 1

Apalachicola Bay/St. Vincent Sound/East Bay

M Y commercial fish Part commercial fish

282

Boca Ciega Bay

Hutton et al., 1956

54

283

Charlotte Harborpine Island Sound/San Carlos Bay

Taylor, 1974

< 74

M Y commercial fish

Halifax = Mosquito Lagoon/ Indian/Banana Rivers

Anderson and Gehringer, 1965; (NOAA, 1985 and 1990)

59-62

Commercial + sport, complex

morida (Atlantic

Oc-01 284

ATLANTIC NORTH WEST

USA p o r t h Carolina]

285

Pamlico/Roanoke/Core Sounds

Epperly and Ross, 1986; (Giese et al., 1979; Epperly, 1984)

70-85

286

Albemarle/Croatan/Currituk Sounds

Epperly and Ross, 1986; (Giese et al., 1979)

70-85

Barnegat Bay

Hillman and Kennish, 1984; (Chizmadia et al., 1984; Durand, 1984; NOAA, 1985)

57-80

Nixon, 1982; (Briggs, 1965; Bruno et al., 1980; Fox, 1981; NOAA, 1985, 1990) Nixon, 1982; (NOAA, 1985, 1990) Nixon, 1982; (Briggs, 1965)

60

Olsen and Lee, 1982; (Nixon and Lee, 1981; Crawford 1984) Olsen and Lee, 1982; (Mulkana, 1966; Stolgitis et al., 1976; Nixon and Lee, 1981; Crawford, 1984)

79

Commercial + sport

79

Commercial + sport

mew Jersey] 287

Commercial only, without fish Commercial OdY

p e w York] 288

Gardiner/Peconic Bays

289

Great South Bay

290

Moriches/Shinnecock Bays

[Rhode Island] 29 1

Charleston Pond

292

Point Judith Pond

"Not used in the data analysis.

=

Nigrinet

60 60

Commercial + sport, complex Commercial + sport Commercial + sport, complex

110

J.-C. JOYEUX AND A. 8. WARD

sources were available, we retained a datum for the time nearest to the yield record period. The names, abbreviations and units for all the variables described in the following paragraphs are given in Table 2. Yield is the ratio of the annual fishery landings (fish, crab, shrimp, prawn and squid) to the lagoon’s water area, expressed in kilograms per hectare per year. In all but four (Laguna Ceuta, 231, Agua Brava, 235, Alvarado, 260, Mexico and Joyuda, 268, Puerto Rico), bivalve landings were not included, since records generally did not distinguish natural catch from aquaculture yield. For the Cuban lagoons (262-267), annual yields were estimated from a period slightly shorter than a year. We computed a mean yield when landing statistics were available for several years. All recorded yields are later than the 1950s, with the exception of Comacchio (41, Italy) for which landing records began as early as 1781 and ended in 1982, providing a data series of 202 years (Ardizzone et al., 1988). This vallicultured lagoon showed a stable yield. To account for the diversity of the fishery data and the fishery techniques (cf. Section 2), the development of an index was required to be used as a dummy variable in the statistical analyses. The index we adopted depended on the quality of the yield data. The index also accounted for particular practices and devices in use within the lagoon, such as valliculture, fry stocking, brushpark, and fish barrage. Called “under-estimation” (UE) the index’s value increases when the part of the yield which is unaccounted for also increases. Definitions of UE‘s seven categories are given in Table 3. When unknown, but apparently not negligible, yields of crustacean and mollusc (crab, shrimp, prawn and squid) were assumed to be less important than yields of fish. Similarly, recreational fishery yields were assumed to be lower than commercial yields. In some specific cases, these general assumptions were false. Crustacean yields were larger than fish yields in Mirim/ Imarui/Santo Antonio (248, Brazil), in the Laguna Madre de Tamaulipas (261, Mexico), and in all the lagoons of Texas (270-275, USA) except Laguna Madre (269) (Kapetsky, 1984). However, most known cases showed that the yield of crustaceans was lower than the yield of fish (cf. Kapetsky, 1984). When a lagoon was reported to possess two or more different UE values, we kept that showing the larger absolute value. UE is neither a surrogate for the yield nor a descriptive variable sensu stricto, and cannot be used alone as a predictor. We used two closely related geographical data, latitude (Lat) and radiation balance (B). The absolute value of latitude, without distinguishing hemispheres, was recorded to the nearest degree. Radiation balance results from the gain and losses of radiation (incident solar radiation, scattered sky radiation, thermal radiation and reflected short-wave radiation) by the Earth’s surface and the atmosphere at the same location (Robinson, 1966).

Table 2 Names, abbreviations, units and transformations for the variables tested. “ln” is Naperian logarithm, “A” is the value of the coefficient for the BoxXox transformation. The sign of the correlation between original and transformed variables is indicated. Results of the normality test, performed on the largest set of value for the considered variable, are given.

Variable

Abbreviation Unit

Yield and anthropological - - variables Y Yield Tonnes per fisherman TIF UE Under-estimation Number of fishermen NbF Density of fishermen NbFIArea Geographical data Lat Latitude B Radiation balance Morphometrical data Area Water area Mean depth zmean Maximum depth zmax Volume Vol Water exchange Variables Basin area BA Basin arealwater area BAIArea Freshwater run-off R Run-offlarea RIArea Total opening TO Total opening/water area TOIArea Tide Tide Tidal prism TP Flushing index FI Physico-chemical variables Temperatures Tmin T 1 max

kg ha-’ yr-] lo3kg F-’

Transformation

+

NbF/km’

~ n ( x 1) or (h(x))’ or ~n(x)or x ’ / ~ h(x) None A = -0.03716 ln(x)

deg. kcal m-’ yr-’

A = 1.78967 A = 1.23051

km2

A = -0.02820 1/(x + 1) A = -0.17912 A = -0.02230

-

m m km3

km’

W )

-

A = -0.17408

lo6 m3 yr-’ lo6 m3 km-’yr-’

A = -0.24314

km km-’ m 10’ m3

+I))~ 1/ x + 1) b

(In(x XI x1/3

+ 1) + 1)

-

ln(x h(x

“C “C

A = 0.52539 on (x A = 2.29973

+ 2)

Correlation

Normality

+, +, +, +

Yes Yes No Yes Yes

+ NIA + + + + + + + + + + + + + + + + +

No No No No No No Yes Yes Yes No No No Yes Yes No No No

Table 2 (continued)

Variable

Abbreviation Unit

[Nitrates]

N/A: not applicable.

N03-N,, PO4-Pmean P04-P,, SG %SG W %W

1l(x+ 1) 1 = -0.37994 on (x ln(x 1) ln(x 1)

+ +

+ 1)

l l ( x + 1)

ll(x + 1) = -0.31549 on (x l/(x + 1) h = -0.15061 on (x )L

x1/3

+ 1) + 1)

Normality No No Yes Yes Yes Yes Yes Yes

(ln(x)>2

[Nitrites]

Biological variables Seagrass area Relative seagrass area Wetland area Relative wetland area

Correlation

xl~3

Salinities

[Orthophosphate]

Transformation

+ + +

-

No YeS No Yes

CONSTRAINTS ON COASTAL LAGOON FISHERIES

113

Table 3 Parameters deciding the attribution of the value of under-estimation

(W. UE value, meaning

Parameters

-3, largely over-estimated

Maximum annual yield, including bivalve mollusc Yield including bivalve mollusc, generally oyster Yield including intensive aquaculture Maximum annual yield Valliculture, hosha, janos, fish reservoirs, etc., without feeding Yield including small extensive aquaculture Bordigue, pantena, lavoriero, barrage Brushpark, acadja Fry stockin Commerciaf + sport, fish + crustaceans + cephalopods, in USA Commercial; fish + crustaceans + cephalopods, anywhere but USA

-2, over-estimated

- 1, little over-estimated

0, no particular technique, device or practice has been recorded in use in the lagoon and the yield datum seemed to be neither partial nor over-evaluated + 1, little under-estimated

+ 2, under-estimated

+ 3, largely under-estimated

Commercial, fish + crustaceans + cephalopods, in USA Commercial, fish, anywhere but USA Commercial, fish, in USA Commercial, crustaceans, in USA (Pacific North East) Commercial, crustaceans, anywhere but USA (Pacific North East) Incomplete commercial, fish + crustaceans + cephalopods, anywhere but USA Incomplete commercial, fish, anywhere but USA Commercial, crustaceans, in USA (Gulf of Mexico, Atlantic North West) Incomplete commercial, fish, anywhere in USA but Gulf of Mexico and North West Atlantic Incomplete commercial, fish + crustaceans; in USA

B was estimated from a map showing the geographical distribution of the yearly sums of radiation balance on the Earth’s surface, in kilocalories per square centimetre per year (Robinson, 1966). Oceanic curves were used to approximate the values of B at the lagoon’s location. Morphometrical data included water area, generally referred to in this study as Area, and mean and maximum Depths (Z,,,, and Zmm).In Graveyron, Certes, Escalopier (1 3-1 5, France) and Valle Nuova (47,

114

J.-C. JOVEUX AND A. B. WARD

Italy), the area included embankments. Other emergent features, such as sandy bars, emerged marshes, swamps or mangroves, may also have been included (Italian valli, Italy; Godavary, 180, India). Area is the mean water surface except in high tidal range sectors where it refers to the surface flooded at high tide (Tillamook, Netarts and Coos, 225-227, USA). For lagoons with intermittent opening, drying in part or in a whole, Area is the maximum flooded area during the humid season (Kelbia Sekbha, 147, Tunisia; CaTmanero-Huitzache, 232, Mexico). In a few cases (Zalew Szczecinski, 1, Poland; Kleiner Jasmunder Bodden, Dad-Zingster-Bodden West and East, 3-5, Germany; Moriches/Shinnecock, 290, USA) the surface area was estimated from published maps (Hgkanson, 1981). Very shallow parts of the lagoon and zones closed to fishing were included. When known, the existence of zones permanently or periodically closed to fishing was indicated in Table 1. For Tillamook, Netarts and Coos (225-227, USA), records for both mean and maximum depths refer to the period of high tide. Volume (Vol) was derived from data for Area and mean depth (Z,,,,) (Hbkanson, 198 I), following the equation: V O = ~ Area

x

z,,,~,, x 10-~

(1)

with Vol in km3, Area in km2 and Z,,,, in metres. It has to be noted that variables derived from a combination of two variables may be subject to large errors. Environmental data group together variables that characterize exchanges of water between a lagoon and its surrounding environments, and some physico-chemical and biological variables. Data quantifying the actual water inputs into or from the lagoon system, or potentially linked to these exchanges of waters, were treated together. The subgroup includes basin area (BA), freshwater run-off (R),total opening to the sea (TO) and tide. The data for the area of the watershed basin (BA) and for the riverine freshwater inputs (R)were compiled from published sources. Run-off data may include precipitation. Data for the total opening to the sea were computed as the sum of the minimal width of each inlet and therefore correspond to the length of shoreline open to the sea. Second- and third-line lagoons, i.e. lagoons that open in another lagoon or estuary, were considered open to the sea. Thus, connections to lower-order lagoons were considered to be inlets (DarD-zingster-Bodden West, 4, Germany; Is Brebeis and Maestrale, 51 and 52, Italy; Ischkeul, 142, Tunisia; Albemarle, 286, USA). Variations induced by closure or seasonal drought were ignored, and the largest TO value was recorded. Because TO data were commonly derived from map measurements and because large lagoons were generally represented at a small scale, we cannot dismiss the possibility that TO was comparatively over estimated for large estuarine systems, although a sig-

115

CONSTRAINTS ON COASTAL LAGOON FISHERIES

nificant bias is unlikely. The mean height of the local tide was recorded without mention of tidal rhythm (diurnal, semi-diurnal or mixed). Second and third-line lagoons were assigned a null tide. Ratios of BA, R, TO and tide to water area were computed and used as independent variables. Flushing potentialities were estimated through computation of the tidal prism ( T P ) and the flushing index (FI; Bowden, 1967): T P = Tide x Area

(2)

and Tide FI=-zmean

Tide x Area --T P Zmean x x Area VO~

(3)

with Area the water area (km’), Vol the water volume (km3) and Zmeanthe mean depth (m). The physico-chemical variables included temperatures, salinities and concentrations of various inorganic nutrients. Generally, minimum and maximum water temperatures (Tminand Tmax) and salinities ( S ~ and n Smm) were the lowest and highest monthly means for the whole lagoon from one or several years of record. Exceptionally, spatial or temporal records were used. Similar remarks apply for mean and maximum nitrite-nitrogen (NOrNmean and NOz-N,,,), nitrate-nitrogen (N03-Nmean and N03-Nmm) and phosphate-phosphorus (PO4-Pmean and P04-Nm,,) concentrations. Salinities were yearly means for Tam Giang (195, Vietnam) and Kitaural Kasumigaura (197, Japan). Air temperature, instead of water temperature, was used for Calich (56, Italy). The biological variables included area of immersed vegetation (phanerogams and macroalgae = SG, for seagrass) and of wetlands (emerged vegetation: mangroves and marshes = W). Data were primarily recorded from compilations on lagoons by West et al. (1985) for New South Wales, Deegan et al. (1986) for the Gulf of Mexico and NOAA (1990) for the US outside the Gulf of Mexico. Ratios between vegetation area and water area (%SG and %W) were computed as SG %SG=Area

(4)

and

%W=

W Area W

+

where SG, Area and W are expressed in km2.

(5)

116

J.-C. JOVEUX AND A. B. WARD

Although fishing practices are extremely diverse and thus difficult to compare, it was important to get an approximation of the fishing effort. A good index would have included the number of fishermen, the number and nature of gear in use, and the crew manning each gear. Unfortunately, the evolution of techniques during the last 50 years vitiates comparison of data for the same lagoon over the span of a few years (see Nauen, 1984). Thus, the number of fishermen (NbF) was recorded without regard to the gear used. The density of fishermen (NbF/Area)was computed as the ratio between NbF and Area. The yearly yield per fisherman (in tonnes per fisherman per year, i.e. CPUE) was estimated and tested as a dependent variable (T/F). The inclusion of sport, occasional or part-time fishermen can significantly decrease the value of yield per fisherman (see Zaouli, 1984).

3.2. Statistical Treatment

Sets excluding missing data were built. All sets included yield and underestimation. Except for UE, all variables (V) within a set were normalized at a = 0.05 through logarithmic or square root transformation. Normality of the distribution was tested by the Kolmogorov-Smirnov-Liellifors method (Liellifors, 1967). When a significant probability level was impossible to reach by any simple transformation, a Box-Cox transformation was performed on the data (Legendre and Legendre, 1983). The new variable ( V ' ) was computed as

v

I

vA-1

=-

A

where A was the fitted parameter providing the best normalization. V' was kept if it reached the 5% probability level or was less a-normal than the other transformations. These (I priori transformations were performed to achieve normality of non-normal distributions and stabilize the error variance (i.e., achieve homoscedasticity). Transformations, results of normality tests, and signs of correlation between original and transformed variables are given Table 2. Various transformations, relating to different data sets, were used for Yield (Table 2). Within each set, the dependent variable(s) was linearly or curvilinearly tested against every explanatory variable. Multilinear regressions including UE were systematically attempted. On occasion, multilinear models with independent variable(s) belonging to other sets were tested. In this case, the additional variable(s) was not transformed anew. Note that the data set including Y, Area, Lat, B and UE was complete (n = 274).

CONSTRAINTS ON COASTAL LAGOON FISHERIES

117

All reported relationships were significant at the 5% probability level. For curvilinear relationships, each initial and derived variable (i.e. Xi', Xi2,Xi3, . ..,X,")presented a significant (a! = 0.05) correlation with Y. Each explicative variable included in multilinear models was significantly correlated with Y (a! = 0.05). Although independent variables should theoretically be independent of each other at the 5% probability level, some regressions with correlated explicatory variables are given. To verify the fit of the regression and search for outliers, residuals were plotted against the predicted values of Y and each explicatory variable (Draper and Smith, 1981). The studentized residuals (Res), i.e., the residuals (computed as observed value minus predicted value ) divided by their standard deviation, of the regression established on latitudinal location, water area and underestimation (Eq. 19 in Table 5 , Section 4) were used as dependent variables, without prior transformation.

4. GEOGRAPHICAL AND MORPHOMETRICAL CONSTRAINT

In this section we report on our analyses of the influence or the correlations of the geographical and morphometrical variables on the yield of the fisheries. Information about geography and morphometry is commonplace, and this fact may explain why several regional and global studies are available for comparison with our work. More importantly, these parameters are appropriate to providing a foundation for subsequent analyses of environmental and anthopogenic constraints influencing lagoon fisheries. 4.1. Descriptive Statistics

Records for the fishery yield varied between 9 months for the Cuban lagoons (262-267) and 202 years for Comacchio (41, Italy) (Figure 2). Variability in fishery yield was high between lagoons and among years (Figure 3). The interannual range of yield depicted in Figure 3 includes yearly fluctuations and long-term changes in the catches. Hence, singleyear data and long-term averages reflect trends in chance and effort, respectively. The geographic origin of lagoons in our sample varied but was highly skewed toward the Mediterranean/Black Sea zone (Figure 4), which was the only fishing area extensively and adequately covered. The under-representation of most areas is a result of the lack of statistics and/or the unavailability of reports. Our limited language skills and the tendency of reference indexes, data bases and inter-library loans to cover Western-oriented material

118

J.-C. JOYEUX AND A. B. WARD

Number of years of record Figure 2 Number of lagoons with respect to the number of years of record for the yield estimates (n = 274).

contributes to the low availability of information from other areas. Underrepresented areas that have numerous lagoons include the Atlantic Central/ South East and Indian West (essentially Africa; cf. Burgis and Symoens, 1987, and Vanden Bossche and Bernacsek, 1991), the Indian East and the Pacific Central/North West (South East Asia, Indonesia, China, Russia; cf. Scott, 1989) and the Atlantic South West (South America; e.g. Schwarzbold and Schafer, 1984). In other geographical areas, such as the Pacific North/ South East (North and South America) and the Pacific Central East (Micronesia), the geological and morphological features of the coast severely restricts the possibilities for a lagoon of Lankford's (1976) Types I, I1 or I11 to develop. In a few countries, such as South Africa, no commercial fishing activity is allowed in lagoons. Descriptive statistics for the variables used in this study are displayed in Table 4. From the median values, a typical fished lagoon shows a surface area of 21 km2,a mean depth of 1.2m, and is situated near the 37th parallel. Such a lagoon yields 53 kg of fish and crustacean per hectare and per year. The distribution of fishery yields is given Figure 5. The mean yield (101 kgha-lyr-') computed from our set of 274 lagoons is slightly higher than the mean (94kgha-'yr-I) calculated by Chauvet (1988) from 139 Mediterranean and African lagoons. The median yield of 53 kg ha-' yr-' obtained in the present study is close to the 51 kg ha-' yr-' that Kapetsky (1984) calculated from a set of 107 lagoons worldwide.

CONSTRAINTS ON COASTAL LAGOON FISHERIES

119

Not surprisingly, under-estimation (UE) was highly correlated with the yield (Y). The most extreme values of UE (-3 and + 3; Table 3; Figure 6) diverged somewhat from the regression line, indicating that the correction factor was incorrectly estimated for these two levels in comparison with the others. Note that all Y values for UE # 0 were within the range of Y for UE=O. The equation (Eq. (7)) relating these two variables is given for descriptive purposes in Table 5, where the other equations determined from the geographical and morphometrical data are displayed. Also UE was highly correlated with both explanatory variables latitude (Lat) and Area. The percentage of explanation brought by the relationship was small, especially considering the presence of two independent variables (Eq. (8) in Table 5). The relationship indicates a strong tendency to obtaining underestimated values of Y for larger lagoons and for lagoons situated at lower latitudes. 4.2. Geographical Data

Two significant relationships were found between yield and latitude. However, the part of the variation explained by the regressions was, at best, poor: the second- (not shown) and third-order curvilinear (Eq. (9) in Table 5; Figure 7) relationships respectively described 3.1 (p < 0.02) and 9.7% (p < 0.0001) of the Y variability. Lat was not linearly related with Y. The frequency distribution of Lat was, even after transformation, definitely non-Gaussian and may have influenced the results of the analysis (Legendre and Legendre, 1983), although this is doubtful. The computed curve showed two maxima, respectively situated at 0" and 4 H 9 " latitude. The second maximum may be a computational artifact; there was no lagoon between these latitudes in our set. Two minima, at 26-31" and 54" (the computed Y reaches 0 at 61") latitude, illustrated the fact that lagoons in sub-tropics and high latitudes (>49") seem to yield relatively little, although the small sample size for these latitudes prohibits any definitive conclusion. Note the steady improvement in yield with increasing latitude in mid-latitude lagoons, between 300-500 after transformation (Figure 7). Most of the data for these latitudes were from Mediterranean lagoons. The multilinear equation that included powers of Lat plus UE (Eq. (10) in Table 5) was theoretically not valid since Lat and UE were highly correlated (p = 0.0007). However, explanatory power improved from 10 to 15% compared to Eq. (9) (Table 5), and all partial probabilities were significant. No linear or curvilinear relationship between Y and B was demonstrated (Figure 8). B and UE were highly correlated 0,= 0.005), a condition which may reduce the use of the significant multilinear regression that included UE

120

J.-C. JOYEUX AND A.

cu c

1 I I

Z

J

I

m

cu I

I

c.

I

*-I 4

1

I

e

I

._.

.

I

B. WARD

-Schlei Fjord (10) -Encanissada (21) -Canal Veil (21) -Tancada (21) -Lea Olles (21) -Mar Menor (7) -Mauguio (5) -Tortoli (22) -0rbetello (21) -Lesina (34) -Colostrai (10) -Comacchio (202) -Fondi (25) -Valle Nuova (29) -Santa Giusta (5) -Port0 Pino (3 units; 1980) -1schkeul (18) -Garh El Melah (14) -Tunis (18) -El Bibans (21) Khniss (8) -Manzalah (9) -Pangalanes (19) -Pang. (6 units; 1967) -Chilka (9) -Leschenht (24) 0 ster Inlet (12) ilson Inlet (24) -Irwin Inlet (12) -Broke Inlet (12) -Kasumigaura (27) -Kitaura (26) -Kitaura/Kasumi aura (26) -Clarence River 0) -Huitzache (6) -Conceicao (12) -Lag. MadrelB. Bay (6) -Corpus Cristi (6) -Armsas (6) -San Antonio 6) -Matagorda (61 -Galveston (6) -Sabine (6) -Calassieu (15) -Barataria (13) -Tam a (27) -Pamlco (16) -Albemarle (16) -8arnegat (17)

+.,

-- I

I

e -

-

Id

8

I I

I

=

..’....,I

10

100

. ....

1000

Yield (kg hr-1 yr’)

Figure 3 Year-to-year variations in fisheries yield in some lagoons around the world. The number of years of record is given in parentheses after each name. Diamonds are the means and heavy vertical bars are minima and maxima for the period of study. Additionally, light vertical bars show the yield of the lagoon’s subunits for the same year: Porto Pino (three units; 1980) shows the yield for the three subunits (La Spiaggia, Is Brebeis, and Maestrale) during the year 1980;Pang. (six units; 1967) shows the individual fish yields for six subunits of the Pangalanes for the year 1967 (Lasserre, 1979b). Unless otherwise expressed, data are reported from the references listed in Table 1; Kitaura/Kasumigaura: pond smelt, goby and shrimp only (Kitabatake, 1987); Clarence River: Metapemeus macleuyi only (Glaister,

CONSTRAINTS ON COASTAL LAGOON FISHERIES

121

150 125 100

75

50 25

0

Geographic area Figure 4 Number of lagoons from the fishing areas of the world as defined by the Food and Agriculture Organization of the United Nations (n = 274). Abbreviations: A. = Atlantic; MED. = Mediterranean and Black Sea; I. = Indian; P. = Pacific; N., C., S. = North, Central and South; E., W. = East and West. Areas are listed in the same order as in Table 1.

(Eq. 11 in Table 5). While the partial probability for B was low (p = 0.04), the yield significantly increased with increasing amounts of cumulative radiant energy.

4.3. Morphometrical Data

The negative relationship between yield and water surface area was highly significant (p < 0.0001) but explained only 6% of the total yield variability Figure 3 (continued)

1978); Broke, Irwin, Wilson, Oyster, and Leschenault from Lenanton (1984); Chilka: only fish; Huitzache: part of Caimanero-Huitzache (Edwards, 1978); Sabine, Galveston, Matagorda, San Antonio, Aransas, Corpus Christi, and Laguna Madre/Baffin Bay: commercial only; Barataria: only commercial shrimp.

Table 4 Yield, under-estimation, geographical and morphometrical characterization of the 274 lagoons used in the data analysis. Data are untransformed. Units are in Table 2.

Variable Yield and under-estimation Y UE Geographical characteristics Lat B Morphometrical characteristics Area zme8n

z-

Vol

Mean

Standard deviation

Coefficient of variation

Median

Range

n

101 0.055

188 0.898

181 1.640

53.0 0

4 to 2242 -3 to 3

274 274

33.6 79.3

11.0 19.6

32.6 24.7

37.0 80.0

5 to 54 30 to 120

274 274

284 1.97 6.08 2.98

1156 2.42 6.64 26.7

407 123 109 898

21.0 1.20 3.10 0.046

0.1 to 14344 0.2 to 25 0.6 to 35 O.ooOo6 to 359

274 184 163 184

123

CONSTRAINTS ON COASTAL LAGOON FISHERIES

IP

0

60: 40-

B

8 8 8 0 0 %

- d l a # s ~ s s G a s 8M U Yield (kg ha-1 y r ' ) Figure 5 Distribution of fishery yields (in kg ha-' yr-') in 274 lagoons worldwide. Each interval of yield covers 25 kg ha-' yr-'. A double slash on the abscissa indicates no record for intermediate values.

-3

-2

-1

0

1

2

3

Under-estimation Figure 6 Yield (transformed data) vs. under-estimation for 274 lagoons worldwide. The regression line (Eq. (7) from Table 5) is overlaid.

Table 5 Selected sigruficant relationships. “Eq.”is the equation number referred to in the text. “Dpdt” is the dependent variable, and “Cste” the fitted constant. Variables 1-5 are the explicatory variables, with their partial probabilities (Pp), regresion coefficients (Coeff.), standard errors of the coefficients (S.E.).Abbreviations as in Table 2. Number of values (n),R-squared (I?) and probability for the models (p) are given. Variables were transformed (see Table 2). Transformation for the value of yield Y+ : In( Y 1).

+

-7

Y+

UE

8

9

Y+

3.96

0.19

5.19

10

Y+

5.19

I1

Y+

5.06

E

Y

4

E

r

UE

-0.36 (0.07)

Lar (0.008)

-9.60 x 104 (3.75 x 104)

Area (0.02)

0.06 (0.02)

Lot (0.0001) Lot (0.oOOl) B (0.34)

-1.76 x (0.33 x 10-2j -1.40 x 10(0.30 x 10-2) 2.66 lo-’ (1.23 x lo-))

La?

5.48 x 10-5 (1.13 x lo-’) 4.20 x I O - ~ (1.14 -0.394 (0.076)

e

-0.127 (0.030) -0.103 (0.033) 1.26 (0.42) -0.374 (0.111) 1.34 (0.42) -8.80 x lo-’ 2.22 x 10-2 -7.53 x 10-2 (2.25 x 1.30 x 10-2 (0.30 x 10-3

a

I3

Y+

4.25

14

Y+

3.52

&

15

Y+

4.45

&

16

Y+

4.54

&-, (0.02) Yo1

17

Y+

3.76

I8

Y+

3.81

19

Y+

5.61

Area (0.W

Yo1 (0.001) Lar

(0.0001)

(0.000l)

La? (0.0003)

UE (0.000l)

UE (0.0001)

UE (0.007)

UE (0.02)

I d (0.009

La? (0.000l)

Lo? (0.001)

-4.66 x 10-8 (1.05 x I04J -3.46xlO UE (1.06 x lo4) (0.000l)

-0.321 (0.077)

-0.310 (0.075)

-0.231 (0.085) -0.205 (0.084) 3.19 I O - ~ (1.13

La? (0.03)

-2.33 x lo4 (1.06 x lo-’)

Area (0.0001)

-0.132 (0.031)

UE (0.0001)

-0.291 (0.075)

274

0.08

0.0001

274

0.06

0.0001

274

0.10

0.0001

274

0.15

0.0001

274

0.09

0.0001

274

0.06

0.0001

274

0.12

0.0001

184

0.05

0.0031

163

0.07

O.ooo9

184

0.08

0.0003

184

0.11

0.0001

184

0.11

0.0001

274

0.21

0.oOol

125

CONSTRAINTS ON COASTAL LAGOON FISHERIES

0

100 200 300 400 500 600 700 800 Latitude (lambda

I

1.78967)

Figure 7 Fishery yield vs. latitude (transformed data) for 274 lagoons worldwide. The third-order curvilinear relationship (Eq. (9) from Table 5) is overlaid. Both variables were transformed.

0a

L

0

7.

6-

0

.

8

0

@

0

+ f

C

4 0

2. 1. 0. 0

0

50

100

150

Balance (lambda

200

250

*

300

= 1.23051)

Figure 8 Scatterplot of fishery yield vs. radiation balance (both transformed) for 274 lagoons worldwide.

126

J.C. JOYEUX AND A. B. WARD

(Eq. (12) in Table 5; Figure 9). Explanatory power increased to 12% with the inclusion of UE (Eq. (13) in Table 5). However, Area and UE were highly correlated (p < 0.0001). Both relations (Eqs (12) and (13)) indicated decreasing fishery yields with increasing water area. This was consistent with the negative linear relationships (Figures 10 and 11) existing between yield and depths; the corresponding equations are given in Table 5 (Eqs (14) and (15)). While both mean and maximum depths were significantly correlated with UE (p=O.O11, 0.0098, respectively), only the mean depth model for yield was improved by the addition of UE (Eq. (16) in Table 5). As a combination of two variables, volume data integrated errors from both origins (Area and Z,,,,,,) and should be considered with circumspection. The relationship found between fishery yield and water volume had low explicatory power (Eq. (17) in Table 5; Figure 12). The multilinear relationship including UE is described by Eq. (18) (Table 5 ) , even though Vol and UE were significantly correlated (p = 0.0015). Following the previous findings that the yield was significantly lower in large lagoons (Eq. (12)) and in deep lagoons (Eqs (14) and (15)), the fishery yield decreased when the water volume increased. In fact, water area was strongly correlated with both depths (R=0.60 and 0.54 for Z,,,,,, and Z,,,,,, respectively, all variables transformed); the correlation between area and volume was lower

c

1

-4

-2

0

2

6

4

Water Area (lambda

8

10

-0.02820)

I

Figure 9 Change in fishery yield according to water area (both variables transformed) for 274 lagoons. The relationship described by Eq. (12) (Table 5) is overlaid.

127

CONSTRAINTS ON COASTAL LAGOON FISHERIES

1-

-

0

0

.1

.3

.2

.4

.5

l/(Mean Depth

.6

.7

.8

.9

+ 1)

Figure 10 Change in fishery yield according to mean depth (both variables transformed) for 184 lagoons worldwide. The regression line from Eq. (14) (Table 5) is overlaid.

-1

1.5

0

.5

1

1.5

-

2

2.5

3

Maxlmum Depth (lambda = 0.17912) Figure 11 Correlation between fishery yield and maximum depth in a set of 163 lagoons worldwide. Equation (15) (Table 5) is overlaid. Both variables were transformed.

128

J.-C. JOYEUX AND A. B. WARD

0 ’

0- - -12 -10 - 8

-6

- 4

0

- 2

Volume (lambda

2

i

4

6

- 0.02230)

I

Figure 12 Fishery yield vs. water volume (both transformed) for 184 lagoons worldwide. Equation (17) (Table 5) is overlaid.

(R =0.23). Attempted multilinear regressions involving either the independent variables Area and Z,,, or Area and Z,,, did not provide any valid model where both explanatory variables were significant. In both cases, only Area remained correlated with Y. Vol remained correlated with Y when Area was added as an explanatory variable. These observations indicate that the volume, more than the combination of surface and mean depth, gives a better description of the yield of the fishery by itself than either area, depth, or area plus depth. 4.4. Multivariate Analysis

Several significant multilinear relationships with Y as the dependent variable were demonstrated. B could not be included in combination with Lat. Combinations were as follows: Lat + Area + UE ( p < 0.0001, R2=0.15), B + Area + UE ( p < 0.0001, R2=0.14). The best regression (Eq. (19) in Table 5) combined linear relations on UE and Area (cf. Eqs (7) and (12), respectively) and a curvilinear relation on Lat (cf. Eq. (9). While all other variables (Lat, Lat2, Lat’, Area and UE) stayed partially correlated with Y,B did not reach significance when forced in Eq. (19).

CONSTRAINTS ON COASTAL LAGOON FISHERIES

129

The studentized residuals (Res) from Eq. (19) were not linearly or curvilinearly correlated with either UE, Lat, Area, Z,,,,, Z,,, or Vol. To begin with, Y was poorly related to Z,,, Z,,,, Vol (Eqs (14H18)). Once the effect of Area on Y is accounted for, no other morphometrical variable brings additional information. 4.5. Discussion

Since surveys are more difficult to carry out on large bodies of water, landing statistics are often incomplete (Eq. (8)) for studies often focus on target species of economical significance. By contrast, practices and devices that are more common in smaller entities, such as valliculture, fry stocking, fish barrage and to a lesser degree brushpark, improve the annual yield of the fishery (Eqs (7) and (8); Fig. 6: category U E = -1) (Welcomme, 1972; Ardizzone et al., 1988; Chauvet, 1988). The negative link between UE and Lat (Eq. (8)) primarily results, in our opinion, from a sample where wellstudied, small-sized, mid-latitude lagoons (Table 4) were numerically dominant. However, reduced man-power, economic possibilities, and weaker reporting regulations in countries in development may have lead to fewer, more partial studies. A low availability of reports may also be implicated. Using a similar statistical approach, Brylinsky and Mann (1973) determined that the most productive lakes and reservoirs were situated between the latitudes 10" and 12"N. For lagoons, the only attempt, which was unsuccessful, to link fishery yields and latitude was carried out by Kapetsky (1983). The weakness of the correlation between these two variables makes Kapetsky's (1983) negative result not surprising owing to the relatively small set of data. Most of the 104 lagoons he used are included in our set. The various models including latitude and radiation balance (Eqs (9)(1 1)) had extremely low explicatory powers (maximum= 15% with Eq. 10). This indicates that these variables were not major determinants of the observed variability in yield (Figures 7, 8). Either latitudinal changes in fauna (number of species, migratory species, pelagic species, etc.), or ecology (predation, competition, etc.), or phenomena linked to seasonality (energetic inputs onto extended vs. contracted periods) may compensate for the diminution of yearly inputs of radiant energy at increasing latitudes. The same hypotheses may explain the third-order curvilinear fit relating yield and latitude. In open littoral environments, such as bays, embayments and continental shelves, several authors have successfully related shrimp fishery yields to latitude (Turner, 1977, 1986; Yaiiez-Arancibia et al., 1985; SoberbnChavez et al., 1986; Pauly and Ingles, 1986). These areas possess open boundaries and are larger and physico-chemically more stable than

130

J.-C. JOYEUX AND A. B. WARD

lagoons. The models developed are valid within the latitudinal range 0"-36", latitudinal range of distribution of most penaeid shrimps, and include a third variable, the intertidal vegetation area. The equations given by Pauly and Ingles (1986), chosen because of their worldwide scope are MSY = 2.158 - 0.01539 x Lat log&zJ (n = 38, R2 = 0 . 1 0 , ~< 0.05)

(20)

and, log(MSY) = 2.41 - 0.0212 x Lat + 0.4875 x loglo(Z.Area)

(21)

(n = 38, R2 = 0 . 5 3 , ~< 0.01) where MSY is the maximum sustainable yield, i.e., the average landing in years with high, stabilized effort (Turner, 1977), Z.Area is the area of intertidal vegetation (salt marsh macrophytes and mangrove) and Lat the absolute value of latitude. MSY is expressed in thousands of tonnes (lo3 t), Area in 103km2 and MSY/Z.Area in kgha-'. These authors added that their equations cannot be used as predictive models because of the influence of the logarithmic operator over a sum of intertidal vegetation areas for the same MSY. Using updated data from a previous work (Turner, 1977), Turner (1986) estimated that loglo(MSY/Z.Area)reaches its maximum for latitudes close to 5"N/S.This result is not contradictory with Pauly and Ingles' (1986) linear relationship (Eqs (20) and (21)). A better equation can be fitted to their own data; this increases the coefficient of correlation to 0.55 and produces a local maximum near 13.5" latitude (Eq. (22)): MSY = 1.73 + 0.0569 x Lat - 0.00209 x Lat2 log1o(Ki) (n = 38, R = 0.55, R2 = 0 . 3 0 , ~= 0.002) (partialp = 0.02 and 0.003, respectively for Lat and Lat2)

The local maxima observed between 0-5" and 13" latitude for the yield of shrimp fisheries are consistent with our finding of the intertropical maximum at latitude 0".However, the limited extent of the geographic distributions of these shrimp at higher latitudes necessarily leads to at least one minimum yield and, therefore, one maximum.

CONSTRAINTS ON COASTAL LAGOON FISHERIES

131

The fishery yield was inversely correlated with the lagoon area (Eqs (12), (1 3) and (1 9)). This observation was already reported by Kapetsky (1 983), whose relation was

Y = 112.39 x (n = 93, R2 = 0.27)

where Y is in kg ha-’ yr-I and Area is the lagoon surface area in km2. On a smaller scale, Amanieu and Lasserre (1981) showed that fishery yield and water area were negatively correlated in Mediterranean lagoons of less than 3000 ha in one part. Their equation was

where Y is in kg ha-’ yr-’ and Area is in hectares. In an overview of fishery yields in various environments (lagoons, marine bays, oceans, temperate and tropical lakes), Nixon (1988) negatively correlated the fished area of marine systems to their yield. No equation was furnished for either marine or freshwater ecosystems. Surface areas of the ecosystems studied were between lo2 and 10”ha. It is unfortunate that data for 19 lagoons given by Goode (1887) and used by Nixon (1988) were included among the 50 data for the marine set. These data furnished the largest component of small-sized saltwater bodies: 61% of the systems smaller than lo6 ha and 77% of those smaller than lo5 ha. We think their inclusion and comparison with more recent data introduce a bias because these old records are difficult to compare with more recent ones (Olsen and Lee, 1982). Nixon (1988) justified it by arguing that “few small marine areas are fished intensively any longer”, and this might be true for New England estuaries. Both this author’s study and the present one are very similar in regard to the extreme scattering of the data points on the yield axis. It is therefore doubtful that any relationship would have been found without the inclusion the data of Goode (1887; see Nixon, 1988). The link between fishery production and lagoon area may not be as direct as suggested by relationships described by Eqs (12), (13) and (19) (Table 5). In fact, the surface area is one of several highly correlated morphometrical features which characterize the lagoon. The individual effects of each of these parameters are difficult to distinguish. More importantly, the way the fishery is managed depends on the size of the lagoon. Fishermen in small lagoons generally rely on small boats and passive gear (e.g., Chauvet, 1988). Large vessels such as trawlers and seiners are heavily used in large and deep lagoons (e.g., Ewald, 1964; North Carolina

132

J.-C. JOYEUX AND A. B. WARD

Division of Marine Fisheries, 1996). In the latter the cost in equipment and maintenance to run large boats and pay crews force fishermen to concentrate their effort either on highly priced catch such as shrimp and prawn or on low price abundant “pelagic” fish. Neglect of other fishery resources, or their destruction as discards, may be a major reason for the relatively low yield observed in large lagoons. Pauly and Ingles (1986) (Eq. (25) below, see also Eqs (20), (21) and (22)) showed that offshore yields of shrimp fisheries worldwide are strongly correlated with the area of intertidal vegetation: MSY

= 2.761 - 0.24194 x loglo(Z.Area)

(25)

(n = 38, R2 = 0 . 2 6 , ~< 0.01) with the same units than for Eq. (20). Although untested worldwide, one can expect offshore fish yields to be also dependent upon the littoral area covered by wetlands, at least for species or group of species that use these habitats as nurseries. Such a relationship was barely demonstrated for the Mexican part of the Gulf of Mexico by Yhiiez-Arancibia et al. (1985), whose equation was lo&(Cupture) = 6.070 + 0.496 x lo&(Z.Area) (TI =

(26)

11, R2 = 0 . 6 9 1 , ~= 0.05)

where Capture is the average commercial fishing catch in lo3kg (including fish, shrimp and other macroinvertebrates) and Z.Area the total area of coastal vegetation (marshes) in km2. In some inland freshwater bodies, yields per unit area have been shown to be negatively correlated with water area (De Silva et al., 1992 (Groups I and 111)). However, most studies reported no relation between these two variables (Moreau and De Silva, 1991; Ranta and Lindstrom, 1989; Hanson and Leggett, 1982; De Silva et al., 1992 (Group 11)). Fish yield decreased as the lagoon increased in depth. The greater explanatory power of maximum depth (7%) compared to mean depth ( 5 % ) was unexpected, because we knew recorded maximum depth sometimes corresponded to a natural or artificial bathymetric features. Artificial morphometrical features include mainly dredged areas. Natural features include extremely localized areas such as in Thau (29, France) where a deep pit in the lagoon’s bottom is occupied by an underwater spring (Jouffre and Amanieu, 1991). Similar situations exist in Paola (37, Italy; Karvounaris, 1963) and Lake Macquarie (204, Australia; Baas Becking et ul., 1959). On the other hand, maximum depth data are obtained from direct measure-

CONSTRAINTS ON COASTAL LAGOON FISHERIES

133

ments, while mean depths are computed from volume to surface ratio or isobath mapping, either possible sources of error. Our results fit with those from Marten and Polovina (1982) who showed that the maximum sustainable yield of demersal multispecies fisheries in intertropical coastal areas was inversely related to mean depth ( R = 0.86, R2 = 0.74, n = 16, no equation given). The depth range was, however, much larger than in our set of lagoons ((r300 m, instead of 0.2-25 m in the present study). In freshwater bodies, such as lakes and reservoirs, authors have successfully related fish yield with the morphoedaphic index (ME4 see Ryder et al., 1974 for review). This index includes the ratio of total-dissolved-solidsconcentration to mean depth. The relationship linking MEZ to fish yields is Y = 0.966 x MEZ = 0.966 x

where Y is the fishery yield in kg ha-’ yr-’, Z,,,, the mean depth in metres and TDS (total-dissolved-solids) in mgl-’. Critics have argued that this ratio makes it impossible to determine the individual effects of each parameter and incorporates errors of measurements for both variables (Ranta and Lindstrom, 1989, 1990). It certainly prohibits any direct comparison with our work. However, Matuszek (1978) considered depth to be a better indicator of fish production than TDS. In studies specifically looking at the relationship between depth (when indicated it is always mean depth) and fish yield in freshwater bodies, the findings were either negative (i.e., no significant relationship: Hanson and Leggett, 1982; Ryder, 1965; Ranta and Lindstrom, 1989; De Silva et al., 1992; Moreau and De Silva, 1991, Philippines) or the yield was negatively correlated with depth (Rawson, 1952; Moreau and De Silva, 1991, Thailand and Sri Lanka; Fryer and Iles, 1972, in Jones, 1982; Jenkins, 1982). Considering the number of independent variables included, the percentage of the variability in yield (21%) explained by the model that integrated powers of Lat, Area and UE (Eq. (19) in Table 5) was extremely low. The actual use of such a relationship for management is, in our opinion, small or null, especially considering that the correlation with Lat might be an artifact. This model, however, confirmed that the characteristics included were similarly affecting the yield when tested alone (independently of each other: Eqs (7), (9) and (12)) or in combination. Studentized residuals (n = 274) were used as a dependent variable to see if unexplained variability in yield was related to morphometric characteristics besides Area. Because the residuals were uncorrelated with Lat, Area and UE, masking effects from these variables were removed (i.e., the variables were “whitened”; Fogarty, 1989). Residuals were tested against Z,,,,, Z,,, and Vol, characteristics which

134

J.-C. JOYEUX AND A. 6. WARD

were themselves highly correlated with Area. No significant correlation was evident demonstrating that these characteristics were redundant once Area was known.

4.6. Conclusions

Several parameters were shown to be influential on the fishery yields from coastal lagoons around the world. Individually, all morphometric characteristics, which included water area, mean and maximum depth, and volume, were inversely correlated with the annual volume of landings per unit area. However, information was redundant because all these variables were intercorrelated. Surprisingly, annual fishery production was largely independent of the amount of solar energy received at sea level and, therefore, relatively unrelated to the lagoon’s latitudinal localization. This counter-intuitive result was not invalidated by our multivariate model. As yet, 79% of the variability of yield remains to be explained. Besides the yield itself, we now have the use of a complementary tool, the residuals from Eq. (19), to evaluate the influence of environmental and anthopogenic factors upon fishery yield in coastal lagoons.

5. ENVIRONMENTAL AND ANTHROPOGENIC CONSTRAINTS

This section will explore the relations between the fishery yield in coastal lagoons and environmental and anthopogenic variables. Specifically, we will attempt to demonstrate that yields are indeed dependent upon the nature and extent of communication between continent, lagoon and ocean.

5.1. Descriptive Statistics

The descriptive statistics for the variables we tested are given in Table 6. Based on medians, the standard lagoon in our sample is used by 110 fishermen and yields 53 kg km-’yr-’ of fish and crustacean. Although a fisherman lands, on average, two tonnes a year, catches may vary between 18kg and 19t. The lagoon connects to the sea through a 390-m-wide inlet. The height of the oceanic tide averages 40 cm. Annual fluctuations in temperature and salinity span, respectively, 20°C and 30 salinity units. Seagrass covers 12% of the lagoon area, and wetlands 10% of the total inundated surface.

135

CONSTRAINTS ON COASTAL LAGOON FISHERIES

Table 6 Yield, fishing effort and environmental characterization of coastal lagoons. Mean, standard deviation, coefficient of variation, median, range of value and number of observations are given. Data are untransformed. Uuits are in Table 2. Variable

Mean

Standard deviation

Coefficient Median of variation

Yield and anthropogenic characteristics Y 101 188 181 3.33 3.96 119 TIF UE 0.055 0.898 1640 NbF 2616 5650 216 NbFIArea 13.6 25.7 189 Water exchange characteristics BA 39.1 97.1 248 BAIArea 63.0 119 190 R 16488 63519 385 RIArea 177 1063 150 TO 1.70 3.47 204 TOIArea 0.0149 0.0274 184 Tide 0.559 0.412 73.7 TP 614 2325 378 FI 0.352 0.472 134 Physico-chemical characteristics Tmin 11.9 8.43 70.7 Tmax 29.6 3.96 13.4 smin 10.9 12.5 114 smax 36.5 25.7 70.4 NOz-Nmean 1.65 2.69 163 N02-Nma, 9.43 23.6 25 1 NO,-Nmea, 13.2 21.1 160 NO,-Nm,, 55.4 151 272 P04-Pmean 4.76 21.7 456 P04-Pmx 16.8 74.9 445 Biological characteristics SG 40.9 125 306 %SG 0.136 0.125 91.6 W 170 328 193 %W 0.200 0.230 115

Range

n

53.0 2.09 0 110 4.00

4 to 2242 0.018 to 18.7 -3 to 3 1 to 27977 0.015 to 114

274 67 274 67 67

4.35 12.4 1990 4.93 0.385 0.0480 0.420 125 0.190

5.95 to 540000 0.999 to 537 0 to 435000 0 to 7531 0 to 20.3 0 to 0.167 0.08 to 1.70 1.180 to 15800 0.030 to 2.60

55 55 50 50 88 88 47 47 46

10.0 30.0 5.00 34.3 0.810 3.90 4.95 15.0 0.745 1.97

-1.4 to 29 14 to 38 0 to 47.1 0.3 to 260 0.017 to 12.2 0.064 to 130 0.076 to 129 0.126 to 1100 0.050 to 150 0.120 to 600

136 135 153 156 31 37 50 57 48 66

1.48 0.123 4.79 0.099

0 to 0 to 0 to 0 to

41 41 49 49

773 0.634 1.300 0.855

5.2. Water Exchange Data

Equations determined from our data and referred to in the text of this section are given in Table 7. No significant relationship was found between basin area ( B A ) , BAIArea, run-off ( R ) , RIArea, total opening (TO), TO/ Area and fishery yield (Y). The residuals from the attempted regression on RIArea showed a strong heteroscedasticity, i.e., the variance was correlated with the mean, and this variance increased as RIArea increased.

Table 7 Selected significant relationships. “Eq.” is the equation number referred to in the text. “Dpdt” is the dependent variable. Variables 1-4 are the explicatory variables, with their partial probabilities (Pp), regresion coefficients (Coeff.), standard errors for the coefficients(S.E.). Abbreviations for the variables are in Table 2. Number of values (n),R-squared (I?) and probability for the models (p) are given. Variables were transformed (see Table 2). Transformations for the value of yield Y : Y o : ln(Y); Y+ : ln(Y 1); Y++: (ln(Y))’; Y x x: cubic root. i = NbF/Area (not transformed).

+

Variable 1 Coeff.

Eq. Dpdt Cste

Variable 2 Coeff.

Variable 3 Coeff. Variable 4

Coeff.

U

~~

Water exchange variables 28 Y o 4.91 Tide 29

Yo

4.60

TP

30

Yo

-3.07

31

Res

-0.197

Tide (0.0220) R

32

Res

-3.32

Ti& (0.0265) 33 Area -8.49 Ti& (0.0006) PW‘d variables 34 Y+ 4.35 T& (0.0016) 35 Y+ 2.78 T& (0.0007) 36 Y” 4.33 s, (0.0147) 37 Y o 4.57 (0.0287) 38 Y o 2.24 kOz-im(0.0091) 39 Y o 2.71 NOz-N,, (0.0343)

-1.56 (0.72) -0.206 (0.063) 15.1 (6.3) 0.00730 (0.00350) 10.4 (4.52) 35.3 (9.54)

T& (0.0085)

-10.6 (3.8)

TO (0.0036)

Tide’ (0.0124) Tidez

-7.23 (2.77) -22.2

(O.Oo04)

(5.85)

UE

-0.239 (0.094) -1.02 (0.28) -0.272 (0.096) -0.287 (0.093) -8.21 (2.83) -6.55 (2.72)

0.121 (0.038) 2.29

(0.01 17) T&

(0.66)

(O.Ooo4)

-0.222 (0.089) -0.0514 (0.0233) 8.92 ’ (3.18) 6.87 (3.08)

UE (0.0053) UE (0.0025) kO2-i:-

(0.0072) NOz-N:(0.023)

.

T L (0.0005)

2.12 (0.67)

TOIArea (0.0392)

0.157 ? (0.044) (0.0008)

3.11 (1.45)

-7.78 x (2.27 x lo-’)

47

0.09

0.0358

47

0.19

0.0021

37

0.42

0.0014

50

0.08

0.0448

47

0.21

0.0063

47

0.25

0.0017

136 0.11

0.0004

136 0.17

0.0001

153 0.08

0.0017

156 0.08

0.0017

31

0.23

0.0252

31

0.36

0.0063

I

UE (0.0263)

-0.407 (0.173)

40

Res

-1.09

Biological variables 41 Y x x 2.58 42

Y x x 6.47

43

Y++

NOz-N,,, (0.0109)

6.22 (2.28)

SG

0.348 (0.128) -3.91 (1.71) 1.83 (0.67) 0.618 (0.136) 0.215 (0.079)

%SG

13.4

W

44 Res

-1.10

(0.0106) SG

45

-0.804

W

Res

N02-Nk(0.0061)

%W (0.0077)

-6.02 (2.03)

-12.6 (4.52)

31

0.25

0.0185

41

0.16

0.0099

40

0.12

0.0281

49

0.16

0.0197

41

0.35

0.0001

49

0.14

0.0089

67

0.12

0.0037

67

0.41

0.0001

67

0.45

0.0001

67

0.41

0.0001

67

0.11

0.0064

67

0.16

0.0009

67

0.07

0.0321

67

0.55

0.0001

67

0.58

0.0001

67

0.41

0.0001

Anthropogenic variables

46

Y+

3.31

NbF

47

Yi

3.61

NbFIArea

48

Y+

3.69

NbFIArea (0.0001)

49

Yi

3.65

NbF

(0.0001) 50

Lat

51

NbF/ 2.20 Area Res -0.732

52

Res

-0.326

NbFIArea

53

TIF

1.24

NbFIArea

NbF

0.177 (0.059) 0.414 (0.062) 0.528 (0.080) 0.500 (0.076) -0.0032 (0.0011) 0.173 (0.097) 0.308 (0.058) -0.566

(NbF/Area)’-O.0495 (0.0364) (0.0231) Area -0.499 (0.0001) (0.090)

(0.064) 54

TIF

1.35

59

Yi

2.67

NbFIArea -0.433 0.0001 (0.082) i 0.5 0.887 (0.0001) (0.154)

(NbFIArea)’ -0.0579 (0.0175) (0.0237) i -0.0695 (0.0001) (0.0152)

A

Y

138

J.-C. JOYEUX AND A. B. WARD

The mean tidal amplitude was marginally correlated with the yield of lagoon fisheries, and this yield decreased with increasing tide (Eq. (28) in Table 7). Yield and tidal prism (TP) were negatively correlated (Figure 13; Eq. (29) in Table 7), which was not surprising since both constituent variables for TP also showed this characteristic. The forcing of UE in the equation furnished a barely insignificant relationship on UE (partial p = 0.0547). The model’s significance was increased by removing the lagoons presenting a UE value # 0 (n = 25, p = 0.0003, R2 = 0.44). The flushing index (H)was not linearly or curvilinearly correlated to Y, and the multivariate regressions dissociating the original variables (Tide plus Z,,,,, and Tide plus Z,,,, plus UE; transformation for Zmean: l / ( x + 1)) were not significant. Because the effects of oceanic tides are exercised through the inlets, we tested a combined data set for Tide and TO for clues on how these variables together related to the yield. Although the number of data were small, a significant model was built (Eq. (30) in Table 7) in which Y was fitted by a second-order curve on Tide. Y was negatively correlated to TO and proportional to TOIArea. We tested all variables involving water exchanges against Res, the residuals from the Eq. (19) (Table 5 in Section 4) based on UE, La? and Area. No linear or curvilinear relationships were detected for BA, BAIArea, TO,

I

2. 1 . -

0

-

1

-

- - . . . - . - . - .- . -

2

3

4

5

6

7

8

.

9

-

r

10

In(TP+l) Figure 13 Change in yield per unit area (Y) according to tidal prism (TP).The set included 47 lagoons. Equation (29) (from Table 7) referred to in the text is overlaid.

CONSTRAINTS ON COASTAL LAGOON FISHERIES

139

TOJArea,TP, and FI. RIArea did not show any significant relationship with Res, but there was a strong heteroscedasticity among residuals for the regressions attempted. A marginally significant linear correlation was demonstrated between Res and R (Figure 14; Eq. (31) in Table 7), and the value of Res increased with the amount of freshwater entering the system. This indicated that in comparable lagoons (i.e. “corrected” for Lat, Area and UE), Y had a slight tendency to increase with more freshwater input. Res was also linearly correlated with Tide (n = 47, p = 0.0498, R2 = 0.083, but again the fit was poor. Lower residuals, i.e., yields, were characteristic of larger tides (cf. Eq. (28)). However a better relationship was obtained by fitting a second-order curvilinear regression onto the data (Figure 15; Eq. (32) in Table 7; cf. Eq. (30)). In this case, lowest Res corresponded to the extremes in the range of Tide. The theoretical maximum Y was situated at Tide= 0.36 m. Interestingly, Tide and Area were curvilinearly correlated (Eq. (33) in Table 7), and the maximum Area was reached at tidal amplitudes close to 0.5m. Small lagoons are encountered in zones where Tide was greater than 1 m or lower than 0.16m. It is probably not coincidental that the optimum value for Area was close to the estimated tide for maximum yield.

(In( R+l))* Figure 14 Influence of freshwater run-off ( R ) on the fishery yield, as indirectly seen through the dependent variable Res. Equation (31) (Table 7) is superimposed. N = 50.

140

J.4. JOYEUX AND A. B. WARD

t

21

0 0

!r

-1.5.

0 0

Tide Figure 15 Effect of tide height on the fishery yield in 47 lagoons from all over the world. The variable Res was used as a proxy for the yield. Equation (32) (Table 7) is overlaid.

5.3. Physico-chemicalData

Yield presented a significant, but weak, correlation with the annual minimal temperature (Figure 16 and Eq. (34) in Table 7). Lat, Tminand T,,, were highly intercorrelated (transformed variables; all 2 x 2 combinations at p=O.OOOl). Tmin was better correlated with Lat (R2 = 0.89) than T,,, (R2 = 0.37). Yield at latitude data were best described by a third-order relationship (cf. Eq. (9) in Table 5, Section 4), a fact which may explain why Y vs. Tmindata could be fitted with a fourth-order curvilinear regression (Eq. (35) in Table 7). The fit was considered an artifact. Yield was not correlated to maximum temperature in any way, and was loosely correlated with minimum (Figure 17 and Eq. (36) in Table 7) and maximum salinities (Eq. (37) in Table 7). No significant relationship was found between Y and N02-Nm,, NO3N,,,,, N03-Nmax, P04-P,,,, and PO4-P,,,. In all cases, Y was significantly correlated with UE (p = 0.0473, 0.0013, 0.0119, 0.0012, 0.0016, respectively). The catch per unit area was curvilinearly correlated with the mean nitrite concentration (Figure 18; Eqs (38) and (39) in Table 7). Note that only four lagoons (13% of the total) allowed the data to be accurately fitted by a curvilinear model; in their absence the relation would have been

141

CONSTRAINTS ON COASTAL LAGOON FISHERIES

.

. 0.

'

8

r !

. Q L - 2

2

0

4

6

T m (lambda

= 0.52539)

t

8

10

Figure 16 Fishery yield vs. minimum temperature (Tmin) in 136 lagoons worldwide. The regression line for Eq. (34) (Table 7) could not be shown. The fourth-order curvilinear relationship mentioned in the text (Eq. (35)) is detectable.

. - . - .

-.

-.5

0

.5

1

1:5

2

2.5

3

3.5

4

Figure 17 Scatterplot of yield vs. minimum salinity (&,id from a set of 153 lagoons. Together, Sminand UE explained 8% of the variability in yield (cf. Eq. (36) in Table 7).

142

J.-C. JOYEUX AND A. 6. WARD

both linear and positive. Also, the data would be better fitted with an asymmetric model (Figure 18). Res was not correlated with either of the following variables : T ~ nT,,,, , NOz-Nmax, N03-Nmean, NO3-NmaX,P04-Pmean, PO4-P,,. Negative linear correlations were demonstrated between Res and S ~ , (, p = 0.0196) and S,, (p=O.O490). In each case, RZ was extremely low (24%). Regression coefficients were of the same sign as those evident during analyses for Y (Eqs (36) and (37)). A curvilinear equation described the relationship between Res and NOz-Nmean (Eq. (38)). Similar to the results for Y, the values of Res were maximum in the middle of the range of NOz-Nmean,and minimum at the extremes. The relation (Eq. (40) in Table 7) explained 25% of the observed variation of Res. 5.4. Biological Data

Fishery yield and area covered by immersed vegetation were linearly correlated (Eq. (41) in Table 7). The data set was composed of 26 small lagoons from New South Wales (all with UE = 0), two large lagoons from the same Australian state and 13 large lagoons from the Gulf of Mexico. The yield

0

1 0

.1

.2

.3

.4

.5

.6

.7

.8

.9

1

t

1.1

Figure 18 Change in yield according to the mean nitritenitrogen concentration in 31 lagoons worldwide. The curvilinear relationship (Eq. (38) from Table 7) is overlaid on the data.

143

CONSTRAINTS ON COASTAL LAGOON FISHERIES

was not correlated with the seagrass area in North American lagoons. Both variables were linearly related for the small NSW lagoons (p = 0.0106, R2=0.24, with a slope of 0.740 f 0.267). The relative surface covered by seagrass was not correlated to the yield unless Laguna Madre/Baffin Bay (269, Texas, USA) was discarded as an outlier (Eq. (42) in Table 7). No regional relationship was found for either NSW or Gulf of Mexico lagoons (outlier removed). No linear or curvilinear relationship was found between yield and either wetland surface areas. The significance of the multilinear relationship ( W plus % W , Eq. (43) in Table 7) was impaired by the strong correlation among dependent variables (p = 0.0003). No significant relationship, linear, curvilinear or multilinear, was demonstrated when the data set was split according to geographic location. Residuals (Res) and biological parameters were strongly correlated, and the relationships described increasing yields with increasing area of submerged (SG) or emergent ( W) vegetation (Figures 19 and 20; Eqs (44) and (45) in Table 7). Only SG remained correlated with Res when both independent variables were included in a multilinear model.

de

-.5

0

.5

1

SO (lambda

1.5

2

2.5

3

= -0.31549)

Figure 19 Correlation between area of immersed vegetation (SG) and Res in 41 lagoons. The dots above the zero-ordinate line indicate yields higher than average for their size and latitudinal location. The regression line from Eq. (44)(Table 7) is overlaid.

144

J.-C. JOYEUX AND A. B. WARD

de -1.5.

it

0 0

- 2.

0

0

-2.5.

.

0 0

I1

W (lambda I -0.15061) Figure 20 The scatterplot of Res vs. wetland area (W)for 49 lagoons shows the association between the yield and the surface covered by emerged vegetation. Equation (45) is overlaid.

5.5. Fishing Effort and Catch per Unit Effort

As expected, the fishery yield was strongly dependent upon the fishing pressure on the lagoon system, both as absolute effort (Eq. (46) in Table 7) or as effort relative to the surface area (Figure 21; Eqs (47) and (48) in Table 7). Attempts at including other dependent variables (UE, Lat, B, Area) generally failed. Area could be added to Eq. 46 (Eq. (49) in Table 7). There was no clear decline in fishery production at highest density of fishermen (Figure 21; Eq. (48)). This feature probably resulted from the effect of higher densities of fishermen at lower latitudes (Eq. (50) in Table 7), possibly combined with a lower fishing efficiency at these latitudes. While Area was not linearly correlated with the density of fishermen, the multivariate model including Area and Lat was barely non-significant on Area (partial p = 0.0570). NbF and NbFIArea respectively explained 16% and 7% of the variation of Res (Eqs (51) and (52) in Table 7). There is no doubt that the relatively poor relation between Res and the fishermen density (Eq. (52)) results from the hidden inclusion of the effect of density into the computation of Res from the latitude (Eq. (50)), and possibly area, data. The yearly yield per fisherman was solely dependent upon the density of fishermen, and data could be fitted with either a linear or a curvilinear model

CONSTRAINTS ON COASTAL LAGOON FISHERIES

145

-

A

+ s .%! L -C

In(NbF/Area) Figure 21 Yield per unit area as a function of fishing pressure, i.e. density of fishermen (NbFIArea) in 67 lagoons worldwide. The relationship referred to in the text (Eq. (48); Table 7) is superimposed.

(Figure 22; Eqs (53) and (54) in Table 7). Unsurprisingly, the catch per unit effort (i.e. T/F)increased when the total effort, the number of fishermen per unit area, decreased.

5.6. Discussion

In general, fishery yields in lagoons were independent of freshwater inputs and related variables (RIArea, BA, BAIArea), and this result was expected. We suspect that, regionally, models including such variables may have proved satisfactory because climatic conditions would be more similar. To our knowledge, only Moreau and De Silva (1991) have shown a link, in some freshwater systems, between fishery yield (total or per unit area) and BA or BAIArea. However, these two variables are more related to the potential input of freshwater or nutrient, than to the input proper. In that respect, our results show that, overall, there is no evidence that freshwater inputs are related to fishery production per se. In a single lagoon or coastal area, however, the influence of freshwater inputs was demonstrated in numerous occasions. Copeland (1966) discerned trends relating the yearly catch of commercial fisheries to the annual freshwater inflow for the previous year in five Texas lagoons. Barret and Gillepsie (1973) noted “an

146

J.-C. JOYEUX AND A. 8. WARD

44

-

3-

b

0

0

0

’

2. 1-

c

0..

-Ec -- 2.1. - 3.

0

om ’

- 4.

-5.

0

-

-

-

.

’

r

Figure 22 Change in catch per unit effort according to the number of fishermen per unit area. The curvilinear regression (Eq. (54) from Table 7) is overlaid.

inverse relationship between shrimp production and river discharge” of the Mississippi into the Gulf of Mexico (white shrimp Penaeus setiferus, caught in offshore and inshore areas). Glaister (1978) positively related the annual oceanic catches, but not the estuarine catches, of the school prawn Metapenaeus macleayi to the annual freshwater discharge for the same year of the Clarence River (196, Australia). The Texas Department of Water Resources (1979 and 1981a,b) developed numerous and complex models to correlate the freshwater inflows entering Galveston (274, USA) and Sabine (275, USA) in to the fishery catches of various shellfish, crustacean and fish species. More recently, Yaiiez-Arancibia et al. (1985) demonstrated significant relationships between the river discharge and the commercial yield of the Gulf of Mexico fisheries landed in the states of Tabasco and Campeche. In both states, the yield of fish, shrimp and other macroinvertebrates was positively correlated with the annual river discharge. In a model integrating time series from three contiguous states, these authors related the annual yield per unit area to the riverine discharge Y = -7.364 loge( LE.Area)

+ 0.626 x lo&(Discharge)

(55)

(n = 19 with n = 5 for Veracruz and n = 7 for both Tabasco and Campeche, R = 0 . 9 8 , ~= 0.001)

CONSTRAINTS ON COASTAL LAGOON FISHERIES

147

with Y in lo3kg, LE.Area the surface area of lagoons and estuaries in km2, and Discharge in 103m3. It is our conclusion that freshwater inflows into estuarine or coastal ecosystems modulate, or adjust, the fishery production in accordance with the environmental requisites of the species or groups of species that are commercially important. On a multiregional basis, information on freshwater inflows is almost irrelevant to the management of estuarine fisheries (Figure 14; Eq. (31)). There is, however, no doubt that time series of freshwater inputs present an interest for fishery management in individual waterbodies, and that species-specific relationships can be successfully applied to similar waterbodies, at least regionally. Because numerous estuarine species of economic importance are transient, i.e. reside in lagoons for a short period of time, typically a few weeks to a few years, we expected a strong relationship between yield and a variable representing, or linked to, the possibilities to locate, penetrate into or leave the estuarine area. When the communications between lagoon and ocean are considered, this variable would be TO. Since we found no such direct relationship, we can conclude that the length of open shoreline, absolute or relative to the lagoonal area, allowing passage to and from the ocean, is not by itself a main determinant of the fishery yield. However, the combined influence of tide and inlet(s) size (Eq. (30)) strongly affects the yield, and overall choked lagoons (Kjerfve, 1986) in regions of low tide are better producers. This indicates that Kjerfve’s (1986) classification of lagoons, based on hydrological and oceanographic features, may have biological implications. Literature on the effects, sometimes beneficial but often devastating, of inlet opening or enlargement is relatively common (Bird, 1982). The interested reader can refer to Olsen and Lee (1982), who detailed the disastrous changes in the hydrodynamics and the fisheries which followed the stabilization of Charleston Pond’s seasonal inlet (291, USA). The effects of tide and opening upon the colonization processes of estuarine-dependent species of marine origin remains speculative because the proportion of anadromous, catadromous and resident species in the catches was not recorded. Nonetheless, as Eqs (28) and (30) suggest, tidally induced movements of water may be inconsequential, or even detrimental, to the colonization or retention processes for immigrating organisms of marine origin. Evidence obtained from time series of freshwater inputs vs. fisheries yields in individual bodies of water, lagoons, estuaries, or coastal waters suggested to us a correlation between salinity and yield. Against all expectations, the effect of salinity on lagoon fisheries was, at best, marginal (Eqs (36) and (37)). However, the possibility of high yield was reduced with minimum salinity above 15% (Figure 17). To some extent hypersaline lagoons were not especially unproductive, owing to the continuous inflow of water and

148

J.-C. JOYEUX AND A.

B. WARD

organisms from neighboring oceanic areas into the least saline part of the lagoon (Hedgpeth, 1957; Barnes, 1980). Literature available on the effects of temperature on coastal catches shows that authors either have concentrated their efforts on invertebrates or have not reported negative results on fishes. Most publications use time series (Fogarty, 1989). The effect of temperature was generally perceptible on catches after several months or years (Flowers and Saila, 1972). In freshwater environments, a more general approach to the problem was conducted by, among others, Schlesinger and Regier (1982) who demonstrated that, in intensively fished lakes, the maximum sustainable yield per unit area was linearly correlated with the mean annual air temperature. For a group of lakes presenting a more variable fishing effort, this MSY was described by a multilinear equation including temperature, fishing effort and MEZ (for MEZ, see Ryder et al., 1974). For comparison, these authors also transformed Turner's (1977) latitude data into temperature data, recomputed Turner's (1 977) relation between stabilized commercial shrimp yield and temperature, and obtained logl,,(%)MSY

= 0.071 x Temp - 0.042

(n = 27, R2 = 0 . 3 4 , ~< 0.05)

where MSY is the maximum sustainable yield (average landings in high, stabilized effort, in lo3t), U r e a the surface of intertidal vegetation (salt marsh macrophytes and mangrove, in 103km2)and Temp the mean annual air temperature (in "C).The R2 of the new relationship (0.34) was less than for the original equation (0.54) derived from latitude data. Schlesinger and Regier's (1982) relationship between yield and mean air temperature is equivalent to ours on minimum water temperature (Eq. (34)). The small number of data these authors used may have, however, hidden a more complex dependence. We previously discarded such a complex model (Eq. (35)), but we doubt whether the fourth-order curvilinear fit could be an artifact. In Section 4, we briefly discussed the curvilinear fit relating yield and latitude and assumed that, possibly, changes in fauna, ecology or seasonality may induce a non-linear response to latitude. The same reasoning applies here. A global synthesis similar to ours could demonstrate strong interactions between the fishery yield and the fauna and ecology of lagoons and estuaries. The relatively poor dependence of the fishery yield on the wetland area (Eqs (43) and (45)) was surprising because positive relationships had been previously demonstrated by various authors in diverse regions of the world, although most studies did not relate to catches in estuarine areas but to

CONSTRAINTS ON COASTAL LAGOON FISHERIES

149

catches on the continental shelf. We have already briefly discussed (Section 4) the findings of Yaiiez-Arancibia et al. (1985), Turner (1986) and Pauly and Ingles (1986). A similar study on regional catches of estuarine-dependent fishes showed equivalent results (Nixon, 1980), i.e., the yield per unit area is positively related to the relative or the absolute surface covered by intertidal vegetation. Why yield would increase with the absolute wetland area and “simultaneously” decrease with the relative wetland area is not clearly understood (Eq. (43)). This is probably a sampling artefact owing to the higher yields reported from large North American lagoons than those reported from small New South Wales estuaries. In addition, the Australian landing data were of greater accuracy, with all UE values at zero. When the effects of latitude, area and under-estimation were accounted for, areas covered by seagrasses and wetlands respectively explained 35% and 14% additional variability (Eqs (44) and (45)). Both relations indicated a favorable effect of aquatic vegetation on fishery yields, though the question remains of how representative was our sample. Another biological parameter of interest is primary productivity, which has previously been positively correlated to fishery yield per unit area in most, if not all, aquatic ecosystems (Marten and Polovina, 1982; Nixon, 1981, 1982; Nixon et al., 1986; Moreau and De Silva, 1991). No data for primary productivity or chlorophyll a standing crop for phytoplankton were compiled for this study. These variables are closely related and dependent upon nutrient availability (Gilmartin and Revelante, 1978; Monbet, 1992). Because nitrogen input from freshwater generally cannot support the primary production observed yearly (Nixon, 1981; Peters and Schaaf, 1991), it has been inferred that the complex food chains based on detritus may be extremely influential (Boesch and Turner, 1984). In any case, we expected that by bypassing phytoplankton production, or chlorophyll a, we could directly link nutrient concentration to fishery yield (Iverson, 1990) and, possibly, quantify the influence of man-made eutrophication (Kapetsky, 1984). However, with one exception, we failed. Perhaps this failure is a result of the inherently ephemeral variations observed on both temporal and spatial scales (e.g., Hodgkin and Birch, 1982; Dufour and Lemasson, 1984; Bono and Vigna Taglianti, 1985). It is significant, in our opinion, that the relationship demonstrated involved the mean concentration of a toxic compound (Eqs (38x40)). Obviously maxima, which are expected to vary more from year to year than means, are problematic when the periods of record for the various variables do not perfectly match. Although the toxicity is expected to vary inversely to the salinity (Crawford and Allen, 1977), it may explain why we obtained a second-order curvilinear relation (Figure 18; Eqs (38) and (40)). It is commonly agreed that greater tidal amplitude accentuates vertical mixing of waters and increases sediment resuspension. Nixon (1988)

150

J.-C. JOYEUX AND A. B. WARD

hypothesized that the physical inputs of energy through tides and, on a smaller scale, wind are responsible for greater fishery production of bays and estuaries compared with lakes (Kapetsky, 1984). By generating vertical mixing and reducing or eliminating stratification, physical inputs “enhance[s] the growth of heterotrophic bottom communitiesand promote[s] the efficient return of nutrient to the water column”. In the lagoons of the Gulf of California, where the tidal range is small (about 1 m), both chlorophyll a standing crop and primary production rates were positively correlated to tidal exchange (Gilmartin and Revelante, 1978). Monbet (1992) demonstrated that, at equivalent nutrient concentration of dissolved inorganic nitrogen, estuaries with a tide < 2 m had greater chlorophyll a concentration than macrotidal estuaries with tide > 2 m. However, one cannot expect equivalent flushing times and tidally induced turbulence in open river mouths and in comparatively enclosed lagoons (cf. Boynton et al., 1982). Our results clearly suggested a curvilinear effect of tide on the fishery yield (Figure 15; Eq (32)). Through what mechanisms tide relates to fisheries is still a matter of speculation. Interaction pathways existing among tide, length of open shore, nutrient inputs and concentrations, phytoplankton productivity, and ultimately fishery yield clearly need to be clarified because circumstantial evidence indicates that these parameters are linked. Fishing pressure was the single most important variable affecting the yield per unit area (Eqs (46)-(49)) and the yield per unit effort (the unit of effort being one fisherman year; Eqs (53) and (54)). Similar relationships had previously been demonstrated by Kapetsky (1984; Eq (57) below was recomputed): Y = 25.6+ 18.4 x

NbF

(57)

(p < 0.001, R2 = 0.78, n = 42)

and lo&(T/F) = 1.371 - 0.392 x lo&

-

(T:ez)

(58)

(R2 = 0.43, n = 42) with Y in kg ha-’ yr-’, Area in km2, and T/Fin t fisherman-’ yr-’. Equation (58) is comparable to Eq. (53). Applied to the present set of lagoons, the first type of equation produced the parameters a = 25.7, b = 12.8f2.22 and c = -0.102 f 0.022, and a R2 = 0.42. Bayley (1988) failed to fit an equivalent model for 13 African lagoons, probably owing to the small sample size. His model, however, was easily applied to our data (Eq. (59) in Table 7), and

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the resulting parameters were close to the ones he had estimated for African lagoons. It is remarkable that globally, although fishing activity was extremely diverse, yield was so much dependent upon fishing pressure. It would make no sense to assume that the maximum sustainable yield per unit area in lagoons lies close to or beyond the top of the catch-effort curve (Figure 21). Nevertheless, estuarine fisheries are apparently able to withstand considerable pressure without obvious ill effects, probably because lagoons are not isolated entities but lie at the interface of the ocean and the continent and are, in large part, populated by transient species. 5.7. Conclusions

In Section 4, we demonstrated that fishery yields in coastal lagoons were dependent upon, or correlated to, their geographic location and morphometric features. In the present section, the influence of several environmental parameters was demonstrated. Overall, the yield is controlled by winter temperature and is limited by higher salinities. Tidal exchanges of water, as expressed by the height of the oceanic tide and the openness of the lagoon to the ocean, are instrumental to fishery production. It is not understood if this influence is accomplished directly through advection or flushing of colonizing organisms or more indirectly through nutrient retention and recirculation within the estuarine area with subsequent phytoplankton production. Emerged and submerged macrophytes are adequate markers of suitable environments for commercial species. However, the most influential parameter on yield is fishing pressure. From the point of view of management, fishing pressure is also the factor most easily controlled, at least theoretically. In that respect, we suggest that lagoons isolated from other estuarine biota may be especially sensitive to overfishing of catadromous species. In clustered lagoons, fishing pressure is probably diverse and repopulation could proceed more readily.

6. FINAL CONCLUSIONS

Without doubt our compilation will be expanded in the future in order to complete, extend or criticize our analyses. Researchers will then be confronted with the same underlying obstacles we encountered, e.g. data dispersion, scarceness, unavailability, unaccountability, absence of time-match with other recorded variables, and non-suitability owing to partial records or non-translatable units. The absence of time-match was probably a powerful generator of statistical noise. Obviously, studies lack homogeneity of

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format in which data are reported, and most works lack focus on how they could fit into a broader perspective. These difficulties were particularly obvious for parameters that display strong temporal and spatial variations such as nutrient concentrations and, to a lesser degree, salinity. Representativeness of the data on both temporal and spatial scales is required to assess fully the correlations between these variables and the yield of the fisheries. This is an especially acute problem for subdivided bodies of water and large lagoons, more than 100 or 1000km2.However, too much information can be difficult to analyse so we suggest future studies provide, perhaps in addition of their own needs, monthly data summaries. Spatio-temporal variations could be summarized monthly according to subunits or, for large systems, according to the pertinent spatial scale (upper reaches, middle, lower reaches). In addition, we wish to emphasize that improving the quality of existing data is no substitute for missing data. Some regions of the world have extremely poor records, or poorly accessible records. Uncorrected, this would continue to influence any global analysis. The two main components of exploited fish and crustacean populations in lagoons are the residents and the migrants. The residents are generally under low fishing pressure. The fishing activities are based essentially on the migrants of marine origin and to some extent the migrants of continental origin. The fishery uses at best the periodic migrations of these organisms between ocean, estuary and continent (e.g., Chauvet, 1988). When estuarine stocks are renewed from marine populations, lagoons can be considered to act as traps within which the trapped populations are subjected to the fisheries. Thus, it appears that the better the trap, the more successful, economically viable and important the fishery. We believe much could be learned from an analysis that would distinguish yields of resident from migrant populations and correlate these yields to the relevant variables. The directions of such further studies are quite clearly suggested by the present work. The morphometrical factors, such as area and depth, and the variables related to the immigration processes, such as the length of open shore and the height of the tide, are likely to be major determinants of the efficiency of the trap. Those factors would also play a role in explaining the efficiency of the fisheries within the lagoon because the effectiveness of passive gear depends on tidal, die1 or seasonal movements. In contrast, vegetation and physico-chemical factors affecting the yield are more likely to influence such circumstances as resident time, growth, survival, and species succession. In most respects, vegetation and physico-chemical factors would act the same way on both residents and migrants. The principles behind the morphoedaphic index used in freshwater environments may be applicable on resident populations in lagoons. These principles are that the system is closed with defined exploited populations and that, therefore, the productivity of

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the environment explains the yield. The productivity can be forecast from both morphometrical and physico-chemical characteristics. We have shown that the yield of fisheries in lagoons throughout the world is dependent, or linked, to numerous parameters. The necessity for enhancing and rehabilitating fisheries is everywhere evident. To answer that concern, two options are available: improving the fishery management of the stocks, which is out of the scope of this study, and manipulating the environment. Adequate manipulation of environmental parameters is difficult. In fact, altering the environment is not an exact science because physical action and biological response are often disjointed or delayed. Specific attempts to improve the fishery yield by modifying the environment are common in small Mediterranean lagoons (Section 4). The practices involved generally rely on natural colonization and subsequent confinement within the lagoon. Extreme alterations, such as breaking down an entire system into smaller, more productive units is being done (Rowntree et al., 1984; Toews and Ishak, 1984; Ardizzone et al., 1988) although it takes a great deal of effort. However, more environmentally sound alternatives are preferable. For purposes of fishery management, the best locations for action occur at the lagoon’s weak points: its junctions with the continent and the ocean. To manage these points would provide control over salinity, nutrients, pollutants and colonizing organisms to restore or enhance fishery production. Specific actions we recommend below may seem obvious considering the amount of literature already available on the subject. However, in most cases, their global pertinence had not been assessed. These recommendations are: improve or restore the quality, quantity, and periodicity of the riverine inputs to control salinity, nutrients, and pollutants; carefully consider the possible effects that tide, floods, and storms will have upon both biological (colonization, etc.) and physical (circulation, sedimentation, etc.) processes after inlet modification, whatever it may be (opening, enlargement, jetty building, door installation, etc.); preserve the aquatic vegetation from destruction.

ACKNOWLEDGEMENTS

This study would not have been possible without the invaluable support of John M. Miller (Department of Zoology, North Carolina State University, Raleigh, USA), GCrard Lasserre (Laboratory of Hydrobiology, University of Montpellier 11, France) and Jean-Pierre Quignard (Laboratory of Ichthyology, University of Montpellier 11, France), who welcomed the first author in their respective laboratories. We were also provided a great

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help by James M. Kapetsky (Food and Agriculture Organization, Inland Water Resources and Aquaculture service, Rome, Italy), who had started to compile data and was kind enough to forward them to us. We also thankfully acknowledge the help of C. Arellano, North Carolina State University, Raleigh, USA; M. Baez-Hidalgo, University La Habaiia, Cuba; J.-L. Bouchereau, University of Montpellier 11, France; J.J. Burchmore, New South Wales Fisheries, Australia; P.S. Economidis, Aristotle University, Thessaloniki, Greece; M. G u m a n Arroyo, University of Guadalajara, Chalapa, Jalisco, Mexico; R. Ibarra, Ministry of Agriculture, Nueva San Salvador, San Salvador; D.S. Jayakody and M.U. Jayasekara, National Aquatic Resources Agency, Colombo, Sri Lanka; K.A. Koranteg, Ministry of Agriculture, Tema, Ghana; G.L. Morales, Department of Agriculture, Metro Manila, Philippines; L.D. Padalkar, Department of Fisheries, Madras, Tamil Nadu, India; D. Rechlin, Institut fur Ostseeficherei, Rostock, Germany; P. Roy, University of Sydney, Australia; P. Sirimontaporn, National Institute of Coastal Aquaculture, Thailand; J.A. Tomasini, University of Montpellier 11, France; A.P. Tsiokas, Ministry of Agriculture, Athens, Greece; A. Yaiiez-Arancibia, Autonomous University of Campeche, Campeche, Mexico; Yu.P. Zaitsev, Institute of Biology of the South Seas, Odessa, Ukraine; A. Zerbi, University of Montpellier 11, France and the personnel of the North Carolina State University Libraries.

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APPENDIX 1 Yearly yields of Mollusc (bivalvia and gastropods) fishery and aquaculture in 109 lagoons. Lagoon number refers to the number given in Table 1. All weights are in the shell. For lagoons 269 and 271-274, the given value was estimated from meat weight data, with meat weight = 0.25 total weight. The value for the yield given in parentheses is the ratio of the landings to the area. Appendix copyright 8 1998 by the authors. Reproduction of this appendix in its entirety by any means is permitted. Lagoon

Fishing yield or activity

13 14 15 17 18 19 20 23 24 25 26 27 29 29 30 31 32 33 34 35 37 40

Aquaculture yield Total yield or activity

Years; comments

0 0 0 0 0 -

0 0 0 0 0 Yes -

540t 0 0

-

-

188438 kg ha-' -

0

0

0 0 150t 200 t 300-600 t Yes 0

c. 85 c. 85 c. 85 c. 85 c. 85 c. 72; clam, cockle c. 72; oyster, mussel ( ~ 7 kg 2 ha-' yr-I) -

c. 84; ( ~ 2 6 kg 3 ha-' yr-I) c. 84; ( ~ 2 5 kg 3 ha-' yr-I) 69-73, 79-82; mussel (~2814525kg ha-' yr-')

-

References Amanieu, 1967 Amanieu, 1967 Amanieu, 1967 Demestre et al., 1977, 1989 Demestre et al., 1977, 1989 Demestre et al., 1977, 1989 Demestre et al., 1977, 1989 Bourquard, 1985 Bourquard, 1985 Bourquard, 1985 Bourquard, 1985 Bourquard, 1985 Amanieu, 1973 Amanieu, 1973 Personal observation Amanieu, 1973 Frisoni et al., 1984 Frisoni et al., 1984 Frisoni et af., 1984 Rossi et al.. 1984 Ardizzone, 1984 Ardizzone, 1984

Appendix I (continued)

Lagoon

Fishing yield or activity

Aquaculture yield Total yield or activity

43

Yes

0

44 45 46 49

0

0 0 0 Yes 0 0 0 25t

Years; comments

References

Oyster

Maghocchetti Lombi et al., 1987-88

60 61 142 143 143 143 144 145 146 147 148 149 150 151 152 155 156 157 158 167

0 Yes 20 kg ha-' 12kg ha-'

-

70-175 t

< 1kg ha-'

-

-

Yes

-

-

Yes

-

0 0 Yes 200 kg ha-'

0 0 0 0 0 0 0 0 0

-

-

-

Mussel 68; cockle 68; cockle

-

c. 84;

oyster

(%2 kg ha-' yr-') c. 84; mussel ( ~ S 1 kg 2 ha-' yr-') c. 84; cockle

-

Cockle -

-

Ostrea tulipa present

-

Oyster, Arca c. 72; clam

Ardizzone, 1984 Ardizzone, 1984 Ardizzone, 1984 Ardizzone et al., 1988 Cottigha, 1970 Cottiglia, 1970 Zaouali, 1979; Kerambrun, 1986 Kerambrun, 1986

Zaouali, 1979; Kerambrun, 1986 Zaouali, 1984 Zaouali, 1979; Kerambrun, 1986 Zaouali, 1979; Kerambrun, 1986 Zaouali, 1984 Zaouali, 1979; Kerambrun, 1986 Zaouali, 1979; Kerambrun, 1986 Zaouali, 1979; Kerambrun, 1986 Brunel, 1985 Morgan, 1982 Frisoni et al., 1984 Pauly, 1975 Texier, 1984 Texier, 1984 Texier, 1984

Jhingran and Gopalakhrisnan, 1973

A

173

-

175 176 177 178 179

-

191

-

58 kg

192

-

141 035 kg

193

-

28 743 kg

194 198

-

Yes

>O 100 529 kg

199 200 20 1

0 0

-

202

-

21 kg

203

-

12 kg

204

-

4201 kg

205

-

25 kg

206

-

1 833 546 kg

207

-

571 kg

c. 88

Samarakoon, 1986

c. 88

Samarakoon, 1986 Samarakoon, 1986 Samarakoon, 1986 Samarakoon, 1986 Chacko et al., 1953

-

-

c. 88

0 0

0 0 1 297 249 kg

c. 88 Ostrea madradensis present 72-82; no oyster (= < 1 kg ha-' yr-') 72-82; oyster ( ~ 3 0 kg 9 ha-' yr-') 72-82; oyster ( ~ 9 kg 9 ha-' yr-')

-

72-82; mainly oyster (=11 kg ha-' yr-') 72-82 72-82 72-82 mainly oyster (%151kg ha-' yr-') 72-82; no oyster (=< 1 kg ha-' yr-') 72-82; no oyster (= < 1kg ha-' yr-') 72-82; no oyster (= < 1 kg ha-' yr-I) 72-82; no oyster ( X < 1 kg ha-' yr-I) 72-82; mainly o ster (e298 kg ha-'yr-') 72-82; no oyster (=< 1 kg ha-' yr-I)

West et al., 1985 West et al., 1985 West et al., 1985 Scott, 1989 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985

Appendix I (continued)

Lagoon

Fishing yield or activity

-. Aquaculture yield Total yield or activity

208 209

0 1438kg

210

22 992 kg

21 1

3104kg

212 213

8 kg 379 161 kg

214 215

20 kg 87 741 kg

216 217 218

l5kg 19 kg 120008kg

219 220 221

0 0 1989kg

222 223

0 44422 kg

224

6834 kg

225

-

225

-

Years; comments

References

72-82 72-82; no oyster ( ~ 3 . kg 5 ha-' yr-') 72-82; mainly oyster ( ~ 5 kg 3 ha-' yr-I) 72-82; mainly oyster ( ~ 7 . kg 4 ha-' yr-' 72-82; (X< 1 kg ha- )yr - ') 72-82; mainly oyster (X190kg ha-' yr- I ) 72-82; (X< 1 kg ha-' y-') 72-82; mainly oyster ( ~ 6 kg 6 ha-' yr-') 72-82; (= < 1 kg ha-' yr-') 72-82; (X< 1 kg ha-' yr-I) 72-82; mainly o ster ( 4 9 0 kg ha-'yr-') 72-82 72-82 72-82; mainly oyster ( ~ 2 . kg 6 ha-' yr-') 72-82 72-82; mainly o ster ( ~ 1 3 kg 9 ha-'yr-') 72-82; oyster ( ~ 9 kg 6 ha-' yr-') 70; oyster ( ~1 kg 3 ha-' yr-') 69-7 1; clam

West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 West et al., 1985 Percy et al., 1974 Percy et al., 1974

a0 h)

< 1 kgha-' < 1 kgha-'

< 1 kg ha-'

226 227 24 1

2-19 kg ha-'

-

242 243 252 258

0 < 1 kg ha-' 1168kg ha-' 7.2 kg ha-'

-

258 259 260 261 268 269 270 27 1 272 273 274 275 278 281 284 285 286 287 29 1 29 1 292

< 1 kgha-'

-

-

7.2 kg ha-'

>O

>O -

263 kg ha-'

=O

-

-

Yes < 1 kg ha-'

0 2 kg ha-' 9 kg ha-' 7 kg ha-' 41 kg ha-' 0 -

0

-

< 1 kg ha-'

0 6 7 4 kg ha-'

5 kg ha-' >O < 1kg ha-' 0 2 kg ha-' 9 kg ha-' 7 kg ha-' 4 1 kg ha-' 0 2 0 < 1 kg ha-' < 1kgha-' 0 -

-

-

132kgha-'

-

-

"-"indicates

2-1 55 no data.

7&71; mainly clam 69-71; clam 78, 90 90 90 54-64 73-85; Crassostrea virginica, C. rhizophorae 3 species of snails 67-7 1; mainly oyster c. 1979 77 c. 1988 68-73 68-73 68-73 68-73 65-74 68-73 66-75 64-72

-

69-62 70-85 70-85 75/76; clam 79; oyster; (x4 kg ha-' yr-') 79; oyster, scallop, clam 79; depending on the area; clam, scallop

Percy et al., 1974 Percy et al., 1974 El Salvador, 1979; Ibarra, pers. comm. Ibarra, pers. comm. Ibarra, pers. comm. Pereira-Barros, 1967 and 1969 Yaiiez-Arancibia and AguirreLion, 1988 Reskndes Medina, 1979 Garcia, 1975; Resendes Medina, 1979 Resendes Medina, 1979 Martinez Mata, 1980 Lopez et al., 1988 Hefferman and Green, 1977 Breuer et al., 1977 Hefferman and Green, 1977 Hefferman and Green, 1977 Breuer et al., 1977 Breuer et al., 1977 Breuer et al., 1977 Swingle, 1976 Lewis and Estevez, 1985 Anderson and Gehringer, 1965 Epperly and Ross, 1986 Epperly and Ross, 1986 Hillman and Kennish, 1984 Olsen and Lee, 1982 Olsen and Lee, 1982 Olsen and Lee, 1982

A

00 0

APPENDIX 2

2

a0

P

Collected data. Lagoon number refers to the number attributed in the Table 1. Consult Table 2 for abbreviations and units (data are not transformed). Appendix copyright 0 1998 by the authors. Reproduction of this appendix in its entirety by any means is permitted. Lagoon

1 2 3 4 5 6 7 8" 9" 10" 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25

Y

61 32' 20 16 55 354' 34' 43 21 5 29 80 139 95 125 81 89 64 177 76 26 209

40 103 62

UE

NbF

Area

0 0 0 0 0 0 0 0 0 0 0 0 -1 -1 -1 0 -1 -1 -1 -1 0 0 0 0 0

-

47 1 164' 34 108 77 510' 277' 100 169 12.5 39 1.6 0.35 1.43 0.4 31.1 5.6 4 2 0.21 170 0.53 12 18 3

-

-

-

-

-

22 20 -

-

30 20 18 2 167

-

10 43 10

Lat

53 54 54 54 54 54 53 54 54 54 54 54

44 44 44 39 41 41 41 41 38 41 43 43 43

B

Z-

Zm-

Tmin

Tmax

Smin

Smax

30 30 30 30 30 30 30 30 30 30 30 30 50 50 50 70 70 70 70 70 80 80 50 50 50

-

-

-

-

-

-

2.6"

0" 0" 0" 0" 0" 0" 0 0 0 0 0

20" 20" 20" 20" 20" 20" 20 20 20 15 20

16"

3"

5"

-

-

2.3' 3.9' 7" 2.4" 9 11 1 3 1.8

4.6' 16.7' 8" 3" 9.8 16 11 18 20

-

-

12 6 13.5" 7.8" 16 12 11 7 2.8 2 2 2 4 2.5 0.8 0.8 1 7

2.1 2.1 5.8" 3.7" 3.1 6 4 2.6 1.3 0.6 0.6 0.6 2.5

-

-

-

-

-

-

1.4 1.6 3.2

0.7 0.6 0.8

2" 2" 2"

28" 28" 28"

15 3 21

49 37 38

0.5 0.7 0.4 0.7 4.3

-

-

-

4

27

4

-

29

-

-

-

-

12.3 3 6.5 3.8

30 27.6 27.3 25.4

0.2 0.6 2 2

0.4 29.5 27.1 36.1

-

-

-

6

30.5

47.1

-

52.9

26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58

58 126 53 100 189 158 113 11 25 100 96 167 47 33 65 16 31 100 406 44 150 51 92 356 63 15 5 689 93 225 167 390 416

0 0 0 0 0 0 -1 0 0 0 -1 -1 -1 0 0 -1 0 -2 -1 0 -1 -1

-1 0 -1 -1 -1 0 0 -1 -1 0 0

43 72 6 -

32 -

15 3 3 150 70 16 110 215 2 60 -

-

-

-

-

-

1.36 54 5.8 75 31.7 150 15 5.7 7.9 32 27 3.9 51 70 1.4 100 0.45 12 0.95 3.8 4.1 19 86 2.2 0.9 1.6 1.9 8.3 0.95

20 0.97 2 1.8

43 43 43 43 43

44 43 42 42 45 42 41 42 42 42 44 40 45 41 41 41 44

44 40 39 39 39 40 39 39 40 39 39

50 50 50 50 50 50 60 60

60 50 50 50 50 50 50 50 70

50 70 70 70 50 60 70 70 70 70 70 70 70 70 70 70

1.1 3.9 1.5 10 1.3 10 3 9 11 4 1.5 13 1.5 5 2.8

0.4 1.9 0.4 4.5 0.8 5.8 1.8 2

2" 7 5.3 5 2.6

28" 25 35 26.6 26.3

-

-

8 8 8 0.4 6.1 3 5 8 5 3.5 5 0.1 5 8" 5

30 28 28 30 28.6 30 30 27 30 29 29

-

-

1" 21 2"

0.8' 9 1 0.5

-

-

-

24 3 18 25.5 3 3 5.5 30 32 4 27.7 12 5 16 27.7' 26 12 3.5 17.3 3 26.5 5

-

-

-

2 2 2 2 1.5

1 0.7 0.7 0.7 1.2 0.6

-

-

-

10 10 10 9

-

27.5 27.5 27.5 24

-

36 42 45 2.9

1 1

-

-

-

-

-

-

6.3

-

30

0.6

-

-

-

-

1.8

-

-

-

-

2 1.1 4.2' 0.8 2 0.8 1 1.5

-

30 29" 30

-

-

-

32 34 40 38.6 37 15 25.5 38 38 31 47.4 29 28 20 47.7b 48 29 37 38 22 45.5 22

-

42 58 68 36.4 77.6 -

39.6

-

A

00

cn

Appendix 2 (continued)

Lagoon

Y

UE

59 60 61 62 63

120 120 131 60 120 370 54 120 140 110 100 20 178 395 156 174 62 88 67 53 86 53 44 179 57 130 615 60 22 1

0 0 0 0 0 0 0 0 0 0 -1 -1 0 -1 0 0 0 0 0 0 0 0 0 -2 0 0 0 0 0

64 65 66 67 68 69 70 71 72 73 74 75 76 77 78 79 80 81 82 83 84 85 86 87

2

m

Area

-

0.38 2.5 8 26 0.75 0.3 2.2 3.4 3.8 10 145 12 3.41 2 5 1.5 7.5 2.8 4.9 3.1

39 40 39 39 39 39 40 40 40 38 38 39 38 40 38 38 38 40 40 41 40 40 40 40 38 39 39 38 39

-

0.5

1.5 3 1.4 2.1 1.2 0.13 0.45 1.7

Lat

70 70 70 70 70 70 70 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80

-

0.2 0.3

-

0.8 1.5

-

0.4 0.5

1.5 0.3 1 0.4 0.8 0.4 0.4 0.7

-

0.6

-

1 1.5

88 89 90 91 92 93 94 95 96 97 98 99 100 101 102 103 104 105 106 107 108 109 110 111 112 113 114 115 116 117 118 119 120

232 81 94 160 133 53 50 100 100 29 21 20 30 50 60 68 53 72 100 40 42 62 55 27 14 163 13 67 244 66 65 10 158

0 -1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

0 0 0 0 0 0 0 -1 0 0 0 -1 -1 0

-

-

-

-

-

-

76

-

-

1.1 29 35 1.9 0.45 0.75 0.54 0.1 1.5 5 3.08 3 2.3 1.3 0.35 4.3 47.5 0.3 0.4

3 2.8 2 2 66 29 15 60 3 1.6 3 11 15 0.95

39 39 39 39 37 37 37 37 39 39 39 39 40 38 39 36 41 39 38 40 40 40 40 37 38 41 39 40 41 40 41 41 41

80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80 80

80 80 80 110 80 80 80

-

4 2 1

1.2 -

1 -

-

4.5 1.5 1 4 -

-

2 3.7 3 -

0.7 1

-

23 2

-

0.5 1.5 0.5 0.6 0.7 0.6 0.5 0.6 0.5 0.5 0.4 0.4 0.3 0.6 1.5 0.7 2 1 2.5 0.4 0.5 0.5 0.5

-

1

-

1

-

-

1.5 25 1

-

-

-

-

12 12 12

26 25 27

19.4 26 26

23.6 31 31

-

-

-

-

-

-

-

-

-

-

-

12 10 5 9 7 12.8

26 29 36.9 30.4 32.5 26

16.5 24.6 12.9 36.1 13.8 13.3

26.3 37.1 47 39.1 37.4 33.9

-

-

29

-

-

14 4.1

-

4.1

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

-

9

26

25

-

-

-

-

-

40

5 -

-

-

27

-

18 -

-

6.5 6.8 5.3

23.5 25 28

0 5 23

27 40 37

-

35 a

5

Appendix 2 (continued)

Lagoon

Y

UE

121 122 123 124 125 126 127 128 129 130 131 132 133 134 135 136 137 138 139" 140 141" 142 143 144 145 146 147 148 149

54 113 11 21 50 14 59 18 56 20 24 132 40 31 336 656 375 lgb 1251 66 30 11 2 17 124 5 8 20 54

0 0 0

0 0 0 -1 0 0 0 0

0 0 0 -1 -1 0 0 0 -1 0 -1 0 0 -1 -1 0 -1 -1

4

83 NbF

Area

Lat

B

Z-

37 3.6 28 47 12 14 55 420 1006 60 113 86 2560 650 848 656 115 2Oob 27.3 8.6 21 95 150 30 42 500 100 230 1.73

41 41 37 37 36 36 37

80 80 80 80 80 80 80 80 50 50 50 50 50 90 90 90 90 90 90 90 90 90 90 90 90 90 90 90 90

5 5 2 5 2

46 46 46 46 46 46 31 31 32 31 31 31 37 36 37 37 37 37 34 36 33 35

-

32 2.4 12 15 21.2 14 3.2 3 1.2 2 1.5 12b -

5 6 2 12 2.5 3.6 10 2.3 6.8 2.5

Zm-

Tmin -

T-

Smn

-

-

1.8 4.4 3.3 4.5 4 0.8 1 1 1 0.8 0.6 1.3 3 0.5 1.2 2.6 1 0.8 2.5 0.8 4 0.5

0 -0.9 -0.8 -0.7 0 0 10.5 14.6 12.5 12.3

32 32.2 30 31 33 35 34 31.9 29 28.5

-

-

12.7 10 11 9.5 10 11 9 12.1 4 6.6 12.8

29 30

-

29.5 29.5 29.5 30 29 38 30 28

S-

260

-

-

-

0 0.8 0.4 0.3 5 0.3 18 38 1.7 1.3 1.3 38 5 20 0.1 5 10.9 30 34.9 40.6 1 38 40

35 15 18 13 17 9.7 144 71 28.4 34.4 18.8 90 8 30 0.3 32 39.7 45 46.9 55.1 100 48.5 50

-

150 151 152 153 154 155 156 157 158 159 160 161" 162 163 164 165 166" 167 168 169" 170" 171" 172 173 174 175 176 177" 178" 179 180 181

20 67 180 163 186 150 861 821 278 107' 85 76 32 139 104 26 29 1 3 10 68 18 80 239 94 2244 49 15 25 100 5 22 36 35

0 2 1 -1 1 1 -1 -1 -1 -1 1 0 -1 1 0 0 0 0 0 0 2 0

0 0 0 1 2 0 0 0 0 0

-

2995 4700 1000 20 8200 1lo00

-

845 9200

-

600

-

-

27977

-

2000

-

260" 300 40 1025

-

1lo00

115 15 230 566 424 1 85 157 14 46.6' 965 26 98 23 14 23 78 500 3 3.6 2.4 4.7 27.6 0.65 237 35 1.97 4.1 1 46 1 330 1036

35 34 5 5 5 6 7 6 6 6 7 6 21 25 23 25 31 9 9 13 12 12 11 7 7 7 6 6 6 14 16 20

90 100 110 110 110 90 90 90 90 90 100 100 100 100 100 100 100 110 110 110 110 110 110 110 110 110 110 110 110 120 120 120

8 4.5 8" 23 17 1 3 3

4.7 0.5 3 4.8 3.8 0.5 1.5 1.7 -

-

-

-

-

1.8 6

1 3

23 26

34.8 31.5

0.3 0.1

18.6 26

-

14 12.8 24 27.4 26.7 18 25 25

28.7 28.7 32 31.2 30.7 34.3 31.5 31.5

30 1.5 0 1

0 0 2.6 0

-

-

-

-

-

14.2 5

4.9 2 3.2 2

21

34

0

-

5

-

-

-

-

3 3.5 4 1.5 8.5 3 3 3" 2.2 2.5 5

1.5 1.6 1.75 0.5 2.8 1 0.5 1" 1 0.75 1.5 3 2.2

-

3.7

-

23

-

31

-

-

25 27 24 23.5 26.5 27.5 28 27.6 25 25" 25 25 24.6 29 18

31 32.4 35 32.5 32.5 34.5 34 30.8 35 3 1" 32 30.5 35 35 32

-

40 36 25 23 10 46 32.9 35

-

2.9

-

-

-

1 25 2.3 2 26.2 2.5 0.02 20 0.1

32 45 33.2 31.5 34.7 40 2.5 36.4 34.5

-

-

0.7 0 0.2

57 34 36

-

A (0

Appendix 2 (continued)

Lagoon

Y

UE

182 183 184 185 186 187 188 189 190 191 192" 193 194 195 196 197 198 199 200 20 1 202 203 204 205 206" 207 208 209 210

51 21 110 2 18 5

0 0 0 0 0 0 0 0 0

35

6 4 12 0.1 16 67 461 46 454 122 3 3 57 66 13 43 53 35 47 22 11 22

0 0 0 -2 1 0 0 0

0 0 0 0 0 0 0 0 0 0 0 0

0

NbF

Area

Lat

B

137 18 24 68 51 16 373 1835 629 3.7 4.6 2.9 1040 78 10 217 89 0.18 5.8 86 9.4 106 115 70 62 36 39 4.1 4.3

32 35 33

70 70 70 70 70 70 70 70 70 70 90 70 110 100 100 70 90 80 80 80 80 80 70 70 90 70 70 70 70

35 35 35 38 38 38 36 36 36 7 16 11 36 29 30 31 31 32 32 33 33 34 34 35 35 35

Za

zm-

Tmin

Tmax

Smin

Smx

48 36"

-

33.5 33 35 20.7 35 21.8 -

0.3

-

36.1 49 -

38.5"

-

21 1 212 213 214 215 216 217 218 219 220 22 1 222 223 224 225 226 227 228a 229 230 23 1 232 233 234 235 236 237 238 239 240 24 1 242"

28 30 5 22 17 36 68 7 57 43 22 23 7 4 9 8 14 30 16 26 88 74 21 24 117 34 53 31 8 7 25' 59

0 0 0 0 0 0 0 0 0 0 0 0 0 0 2 2 1 2 2 2 -2 2 2 2 -2 0 2 2 2

2 0 0

-

-

-

-

-

-

-

-

-

4.2 3.2 20 6.3 13 1.7 1.4 6.3 1.7 0.64 7.8 0.81

3.2 0.71 36 9.7 39 48 132 38 70 175

360 530 110 113 140 140 35 4 184' 17

35 35 35 36 36 36 36 36 36 36 36 36 36 36 45 45 43 16 27 27 22 22 24 25 22 16 22 22 22 22 13 13

70 70 70 70 70 70 70 70 70 70 70 70 70 70 50 50 50 100 110 110 110 110 110 110 110 110 110 110 110 110 110 100

-

3.4 10

-

12 7

7.3 7 9 25.9 -

-

-

2.5

22.5

-

-

16.6' 4

22 14 21 33.6 31.6

32.1 29

0.1 13.9 0.5 1.2 -

-

0

33.9 34 31 4.6 36.7

38.6 13

A

Appendix 2 (continued)

Lagoon

Y

UE

243" 244 245 246 247 248 249 250 251 252 253 254 255 256 257 258 259 260 26 1 262 263 264 265 266 267" 268 269 270

4 4

0 2 2 2 0 0 0 2 2 2 0 0 0 0 0 0 0 -2 0 0 0 1 2 2 -1 -3 0 0

2 6 18 121 168 35 3 97 128 49 16 107 117 69 46 176 37 506' 97 28' 174' 593' 220 500

16 12

(D h)

NbF

Area

Lat

B

60 94 80 3770 9730 171 20 100 796 23 54 22 14344 63 450 2500 659 500 2156 1 0.18 2.2 2.8 0.4 0.69 1.2 1219 516

13 35 35 33 31 29 27 25 22 10 10 10 10 10 11 18 22 18 25 21 21 21 21 21 21 18 27 28

70 70 80 80 80 80 90 90 120 110 110 110 110 100 110 100 110 100 110 110 110 110 110 100 110 90 90

100

Zmax

Zm-

Tmin

Tmax

-

8.6

5 -

-

20 7 8.7' 20

5 0.7 1.7' 6

-

-

-

-

11 18 15' 19.8

-

25 24 24" 25 26.4 27 22 15 15 22.5 21.2 24.6" 17.7

27.7 31 37 37" 30.6 33.8 33 31 32.2 35 31 34 33.2" 31.6

8.6

2 1 0.8 25 0.8

30.5 33 92 67.5" 5.7 39.5 40 36.5 36 36

5

1 3 35 2.4 2 10 3 11 4.5

-

1.5

3 2.5 1 1.6 0.6 0.7" 0.3

Smin

Smax

-

-

-

-

-

-

-

-

-

28 27.5 30"

-

0

2

15

21 16.5" 2.6 0.5

0 0 15 0 22 7 10 13

32.8

-

32'

48

43.6 2.6 46

-

-

-

-

-

-

-

-

-

-

-

-

2.7 3.1 3.1

1.3 1.2 2.7

31

-

-

15

30

15 1.8 1

36 57 37

25

33.5

-

42.2

27 1 272 273 274 275 276 277 278 279 280 28 1 282 283 284 285 286 287 288 289 290 29 1 292

29 27 16 36 30 26 37 16 17 24 59 69 100 38 34 35 4 24 8 36 47 25

0 0 0 0 0

-

3 1 1 1 2 2 3 2 0 3 3 1 0 0 0 0 0

-

-

-

-

375 -

27 -

-

504 460 1142 1432 226 253 82 1 1070 11 452 967 78 720 800 5335 1243 167 510 250 66 5.3 6.4

28 28 29 30 30 30 29 31 30 29 27 27 26 29 35 36 40 41 40 40 40 41

"Not used in the statistical analyses on fishery yield. bModified after statistical treatment. "-"indicates no data.

90 90 90 90 90 90 90 90

90 90 90 90 90 90 80 80 70 70 70 70 70 70

4.3 2.1 4 3.1 12

-

9.1 -

3.6 13 13.8 15.3 -

7.5 9 6 27 7.7" -

2.5 8.1

1.9 1.2 2.4 2.4 2 2.7 2 3 1.2 2.5 3.5 1.2 3.3 2.1 5 5.3 1.5 6 2.7 1.8 1.2 1.8

-

12

-

29

1 1 1 1 0 4.4 1 5 14.5 0 1 18 20

33 32 31 27 30 10.8 27 22 24 35.6 33.7 37 34.2

-

-

8.9 8.9 9 6.5 8 5 11 12.2 16.5 11" 6 3 -1.4 0 3"

35.5 35.5 31 31.2 35 33 32 33 30.9 32" 29 28 28 26 26"

-

-

-

-

0 0

25.4 24

23.5 27

30 32

-

-

2

30 5 32 32 30.5"

0 12 23.5 20.5"

APPENDIX 3

A (D

Collected data. Lagoon number refers to the number attributed in the Table 1. Consult Table 2 for abbreviations and units (data are not transformed). All lagoons without any data for these variables were removed from the table. Appendix copyright 0 1998 by the authors. Reproduction of this appendix in its entirety by any means is permitted. Lagoon

NOZ-N,,

1 2 4 6 7 8" 9" 10" 11 12 13 14 15 16 17 18 19 21 23 24 25 26 27 28 29 30

-

NO2-N,,

N03-Nm-

NO3-N,,

P04-P,,

P04-P,,

-

-

-

-

-

214"

1095"

-

-

-

-

-

-

67.6" 239" 78 180

290" 975" 463 210

91.2" 447" 56 108 208 12.2 21.8

128"

591"

-

-

0.81 3.47

1.8 7.78

41.8 52.5

128 282

38.3" 194" 31 75 132 4.25 14.31

-

-

-

-

-

-

0.67 1.4 0.4 0.5

2.61 3 3 2

4.14 9.4 2.2 2.1

11.5 25 12 15

4.19

5.14 5 2 1.5

-

-

-

-

-

-

-

0.25

-

-

0.9

-

4

-

0.3

-

-

-

-

0.5

1.08 16.2

4.3 37.8

2.91 1.93

-

5 0.76

TO

R

BA

Tide SG

W

*

31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 49 56 69 89 90 91 96 97 98 99 101 103 104 108 129 130 131

-

-

-

-

-

-

0.15 -

-

-

-

-

-

-

-

-

0.88 0.807" 0.45

3.9 3.7" 2.64

4.23 1.5 0.57 2.3 0.94 11.4" 0.13

-

-

0.71 0.63 0.32 0.78 0.19 1 .2O 0.05

-

-

1100 83 100 113 6.3 13" 33

-

-

8 6.6 4.9 31.2 1.73 6.43O 8.51

-

-

-

0.89

12.7 -

56.7 9.3

0.09

-

3.42 5.1

0.42 0.24"

-

-

-

-

-

-

0.5 0.936" 0.682" 0.658"

5.78 16.3" 2" 5.54"

8.7 1.59" 20.8" 1.76"

50.5 7.6" 82.5" 6.3"

0.05 1.78" 1.71" 0.834"

0.45 3.4" 4.5O 2.1"

-

-

-

-

-

-

-

-

-

-

-

-

-

3.04" 2.61" 1.74" 1.95" 1.95" 23.2a 46.5" 31" 19"

-

-

-

-

1.3 9.99 12.2

-

-

-

7.59" 3.23" 8.07" 13.4"

13.c

-

-

-

-

4.84" 7.67" 7.42" 7.54" 75.c 1.01" 5.09" 5.09" 1.82"

-

57.6" 115" 74.9" 191"

-

-

-

-

-

199"

-

10.2"

-

282" 9.43" 15.5 2.58 2.89

-

5.65 0.646 -

-

40.4 27.1 16.8

-

-

-

2.39 0.5 0.82

0.82 -

132 134 135 136 137 140 142 143 144 145 146 148 149 150 151 152 153 154 155 156 157 159 160 162 164 167 168 172 174 175

-

3.26

4.84 0.064

0.627

-

-

73 1.79

0.71

-

-

10.5

-

19 6.5 1.4

-

5.71

-

4.99

-

-

0.126

-

240 21 35.7 22.8 4.14 16.4 11

-

-

-

-

-

-

-

1.48 1.39 1.13

8.9 10 4.9

3.16 4.6 2.22 129

9.3 33 10

-

-

-

3.14

-

1.36 0.39 0.85

14.5 0.337 61 1.74 1.9

-

-

0.19

1.94 8

-

0.73 0.81 0.93 0.95

-

1.7 2.9 2.9

-

-

0.03 0 0.2 0.35 0.765 3.6 2.5 0.042 0.19

-

0.37 0.63

-

-

-

-

-

34.3

50

2.53

-

3.68

-

-

1.47

4 36.8 2 0.42

-

-

0.93

-

-

54.3 1.71

-

6.5

-

-

-

0.25

-

-

0.69 1.2

0.5 0.7 1.7 0.12

176 179 180 181 182 183 184 185 186 188 189 190 191 193 194 196 197 198 199 200 20 1 202 203 204 205 206" 207 208 209 210 21 1 212 213

-

-

13.6 38.7

-

-

-

0.2 0.07

1.94 1.61

0.135 0.7

-

-

-

-

0.04 0.2

-

-

60 21 32

100 72 12

0.61 0.16"

-

-

-

-

4352

-

70

-

-

618"

1035 7000

0.8 0.3" 0.3 0.4" 0.5 -

-

-

-

-

840 2823 loo00 1 9700 3250 -

-

-

-

-

0.647 0.014 0.868 0.637

-

2.7 10.2

-

-

-

-

-

4.34"

1.26 -

-

0.6'

-

0.74 0.5 1.4

-

-

-

-

-

-

20000"

-

19.1 0 0.007 30.8 2.08 2.82 13.4 11.6

7.2 0.002 6 4.8 0.003 2.81 1.7 0.007

-

-

-

9000'

-

-

-

-

-

-

-

-

-

1.61

0.43

-

-

-

-

-

-

1 0.2"

-

-

399 670

-

270"

0.17" 5.12 8.54 0.587 0.527 0.508 0.509 0.092

-

-

-

-

-

-

-

0.203 0.288 0 13 0.157 0.046 3.3

Appendix 3 (continued)

Lagoon

N02-Nm-

a

N03-N,,

N02-N,,

N03-N,,

P04-Pm-

PO,-P,,

TO

R

BA

0.5"

-

23" 1700" -

Tide SG

-

-

W

~

214 215 216 217 218 219 220 22 1 222 223 224 225 226 227 228" 229 232 234 235 241 247 249 250 253 254 255 256 257 258 259

1.46"

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1.59

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260 26 1 262 264 268 269 270 27 1 272 273 274 275 276 277 278 279 280 28 1 282 283 284 285 286 287 288 289 29 1 292

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The Effects of Fishing on Marine Ecosystems Simon Jennings' and Michel J. Kaiser2

School of Biological Sciences. University of East Anglia. Norwich NR4 7TJ. UK 2Centre for Environment. Fisheries & Aquaculture Science. Conwy Laboratory. Conwy. LL32 8UB. UK (Present address: School of Ocean Sciences. University -of Wales. Bangor. Menai Bridge. Anglesey. LL59 5EY. UK)

.

1 General Introduction ............................................. 2. Benthic Fauna and Habitat ........................................ 2.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2. Direct Effects of Fishing Gears .................................. 2.3. Indirect Effects on Habitat ..................................... 2.4. Natural Versus Fishing Disturbance .............................. 2.5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3. Fish Community Structure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Diversity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. SizeStructure . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4. Life History Traits ............................................ 3.5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4 Trophic Interactions ............................................. 4.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Predator Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Prey Removal. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.4. Species Replacement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.5. Scavengers and Discards ...................................... 4.6. Reversibility of Fishing Effects .................................. 4.7. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5. Study of Fishing Effects .......................................... 5.1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2. A Statistical Basis for Correlative Studies ......................... 5.3. Investigating Marine Food Webs ................................ 5.4. Modelling Ecosystem Processes ................................ 5.5. Selection of Research Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.6. Spatial and Temporal Scales of Study ............................ 5.7. Conclusions ................................................

.

ADVANCES IN MARINE BIOLOGY VOL . 34 ISBN 0-12-0261344

203 208 208 209 227 233 235 236 236 237 246 248 256 257 257 259 265 276 283 290 291 292 292 294 295 296 299 300 302

Copyright 0 1998 Academic Press Limited A / / rights of reproduclion in m y form reserved

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6. Management................................................... 6.1. Introduction.. .............................................. 6.2. The Role of Fisheries in Marine Ecosystems ....................... 6.3. From Population to Ecosystem Management. ...................... 6.4. Managing Trophic Interactions and Ecosystem Processes. 6.5. Marine Protected Areas ....................................... 6.6. Conclusions ................................................ 7. Summary ..................................................... Acknowledgements .............................................. References ....................................................

............

303 303 303 305 306 308 310 311 313 314

ABSTRACT We review the effects of fishing on benthic fauna, habitat, diversity, community structure and trophic interactions in tropical, temperate and polar marine environments and consider whether it is possible to predict or manage fishing-induced changes in marine ecosystems. Such considerations are timely given the disillusionment with some fishery management strategies and that policy makers need a scientific basis for deciding whether they should respond to social, economic and political demands for instituting or preventing ecosystem-based management. Fishing has significant direct and indirect effects on habitat, and on the diversity, structure and productivity of benthic communities. These effects are most readily identified and last longest in those areas that experience infrequent natural disturbance. The initiation of fishing in an unfished system leads to dramatic changes in fish community structure. As fishing intensity increases the additional effects are more difficult to detect. Fishing has accelerated and magnified natural declines in the abundance of many forage fishes and this has lead to reduced reproductive success and abundance in birds and marine mammals. However, such donor-controlled dynamics are less apparent in food webs where fishes are the top predators since their feeding strategies are rather more plastic than those of most birds and mammals. Fishers tend to target species in sequence as a fishery develops and this leads to changes in the composition of the fished communities with time. The dramatic and apparently compensatory shifts in the biomass of different species in many fished ecosystems have often been driven by environmental change rather than the indirect effects of fishing. Indeed, in most pelagic systems, species replacements would have occurred, albeit less rapidly, in the absence of fishing pressure. In those cases when predator or prey species fill a key role, fishing can have dramatic indirect effects on community structure. Thus fishing has shifted some coral reef ecosystems to alternate stable states because there is tight predator-prey coupling between

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

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invertebrate feeding fishes and sea urchins. Fishing has reduced, and locally extirpated, populations of predatory fishes. These reductions do not have a consistent effect on the abundance and diversity of their prey: environmental processes control prey populations in some systems, whereas top-down processes are more important in others. By-catch which is discarded during fishing activities may sustain populations of scavenging species, particularly seabirds. We conclude by identifying the circumstances in which new research is needed to guide managers and stress the importance of unfished control sites for studies of fishing effects. We discuss the advantages and disadvantages of closed area management (marine reserves) and the conditions under which such management is likely to provide benefits for the fishery or ecosystem.

1. GENERAL INTRODUCTION

Fishing is the most widespread human exploitative activity in the marine environment. Pauly and Christensen (1995) estimated that over 20% of primary production is required to sustain fisheries in many intensively fished coastal ecosystems (Figure 1). Previous estimates of the primary production required were much lower and led Vitousek et al. (1986) to conclude that fishing had few fundamental effects on the structure or function of marine ecosystems apart from those on fished species. These views were widely accepted at the time since they were in accordance with the overriding philosophy of many fisheries scientists who based their assessment and management actions upon the short-term dynamics of target fish populations (Frank and Leggett, 1994; Smith, 1994). However, studies such as those of Pauly and Christensen (1995), coupled with empirical evidence for shifts in marine ecosystems, imply that the actions of fishers may have important effects on ecosystem function (Sherman and Alexander, 1986). As a result the emphasis of marine fisheries research is beginning to shift from population- to ecosystem-based concerns (Langton et al., 1996; Auster et al. 1996), and this has been reflected in a number of recent reviews describing the effects of fishing on ecosystem structure and processes (Munro et al., 1987; McClanahan and Muthiga, 1988; Hutchings, 1990; Russ, 1991; Jones, 1992; Gislason, 1994; Hughes, 1994; Matishov and Pavlova, 1994; Anon., 1995b; Dayton et al., 1995; McClanahan and Obura, 1995; Roberts, 1995; Jennings and Lock, 1996; Jennings and Polunin, 1996b; Birkeland, 1997). As early as the fourteenth century there were concerns about the effects of fishing on the marine environment and these effects were discussed in detail by a number of Government Commissions in the United Kingdom

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Open ocean Coastal/ reef systems

S. JENNINGS AND M.J. KAISER

w

Tropical shelves Upwellings Non-tropical shelves

0 10 20 30 40 50 60 70 80 90 Primary production required (%) Figure 1 Global estimates of the primary production (&95% CL) required to sustain fisheries in five marine ecosystems. Estimates calculated from the size and trophic composition of the global fish catch and by-catch from 1989 to 1991, assuming a 10% transfer efficiency between trophic levels. (Data from Pauly and Christensen, 1995.)

(Anon., 1885). However, the scientific basis for the management of fisheries was founded in the study of exploited fish populations and it was not surprising that these were the primary unit of concern since species or intraspecific stocks were the targets of fishers, the categories favoured by buyers or consumers, and the groupings in which fishing effects were most readily recognized. Early studies of fish population dynamics (Petersen, 1894; Garstang, 1900; Hjort, 1914; Hjort et al., 1933) were paralleled by the wide-ranging studies of ecologists and mathematicians such as Malthus (1798), Lotka (1925) and Volterra (1926) who discussed the impacts of resource limitation on population growth and mathematical approaches for describing population fluctuations. In subsequent years, mathematical descriptions of fish population dynamics and the effects of exploitation were developed by Russell (1931, 1939), Graham (1935), Schaefer (1954), Beverton and Holt (1957) and Ricker (1958) whose models had considerable influence on wider scientific thinking at that time. As applied fisheries science continued to develop, there was relatively little concern about the functional role of fishes within the marine ecosystem and the indirect effects of changes in their abundance or diversity. Indeed, the study of food webs

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

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(Elton, 1927) had considerable influence on the development of terrestrial ecology, but limited impact on the development of fishery science. Throughout the early twentieth century the discussion between fishery scientists and those working on the dynamics of other populations continued, but in subsequent years it was notable that fisheries science was increasingly viewed as a field of study in its own right. With few exceptions (May, 1984), fisheries scientists did not respond to many major theoretical advances in the description of ecosystem structure and function (May, 1973; Pimm, 1982, 1991). This was strange since influential studies of the processes driving fish production had been conducted by scientists who worked in laboratories charged with fisheries assessment and whose interests freely transcended the boundaries between fish population biology, marine ecology and oceanographic science (Hjort, 19 14; Ryther, 1969; Gulland, 1970; Steele, 1974; Cushing, 1975, 1982; Andersen and Ursin, 1977, 1978; Steele and Henderson, 1984; Sharp, 1988; Southward et al. 1988). Their ideas, however, were rarely translated into practical management advice, since social and political pressures focused attention on the fish stock as the prime management unit and the study of management measures such as changes in mesh sizes or fishing mortality had become a central component of population-based research (Smith, 1994). The study of fish population dynamics continues to provide the basis for most present-day management decisions. In those countries with the resources to implement procedures such as virtual population analysis and multispecies virtual population analysis (Pope, 1979; Sparre, 1991), the short-term predictions of stock structure and potential yield are remarkably accurate. Indeed, many practitioners continue to argue that the methods of enforcing management advice are of vastly more concern than the scientific details of the stock assessment models which they prefer (Anon., 1997). From the fisheries scientists’ viewpoint, it is the lack of clear necessity that kept the ecosystem’s perspective from advancing in a field whose pragmatic concern is the mechanics of short-term fishery management. The existing concerns of fisheries scientists in relation to human activities have largely focused on the dramatic collapse of a few stocks such as the Peruvian anchovy Engraulis ringens Jenyn and Atlantic cod Gadus morrhua L. (Idyll, 1973; Myers et al., 1996). Wider concerns that the rate of increase in the global fish catch is declining as demand is increasing (Anon., 1997), that the proportion of fish caught in many fisheries leaves little latitude for recruitment failure (Myers et al., 1995; Cook et al., 1997) and that unwanted by-catch often forms a relatively large proportion of the total catch (Alverson et al., 1994; Hall, 1996) also focus upon fish populations. Conversely, the concerns of terrestrial ecologists are rapidly moving from species to focus on habitats and ecosystems in recognition of the need to

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maintain ecosystem integrity and function rather than simply preserving entities (Gaston, 1996). Many of the key scientists in this field are now assessing the impact of human activities on the structure and function of ecosystems and the ways in which production processes and ecosystem stability are affected by reductions in species diversity (Ehrlich and Wilson, 1991; Walker, 1992; Ehrlich and Daily, 1993; Huston, 1994; Lawton, 1994; Naeem et al., 1994, 1995; Tilman and Downing, 1994; Anon., 1995f; Gadgil, 1996; Johnson et al., 1996; Kunin and Lawton, 1996; Vane-Wright, 1996). The possibility that fisheries have major effects at the ecosystem level and that the ecosystem should be considered as an assessment and management unit have been expressed by some marine ecologists (Sherman and Alexander, 1986; Sherman et al., 1991,1993).Fishing has a number of direct effects on marine ecosystems because it is responsible for increasing the mortality of target and by-catch species and disturbing marine habitats. The direct effects of fishing have many indirect implications for other species. Thus fishers may remove some of the prey that piscivorous fishes, birds and mammals would otherwise consume, or may remove predators that would otherwise control prey populations. Moreover, reductions in the density of some species may affect competitive interactions and result in the proliferation of non-target species. The activities of fishers also provide food for scavenging species since fishes, benthic organisms and other unwanted by-catch are often discarded and because a range of species are killed, but not retained, by towed gears. Our aim is to compile evidence for the effects of fishing on ecosystem structure or function and to determine whether there is a scientific basis for the prediction or management of changes in marine ecosystems. We believe that ecological questions relevant to the marine environment must be studied on many spatial and temporal scales and suggest that an understanding of fishing effects requires the integration of population and ecosystem-centred research. This review describes the effects of fishing on benthic fauna and habitat, community structure and trophic interactions. The divisions serve to structure the review, but they are primarily artificial because the effects of fishing are not mutually exclusive. For example, the intensive fishing of invertebrate-feedingfishes on Kenyan coral reefs has led to reductions in fish species diversity and reduced predation on sea urchins. In the longer term, sea urchin biomass has increased leading to a reduction in algal biomass and bioerosion of the reef matrix with corresponding reductions in the numerical abundance and diversity of fishes which can use the reef habitat (McClanahan and Muthiga, 1988; McClanahan, 1990, 1992, 1994a, 1995a,b). Our review is wide-ranging but selective. In particular, we have largely disregarded the direct effects of fishing on target and by-catch species. The effects of fishing on target species are described in a number of texts

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and form the basis of traditional fisheries science (Beverton and Holt, 1957; Beverton, 1963; Cushing, 1968; Nikolskii, 1969; Gulland, 1977; Munro and Williams, 1985; Hilborn and Walters, 1992; Appeldoorn, 1996) and there are a number of reviews that have described the type and magnitude of by-catch in many fisheries (Beddington et al., 1985; Lo and Smith, 1986; Poiner et al., 1990; Andrew and Pepperell, 1992; Francis et al., 1992; Alverson et al., 1994; Jefferson and Curry, 1994; Anon., 1995b, 1996c; Hall, 1996; Simmonds and Hutchinson, 1996). In examining the indirect effects of fishing on the ecosystem, it is important to recognize that some by-catch species have been affected dramatically by fishing. For example, Smith (1983) estimated that the population sizes of three dolphin species caught by tuna boats in the eastern tropical Pacific were reduced to 20%, 35-50% and 58-72% of pre-exploitation levels by 1979. By-catches of marine mammals, birds and sea turtles have become a dominant political issue in the management of many fisheries (Hall, 1996) and even when rates of mortality have negligible effects on the populations, they may be politically significant. To some extent the value of biological significance is arbitrary (Hall, 1996). Thus the proposal that a by-catch mortality equivalent to 0.5% of cetacean population size is acceptable (Perrin et al., 1994) may satisfy those who hope that the removal of dolphins will not affect ecosystem function but will not placate those concerned for the welfare of individual dolphins. Such issues are outside the scope of this review. Our consideration of by-catch is limited to the indirect effects of discarding on scavenger communities. The portion of the total catch which is discarded dead or dying, either because it is illegal to land or because there is little or no economic gain associated with sorting or retaining it, constitutes approximately 27% of global fish catches (Alverson et al., 1994). Thus fishing activities subsidize marine foodwebs with carrion that would be unavailable under natural conditions, which may have profound effects on scavenging species (Britton and Morton, 1994). Moreover, in some fisheries, the ratio of by-catch to landings is so high that the removal of target species may not constitute the main impact of the fishery (Figure 2). Having reviewed the evidence for the ecosystem effects of fishing we consider whether description can be used as a basis for prediction and whether an understanding of the ecosystem effects of fishing could be translated into effective management. Such considerations are timely given the increasing public and scientific disillusionment with some existing fishery management strategies and that policy makers need a scientific basis for deciding whether they should respond to social, economic and political demands for instituting or preventing ecosystem-based management.

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7 1 Tunas, bonitos Miscellaneous

I

Squid, cuttlefish 1 &octopuses

Mackerels, snooks 8 cutlassfishes

0

2

4

6

Discards (t x

8 1 0 0

lo6)

5 1 0 1 5 2 0 2 5 0 1 2 3 4 5 6

Landings (t x lo6)

Discards: landings

Figure 2 Estimates of global discards and landings, and the ratio between discarded weight and landed weight, in fisheries targeting 11 groups of marine species. (Data from Alverson et al., 1994.)

2. BENTHIC FAUNA AND HABITAT 2.1. Introduction

Fishing activities lead to changes in the structure of marine habitats and can determine the diversity, composition, biomass and productivity of the associated biota. The effects of fishing on habitats are often large-scale ramifications of the cumulative effects on many individual plants and invertebrates since habitats such as kelp forests, coral reefs or bryozoan beds are formed by living organisms. Many fishing gears have direct effects on habitat structure, but indirect effects occur when fishing initiates shifts in the relationships between those organisms responsible for habitat development and degradation. The direct effects of fishing vary according to the gears used and the habitats fished, but they usually include the scraping, scouring and resuspension of the substratum. The magnitude of changes which can be

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

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attributed to fishing often depend upon the nature of the physical environment in which a given habitat is found. Thus the effects of fishing on communities of short-lived burrowing worms that temporarily inhabit mobile sediments in shallow shelf seas will be harder to detect than the effects on coral communities that structure equatorial coral reefs. The indirect effects of fishing on non-target fishes and invertebrates may lead to changes in community structure and habitat type. For example, there is increasingly good evidence to show that the indirect effects of fishing have caused some reef communities to shift from coral to algal or urchin-dominated phases. In Section 2.2 we describe fishing methods that impact the marine ecosystem directly and their effects on habitat structure, benthic communities and nontarget species. In Section 2.3 we discuss the indirect effects of fisheries on habitat structure and production processes, and the ramifications of such changes for fish and invertebrate communities. Finally Section 2.4 considers the relative roles of natural and fishing disturbance in the marine environment.

2.2. Direct Effects of Fishing Gears

Fishing techniques that affect benthic fauna and habitats can be grouped into two categories: active and passive. Active fishing methods usually involve towing trawls or dredges on continental shelves. However, artisanal fishers operating on tropical coasts also use a range of active techniques such as drive netting, spearing and fishing with chemicals or explosives. Passive fishing techniques include the use of pots or traps, baited hooks on set lines, gill nets and drift nets. Actively or passively fished surface, midwater and bottom fishing gears can have direct effects on non-target animals such as birds, marine mammals and reptiles and fishes which are taken as by-catch. In addition, the actions of fishers and their gears extensively modify seabed habitats and their associated benthic communities. In this section we describe the main fishing gears and their effects on benthic fauna and habitats. 2.2.1. Active Fishing Techniques 2.2.1.1. Trawls and dredges The majority of mobile demersal fishing gears can be described as trawls or dredges. Both types of gear are used to capture species that live or feed in benthic habitats, and thus they have been designed to maximize their contact with the seabed. Fishing techniques and equipment have been fine-tuned to exploit the behaviour and habitat preferences of target species and to achieve the maximum catch-per-

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unit-effort. Presumably, fishers use the most effective techniques currently available. As commercial stocks have diminished, so fishing gears have been modified to maintain yield. The increasing power of fishing vessels has permitted the use of larger and heavier trawls and dredges. Towing larger gears incurs higher fuel costs which have to be offset by higher catches. However, these financial considerations take no account of the concomitant increase in environmental damage to non-target benthic communities. Trawls generally fall into two categories, otter and beam trawls. Otter trawls derive their name from the two rectangular otter boards or doors, attached to the towing warps, which act as paravanes to maintain the lateral opening at the mouth of the net. The boards can weigh several tonnes in air and are towed at an oblique angle across the seabed (Jones, 1992). When fished over fine muddy sediments the boards are sometimes fitted with metal shoes up to 30 cm wide which are designed to prevent the boards digging too far into the sediment (M.J. Kaiser, pers. obs.). Nevertheless, Krost et al. (1990) estimated that otter boards penetrated soft mud to a depth of 15cm in the Baltic Sea. In the simplest otter trawls, the ground gear comprises a foot rope protected by sacrificial twine or rubber bobbins, which will be less intrusive than the otter doors. However, when used to catch flatfishes, varying numbers of tickler chains are attached between the otter boards (Harden Jones and Scholes, 1974; Sainsbury, 1987). Rockhopper gear represents the most extreme type of ground gear fitted to otter trawls. As its name suggests, this gear is used over rocky substrata. The groundrope is fitted with large rubber discs (>50cm diameter) and metal bobbins, which each weigh > 10kg. The discs are held in position by a wire which runs the length of the ground rope and is threaded through their rear half. When the discs foul, they partially rotate against the tension imposed by the wire and then “spring” clear, allowing the gear to hop over solid obstructions. Otter trawls are used at depths of up to 1500m, which is far in excess of any other towed fishing net (Jones, 1992; Clark, 1996). To date, the incidental effects of these deep water fisheries are unknown although some unpublished data on bycatches are now available. These fisheries target pelagic species such as orange roughy, Hoplostethus atlanticus Collett, which aggregate in association with reef structures over which the gear is towed. Physical contact of these trawls with the reef substratum is likely to damage the epifaunal community. Beam trawls comprise a rigid beam held off the seabed by two beam shoes. The net headline is attached to the beam and the footrope is attached to the beam shoes; thus the mouth of the net is fixed in an open position. Beam trawls are towed at speeds of up to 7 knots (Kaiser et al., 1996b). Decreasing fish stocks have necessitated gear modifications such as increasing beam width and the addition of more tickler chains or the use of chain mats and flip-up ropes. Consequently, beam trawls have increased in weight

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

21 1

from a mean of 3.5 t in the 1960s (Cole, 1971) up to 10 t in the early 1980s (Beek et al., 1990). Beam trawlers specifically target benthic species such as sole, Solea solea (L.), plaice, Pleuronectes platessa L. and shrimp, Crangon crangon L. which are normally buried in, or rest on, surface sediments. The number of tickler chains fitted to the gear depends on the sediment characteristics of the fishing grounds, and 17 to 25 may be used on some of the largest trawls (Polet et al., 1994). The tickler chains are specifically designed to penetrate the upper layers of the sediment, disturbing those target species that are buried in the sediment so that they swim up into the path of the trawl. Successive chains dig deeper as the leading chains “fluidize” the sediment. The heaviest trawls are used over rough grounds and are fitted with a chain matrix (“stone mat” gear) which prevents large rocks entering the net and causing damage to the gear and catch. Dredges can be categorized as hydraulic or mechanical. Hydraulic dredges lift the sediment, non-target and target species, whereas mechanical dredges physically dig target species out of the sediment. Hydraulic dredges use jets of water or air to create a venturi effect, which lifts the dredgings onto a boat for further processing on fixed or mechanical riddles (Meyer et al. 1981). Some of the largest commercial hydraulic dredgers harvest lugworms, Arenicola marina L., in the Dutch Wadden Sea. These leave furrows l m wide and 40cm deep (Beukema, 1995). Similar devices are used to harvest cockles, Cerastoderrna edule (L.) and Manila clams, Tapes philippinarum Adams and Reeve, at mid to high tide on sandflats in northern Europe (Hall and Harding, 1997; Spencer et al. 1997). Suction dredges are also used on a much smaller scale by divers to remove razor clams, Ensis siliqua (L.); although the area disturbed is relatively small, pits are often excavated to depths of 60cm (Hall et al. 1990). Mechanical dredges differ from trawls because they are designed to dig further into the substratum than beam trawls. Most dredges are used to target epi- or infaunal bivalves such as scallops, Pecten maximus (L.), clams, Mercenaria mercenaria (L.) and razor clams. Most dredge designs incorporate similar features such as a heavy duty bag or net attached to a rigid metal frame. Tooth bars or cutting blades of various designs are usually fitted to the frame. For example, the Newhaven dredge is fitted with a tooth bar bearing teeth approximately 11 cm long. The tooth bars are designed to disturb scallops which lie in shallow depressions in the seabed. Since scallop dredges tend to be used over rough ground, steel ring bellies are usually fitted to the net bag. Large scallop boats fish between 36 and 40 dredges simultaneously and these are attached to beams fitted with rollers that reduce drag. The total width and weight of a set of scallop gear is comparable with some of the larger beam trawls (Kaiser et al., 1996b). Deep burrowing species such as razor clams are caught in dredges fitted with teeth up to 30cm long (Gaspar et al., 1994). The drag created by

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such a deep-digging dredge prevents small inshore boats fishing more than two at a time. Dredges are rarely towed at speeds in excess of 2.5 knots (4.6 km h-’) since the gear is less efficient at higher speeds (Caddy, 1968, 1973; Dare et al., 1993). Consequently, scallop dredges disturb smaller areas of seabed per unit time than beam trawls (Anon., 1995b; Kaiser et al., 1996b). Trawls and dredges have marked impacts on the substratum. Physical disturbance of the substratum results from direct contact with the fishing gear and the turbulent resuspension of surface sediments. The magnitude of the impact is determined by the speed of towing, physical dimensions and weight of the gear, type of substratum and strength of currents or tides in the area fished. The effects may persist for a few hours in shallow waters with strong tides or for decades in the deep sea. In many shelf seas fishing intensity is very high and most fishable grounds will be impacted at intervals of less than one year. On Georges Bank, Caddy (1973) reported that 3% of the seabed in his study area was covered with trawl marks, but the persistence of the marks was unknown. More recently, Churchill (1989) estimated that 18% of a 259km2 area in the Middle Atlantic Bight was trawled in a 6-d period of intense fishing activity and Twichell et al. (1981) recorded up to 20 trawl tracks per 100m’ in the New York Bight, at a depth of loom, where current action was weak. Similarly, Krost et al. (1990) found that trawl tracks occupied 19% of their muddy and relatively deep study area. Churchill (1989) calculated that the effective area trawled on an annual basis in a number of 30’ latitude by 30’ longitude areas in the Middle Atlantic Bight was up to three times their actual area and Welleman (1989) and Rijnsdorp et al. (1991b, 1997) reported that some intensively fished regions of the southern North Sea were swept by trawls several times each year (Plate 1). If observations of trawl marks are to be used to provide an index of fishing intensity then some knowledge of their persistence, as determined from experimental studies, is required. High-resolution video images of sediment surfaces before and after otter trawling indicate that trawling reduces the overall surface roughness of the seabed (Schwinghamer et al., 1996) although trawl doors may leave depressions. Ripples, detrital aggregations and surface traces of bioturbation are smoothed over by the mechanical action of the trawl and the suspension and subsequent redeposition of the surface sediment. Acoustic data collected on trawled experimental sites on the eastern Grand Banks, Canada, showed the effects of trawling could be detected to a depth of at least 4.5cm within the sediment (hard packed sand), and there was a general, although uneven, reduction in the complexity of the internal sediment structure (Schwinghamer et al., 1996). The physical disturbance of sediment can result in a loss of biological organization and reduce species richness (Hall, 1994).

Plate 1 Mean beam trawling intensity by Dutch beam trawlers with engines >300hp from 1 April 1993 to 31 March 1996 in areas of 3 x 3 nautical miles. Beam trawl intensity within each square is expressed as the mean number of times that each square metre in that area is trawled each year. Data collected from loggers which recorded the location of vessels at 6-min intervals. (Further details in Rijnsdorp et al., 1997. Reproduced with permission from Rijnsdorp et al., 1997.)

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It is clear that all mobile bottom gears scrape the surface of, or dig into, the seabed to varying degrees. Hence, it is not surprising that non-target fishes and benthic invertebrate species form a large proportion of the catch in some fisheries (Andrew and Pepperell, 1992; Robin, 1992; de Groot and Lindeboom, 1994; Anon., 1996b; Raloff, 1996). While gear modifications such as the addition of extra tickler chains increase the catch of target species, there is an unavoidable and concomitant increase in the catch of non-target species (Cruetzberg et al., 1987; Kaiser et al., 1994). While nets have been refined to reduce by-catch of non-target and undersized commercial species (e.g. Briggs, 1992), few attempts have been made to reduce bycatch or the physical effects of fishing gears on invertebrate benthic species. For the purposes of this review we define infauna as those animals living entirely within the sediment, whereas epifauna are defined as those animals living on, protruding from, anchored in, or attached to, the sediment. 2.2.1.2. Effects of trawls and dredges on infauna By-catches of non-target infaunal species indicate the extent to which benthic communities are perturbed by a particular gear. For example, the occurrence of the infaunal bivalve, Arctica islandica (L.), and the heart urchin, Echinocardium cordatum (Pennant), in a 12-m beam trawl catch suggested that the tickler chains had penetrated hard sandy substrata to a depth of at least 6cm (Bergman and Hup, 1992); although Steve Lockwood (CEFAS Conwy, pers. comm.) has reported catches of E. cordatum from trawls which penetrate less than 1 cm. The position of small urchins within the sediment column, and not their size, makes them vulnerable: smaller size-classes of heart urchins are found closer to the sediment surface and are most vulnerable to physical damage. Bergman and Hup (1992) emphasized the importance of considering the vulnerability of animals at different stages of their life history. In their study, it was estimated that 90% of the A . islandica in the catch had broken shells; although they did not provide information on the number that were damaged but remained in the sediment and were not able to sample this species adequately in order to determine changes in density. However, the prevalence of A. islandica in the stomach contents of Atlantic cod at times of intensive otter trawling in Kiel Bay indicates that large numbers of these bivalves are damaged by trawling (Arntz and Weber, 1970). Rumohr and Krost (1991) found larger numbers of damaged A . islandica in a dredge towed directly behind an otter board than in the centre of the net. Furthermore, damaged A . islandica have been observed by divers while surveying areas of the seabed disturbed by beam trawls (Kaiser and Spencer, 1996a). Although A . islandica are vulnerable to damage by trawls, those that are slightly damaged can repair cracks in their shell matrix. As a consequence of physical damage, sand grains become lodged between the mantle and the

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growing edge of the shell, eventually becoming incorporated into the shell matrix (Gaspar et al., 1994; Witbaard and Klein, 1994). Witbaard and Klein (1994) studied annual growth rings in the shells of A. islandica, and were able to back-calculate the years in which they had been damaged by noting the occurrence of sand grains in the shell matrix. The incidence of shell damage correlated with increasing beam trawling activity between 1972 and 1991 at a study site in the southern North Sea (Figure 3; Witbaard and Klein, 1994). Witbaard and Klein (1994) concluded that their study site had been disturbed by demersal fishing gear at least once per year during this period.

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that year. Beam trawl effort is expressed on a relative scale based on days fishing standardized for beam size and vessel power. (Data from Witbaard and Klein, 1994; Heessen and Daan, 1996.)

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While it has been relatively simple to detect the changes in abundance of large macroinfauna which result from fishing disturbance, the smaller fauna (-=10 mm) show conflicting responses. Furthermore, a recent study suggests that fauna below a certain body size or mass are resuspended by a pressure wave in advance of otter trawl doors, and are redistributed to the sides of the gear (Gilkinson et al., 1998). Bergman and Hup (1992) found both decreases and increases in the abundance of smaller invertebrates after fishing an area of seabed with a beam trawl. A species-by-species analysis of responses to fishing gear disturbance (Bergman and Hup, 1992; Eleftheriou and Robertson, 1992) may have been less effective than the multivariate approaches adopted in more recent studies (Thrush et al., 1995; Currie and Parry, 1996; Kaiser and Spencer, 1996b). Furthermore, studies in the southern North Sea have been hampered by the inescapable fact that this area has already been disturbed by fishing for at least 100 years (Figure 4). Kaiser and Spencer (1996b) studied the effects of beam trawl disturbance at a site 2 7 4 0 m deep in the Irish Sea. Their experimental site encompassed two distinct habitats: stable sediments composed of coarse sand, gravel and shell debris, which supported a rich epifaunal filter-feeding community of soft corals and hydroids, and mobile sediments characterized by ribbons of megaripples with few sessile epifaunal species. Despite a robust experimental design with paired treatment and control areas, the effects of beam trawl disturbance were undetectable in the mobile sediments. This is not surprising given the levels of natural disturbance experienced in megaripple habitats (Shepherd, 1983). Similarly, de Wolf and Mulder (1985) reported that they could not provide accurate estimates of the abundance of benthic species in megaripple habitats because of the spatial variability within this type of habitat. In addition, animals living in the troughs of megaripples were less likely to be disturbed by fishing since the gears rode over the crest of each sand wave. Brylinsky et al. (1994) were also unable to detect any adverse effects of otter trawling over intertidal mud flats that are regularly exposed to large-scale disturbances such as ice-scour. Conversely, in stable sediments the effects of fishing are more noticeable. Kaiser and Spencer (1996b) found that the number of species and individuals in the stable sediment community was reduced by two- and threefold respectively. Their analysis also revealed that less common species were most severely depleted by beam trawling. In a similar study, Thrush et al. (1995) studied the effects of scallop dredging on a coarse sand community at a depth of 27m. They were able to detect changes in the populations of individuals and compositional differences in the community that lasted for at least 3 months after initial disturbance. Thrush et al. (1995) emphasized that their study was conservative as they were unable to simulate the effects of an entire fishing fleet, implying that at larger scales of disturbance recolonization may take longer. Infauna that live within a few centimetres of the sediment surface at

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Figure 4 Expansion of the area trawled by the English North Sea fleet before 1900. (Redrawn with modifications from Smith, 1994.)

depths < 30 m tend to be small opportunistic species (e.g. spionid and capitellid polychaetes and amphipods) that quickly recolonize areas after disturbance (Dauer, 1984; Levin, 1984). As a consequence, the effects of trawling on this component of the infaunal community are unlikely to last more than 6-12 months. However, a recent study (Posey et al., 1996) suggested that deeper burrowing fauna were not affected by severe episodic storms. The study site was at a depth of 13m, and samples were collected to a depth of 15cm. “Deeper burrowing” was not defined, but it implies

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fauna living at a depth of 7-1 5 cm which is well within the depths disturbed by trawls and dredges (Krost et al., 1990; Bergman and Hup, 1992). If these fauna are less well adapted to periodic natural disturbances, they may be more severely affected by trawling activity. In general, it seems reasonable to predict that the effects of physical disturbance will be short-lived in communities adapted to frequent natural perturbations in contrast to those communities found in habitats exposed to fewer disturbances. An extreme example of the former situation is Hall and Harding’s (1997) study of the effects of mechanical and suction dredging and the scale of disturbance on an intertidal benthic community in the Solway Firth, Scotland. The immediate effects of cockle harvesting were obvious with a drastic reduction in the abundance of individuals; however, the community in disturbed areas was comparable to that in control undisturbed areas after only 8 weeks. This rapid recolonization was attributed to the immigration of adults against a background of seasonal recruitment (Hall and Harding, 1997). This study contrasts with an investigation of the effects of suction dredging for manganese nodules on the abyssal plain of the Pacific Ocean (Thiel and Schriever, 1990; Thiel, 1992). Trenches created by the suction dredge head persisted for at least 2 years in this stable environment. While the persistence of disturbance effects may be approximately correlated to the level of natural disturbance experienced in a particular habitat, there are some exceptions. This is well illustrated in a recent study in which the effects of the scale of sediment defaunation were studied on an intertidal sandflat in New Zealand (Thrush et al., 1996). In contrast to Hall and Harding’s (1997) findings, recolonization rate was reduced at larger scales of disturbance. The main difference between these two studies was the presence, in the New Zealand study, of dense mats of tube-building spionid worms which stabilized the sandflat sediments. Removal of these animals destabilized the sediment and exacerbated the effects of disturbance. Furthermore, while the changes associated with disturbance are relatively short-lived for the majority of small species, longer-lived organisms recolonize more slowly. For example, Beukema (1995) reported that the biomass of gaper clams, Mya arenariu L., took 2 years to recover after commercial lugworm dredging in areas of the Wadden Sea, whereas small polychaetes and bivalves had recolonized the dredged areas within 12 months. Longlived epifaunal organisms frequently have a structural role within benthic communities, providing a microhabitat for a large number of species (see Section 2.2.1.3). Calcareous algae of the genus Lithothamnion are amongst the oldest living marine plants in Europe and provide a substratum that can take hundreds of years to accumulate (Potin et al., 1990). The branching structure of the thalli provide a unique habitat for a diverse community of animals including commercial species such as scallops. Not surprisingly, scallop dredging in this habitat causes destruction of the interstices between

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the thalli and causes long-term changes to the composition of the associated benthic fauna (Hall-Spencer, 1995). To date, most studies have been centred on the hypothesis that ". . . trawling/dredging has the potential to bring about long-term changes in community structure. . .", and have measured changes observed after an experimental fishing disturbance (Graham, 1955; Bergman and Hup, 1992; Eleftheriou and Robertson, 1992; Thrush et al., 1995; Currie and Parry, 1996; Kaiser and Spencer, 1996b; Pitcher et al. 1997). An alternative approach is to examine benthic community changes after closing areas that have been subjected to fishing disturbance for many years. Hill et al. (unpublished data) examined changes after the closure of an area within a scallop ground that had been heavily fished for over 50 years (Brand el al., 1991). After several years of closure, they found that the variation between infaunal samples within the closed area was greater than in the adjacent dredged areas. This suggests that intensive dredging leads to a more homogeneous environment, in a manner analogous to a tractor ploughing a meadow. Van Dolah et al. (1991) studied changes in infaunal communities over a period of 5 months within areas closed to fishing and in adjacent areas fished by shrimp trawlers. They concluded that seasonal reductions in the abundance and number of species sampled had a much greater effect than fishing disturbance. However, in a power analysis of their sampling strategy, only changes in the abundance of individuals and the number of species were considered. This assumes that the response of the infauna to trawling disturbance was unidirectional, whereas a consideration of changes in partial dominance might have been more sensitive to subtle changes in the fauna. While their results should be interpreted with caution it remains plausible that light shrimp trawls do not cause significant disturbance to communities in poorly sorted sediments in shallow water (Van Dolah et al., 1991). In addition, Van Dolah et al. (1991) sampled fauna from fished areas located between shoals and their study indicated that the local sediments were probably mobile and inhabited by fauna adapted to frequent disturbance (Kaiser and Spencer, 1996b). Thus far, we have only considered the effects of fishing on infaunal communities living in coarse substrata. Most animals are found within the top lOcm of these sediment habitats. However, in soft mud communities a large proportion of the fauna live in burrows up to 2m deep (Atkinson and Nash, 1990). Consequently few of these deep-burrowing fauna, such as thalassinid shrimps, are likely to be affected by passing trawls. Although upper burrow structures are collapsed by passing fishing gear, they are rapidly reconstructed (R.J.A. Atkinson pers. corn.). However, the energetic costs of repeated burrow reconstruction may have long-term implications for the survivorship of individuals. In addition, die1 variation in

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behaviour may periodically increase the vulnerability of some species to fishing activities. For example, the burrowing shrimp Jaxea nocturna Nardo moves to the entrance of its burrow to feed at night (Nickel1 and Atkinson, 1995). These animals, along with other bioturbators, have an important role in maintaining the structure and oxygenation of muddy sediment habitats (Reise, 1981; Rowden and Jones, 1993; Fenchel and Finlay, 1995; Fenchel, 1996). Consequently, any adverse effects of fishing on these organisms would presumably lead to changes in habitat complexity and community structure. 2.2.1.3. Effects of trawls and dredges on epifauna Intuitively, sessile epibenthic species are most likely to be vulnerable to the passage of bottom gears. Accordingly, observations of the changes in epifaunal communities in heavily fished areas have provided some of the first indications of the potential long-term effects of fishing on benthic communities. The disappearance of reefs of the tube-building worm, Sabellaria spinulosa Leukart, and their replacement by small polychaete communities, indicated that dredging activity had caused measurable changes in the Wadden Sea benthic community (Riesen and Riese, 1982). Similarly, Sainsbury (1987) reported a measurable decrease in the biomass of the sponge by-catch in the Australian northwest shelf pair-trawl fishery from 1967 to 1985. Loss of the sponge community and associated fauna such as alcyonarians and gorgonians led to a reduction in the catches of emperors, Lethrinus spp., and snappers, Lutjanus spp. which sheltered and fed among the emergent fauna (Sainsbury, 1988). Langton and Robinson (1990) observed a c. 26% reduction in the mean density of the sabellid worm, Myxicola infundibulum (Renier) and the cerianthid anemone, Cerianthus borealis Verrill after one season of intense commercial scallop dredging on the Fippenies Ledges, Gulf of Maine. In addition, the significant negative association between these species became random after intensive fishing (Langton and Robinson, 1990). These authors hypothesized that cerianthid predation of scallop and sabellid worm larvae was an important factor controlling their spatial distribution. Thus the species association was broken down by dredging disturbance. Using a combination of fishing effort data and direct observations from side-scan sonar surveys, Collie et al. (1997) were able to identify comparable substrata that experienced different intensities of scallop dredging on the Georges Bank, northwest Atlantic. Areas that were less frequently fished were characterized by abundant bryozoans, hydroids and worm tubes which increased the three-dimensional complexity of the habitat (Figure 5). Furthermore, examination of evenness within the community suggested dominance by these structural organisms, which indicated that this environment was relatively undisturbed. In contrast, the more intensively dredged areas had lower species diversity, lower

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Figure 5 Photographs of undredged (a) and dredged (b) gravel habitats on the eastern Georges Bank, United States. Epifaunal species are almost absent at the dredged site. (Photographed by Dann Blackwood and Page Valentine, US Geological Survey. Further details in Collie ef al., 1997.)

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biomass of fauna, and were dominated by hard-shelled bivalves (e.g. Astarte spp.), echinoderms and scavenging decapods. The higher diversity indices observed at the less intensively dredged sites were attributable to the large number of organisms, such as polychaetes, shrimp, brittle stars, mussels and small fishes, that were associated with the biogenic fauna (Collie et al., 1997). Many of these associated species were also important prey for commercially exploited fishes such as cod (Bowman and Michaels, 1984). Similarly, Auster et al. (1996) reported a reduction in habitat complexity as a result of fishing (trawling and scallop dredging) activity at three sites in the Gulf of Maine. Video observations made with a remote-operated vehicle (ROV) revealed cleared swaths in the epifaunal cover on the border of the Swans Island conservation area which has been closed to fishing with mobile gears since 1983. As in other studies (Bradstock and Gordon, 1983; Sainsbury, 1987; Collie et al., 1997), hydroids, bryozoans, sponges and serpulid worm matrices were greatly reduced in the fished areas. In addition, there was a reduction in the habitat features produced by some of the target species, e.g. pits created by scallops and crabs (Auster et al., 1996). The Jeffreys Bank site was surveyed by submersible in 1987 and again in 1993. Boulders, 2 m wide, were a prominent feature of the site where towed fishing gear had been excluded until 1987. However, when the site was resurveyed, the percentage cover of sponges was greatly reduced, the thin mud veneer that previously covered the underlying gravel was no longer evident, and boulders appeared to have been moved across the seabed. The Stellwagen Bank area ranged in depth from 20 to 50m, with a mixture of sand, gravel and shell debris habitats formed by large storm waves. These storm events are intermittent compared with the daily scallop dredging activity in the area. ROV surveys revealed that the area was characterized by dense aggregations of the hydrozoan Corymorpha pendula (Agassiz) which provided shelter for shrimp, Dichelopandalus leptoceros (Smith). Wide linear swathes through benthic microalgal cover indicated the occurrence of recent trawling and scallop dredging activity. The hydrozoans and associated shrimps were absent from these fished areas (Auster et al., 1996). Recovery from disturbance may be rapid. Collie et al. (1997) found that the biogenic epifauna at a site, which had previously been dredged for scallops, and then closed to fishing, showed signs of recovery after 2 years and Kaiser et al. (1998) found that epifaunal communities that had been trawled over experimentally in relatively shallow (35 m) water were indistinguishable from control unfished areas after 6 mon hs. The effects of fishing on epifaunal communities may have ramifications for plankton communities which are often dominated by the larvae of invertebrates. The mesozooplankton taken in continuous plankton recorder samples in the central North Sea were numerically dominated by calanoid

t

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copepods from 1958 to the late 1970s, whereas samples taken from the same stations from the early 1980s to early 1990s were dominated by the pluteus larvae of echinoid and ophiuroid echinoderms. This trend is consistent with the reported increases in the abundance of echinoderms in benthic communities which may have been stimulated, in part, by bottom trawling (Lindley et al., 1995). Where fishing occurs in shallow clear waters, marine plant communities are likely to be affected. In particular, seagrass meadows are vulnerable to physical disturbance as dredges and trawls reduce plant biomass and abundance by shearing off fronds, exposing rhizomes,digging shoots from the substratum and increasing local turbidity through sediment resuspension (Fonseca et al., 1984). GuillCn et al. (1994) reported that 45% of a Posidonia meadow in SE Spain was damaged by trawling and that in some areas the meadow no longer bound sediment effectively. Seagrass meadows are highly productive, support complex trophic food webs, provide sediment and nutrient filtration, enhance sediment stabilization and act as breeding and nursery areas for species of commercial importance (Short and Wyllie-Echeverria, 1996). The studies that we have reviewed clearly illustrate the two main effects of mobile gears on epifaunal communities: (i) modification of substrata (shell debris, boulders, mud veneers) and (ii) removal of biogenic taxa and a consequent decline in the abundance of fauna associated with them (see Section 2.3). The loss of biogenic species not only reduces the supply of important prey species, but also increases predation risk for juvenile commercial species, thereby lowering subsequent recruitment to the adult stocks (Walters and Juanes, 1993). Bradstock and Gordon (1983) reported the removal of extensive beds of bryozoans as a result of trawling activity and advocated the protection of these communities, noting that they provided an important habitat for juvenile commercial fishes. Moreover, Dayton et al. (1995) discuss the importance of different functional groups in maintaining community structure. Communities dominated by long-lived suspension feeders are most likely to be replaced by a community of opportunistic deposit-feeding species and mobile epifauna when subjected to large-scale and intense fishing disturbance. More dramatically, biogenic structures that increase the complexity of the epibenthic habitat (e.g. corals, bryozoans, worm tubes) create specialized environmental conditions by altering local hydrographic conditions that encourage the development of a specialized associated community. Loss of such structures will also affect the survivorship of any associated species and prolong the recolonization process.

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2.2.2. Static Fishing Gears

Static bottom gears are anchored to the seabed and left to fish passively. The most commonly used are gill, trammel or tangle nets, which are designed to capture target species by enmeshing or tangling them (Millner, 1985; Potter and Pawson, 1991). Traps and pots are commonly anchored to the seabed in fleets, each pot or trap is baited to attract target species through one or more entrances into chambers in which the animals are trapped. Reefs are frequently damaged by the hauling of set nets, and the problem has been exacerbated by the use of mechanical net haulers or power blocks (Munro et al. 1987). The effects are regarded as minor in comparison with those attributable to drive netting and other active fishing techniques (see Section 2.2.3). Since set net and pot fisheries are static, the areas of seabed affected by each gear are likely to be insignificant compared with the widespread effects of mobile fishing gears. However, effort may be significant if concentrated in relatively small areas with communities of long-lived fauna. A recent study evaluated the effects of pot deployment and retrieval on supposedly fragile epifauna that are the subject of conservation interest in northern Europe (En0 et al., 1996). Not surprisingly, pots that landed on, or were hauled through beds of the foliose bryozoan Pentapora foliacea (Ellis and Solander) caused physical damage to the brittle colonies. However, contrary to expectations, sea pens, Pennatula phosphorea, Virgularia mirabilis O.F. Muller and Funiculina quadrangularis Pallas bent in response to the pressure wave created by the descending pot and lay flat on the seabed. Moreover, when uprooted, the sea pens were able to reestablish themselves in the sediment. Sea fans, Eunicella verrucosa (Pallas) were also found to be more flexible than anticipated, and were not severely damaged when pots were hauled over them (En0 et al., 1996). These observations were interesting, because sea pens and sea fans were considered to be highly vulnerable to fishing activities (MacDonald et al., 1996). The study of Eno et al. (1996) suggests that the direct contact of fishing gears with fauna may not be the primary cause of mortality and the frequency and intensity of physical contact is more likely to be important. When nets or pots are lost, either because of bad weather, snagging or when towed away by mobile fishing gears, they may continue to fish. This phenomenon is known as “ghost-fishing”. In contrast to the numerous records of bird, reptile and cetacean entanglement in set gears (see Dayton et al., 1995 and references therein), little is known about the frequency of net loss or for how long lost gear is likely to fish. This lack of knowledge results from the reluctance of fishers to report such incidents and the difficulty in undertaking long-term studies in a realistic manner. Estimates of the proportion of nets lost from commercial fleets have been reported in a variety of studies reviewed by Dayton et al. (1995). Losses of

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gear appear to be substantial. Approximately 7000 km (20-30% of the total set each day) of drift nets were lost per year in a north Pacific fishery (Eisenbud, 1985). Complaints by fishers, prompted a grapnel survey of the seabed on Georges Bank which yielded 341 actively fishing ghost nets from 286 tows (Brothers, 1992). The phenomenon of ghost fishing was clearly perceived to have negative effects on commercial stocks by commercial fishers in the Greenland halibut fishery, who instigated their own voluntary clean-up programme (Bech, 1995). Considerable numbers of pots are also lost each year in North America. It was estimated that the 31 600 pots lost in the Bristol Bay king crab fishery removed c. 80000kg of crabs from the stock (Kruse and Kimber, 1993). In another study, Breen (1987) reports an annual loss of c. 11% of the traps used in the Dungeness crab Cancer magister Dana fishery in British Columbia. Both lost nets and pots can persist and continue to fish in the marine environment for several years (Carr et al., 1992), although their actual persistence will depend on the prevailing environmental conditions. Nets lost in areas exposed to large swells and storm activity are rapidly destroyed by physical forces (E. Puente, pers. comm.). Those lost in shallow, clear waters are rapidly overgrown with epibiota which makes them highly visible, reducing their fishing capabilities (K. Erzini pers. comm.). However, in circumstances where nets or pots are snagged onto rocks, holding the net in place, or lost in deep water in a relatively stable environment, they may continue to fish indefinitely (Carr et al., 1992). Recent studies have shown that in these cases, a typical pattern of capture is observed. Over the first few days, catches decline almost exponentially as the increasing weight of catch causes the net to collapse. Then, for the next few weeks, the decaying bodies of fishes and crustacea attract a large number of scavenging crustaceans, many of which are valuable commercial species and also become entangled in the net. Thereafter, there appears to be a continuous cycle of capture, decay and attraction for as long as the net has some entanglement properties (Carr et al., 1992; Kaiser et al., 1996a). Pots tend to be constructed of robust materials and have a rigid structure which means that lost pots are likely to maintain a higher capture efficiency for much longer than lost nets. Not surprisingly, ghost-fishing mortality rates of up to 55% of the mortality rates recorded in attended pots have been reported (High, 1976; Miller, 1977). A rebaiting cycle occurs in lost pots as described for lost nets above, which suggests that an intact pot could fish indefinitely (B. Bullimore, pers. comm.). The “ghost-fishing” potential of pots also varies for different fisheries and pot designs. For example, Parrish and Kazama (1992) found that the majority of Hawaiian spiny lobster, Palinurus marginatus and slipper lobster, Scyllarisdes squammosus were able to escape traps, whereas parlour-type traps lead to mortalities of 12-25% for American lobster, Homarus americanus (Smolowitz, 1978).

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Compared with the proportions of target species removed by mobile fishing gears, the number of organisms removed by ghost-fishing is probably small. However, these fisheries tend to be highly localized leading to a concentration of lost gear within relatively small areas. Consequently, the proportion of local stocks removed can be significant (Kruse and Kimber, 1993). Furthermore, many of these species have a high individual value and hence represent a large economic loss to the local fishing industry. In order to reduce these losses for undersized specimens, escape panels are now fitted to many pots used in North America and biodegradable materials are used to ameliorate losses from “ghost-fishing” (Guillory, 1993; Polovina, 1994).

2.2.3. Drive Netting, Poisons and Explosives Techniques such as drive-netting, pull-seining, poison and explosive fishing are principally used by small-scale and artisanal fishers fishing on tropical reefs. Although the effects attributable to the activities of individual fishers are often 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 er al., 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 explosive fishing (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 nontarget fishes and invertebrates (Rubec, 1986; Eldredge, 1987; McAllister, 1988; 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 bioerosion. Fishing affects reefs directly when gears contact the reef substratum or indirectly by altering the relationships between those communities of plants, invertebrates and fishes which determine rates of reef accretion and bioerosion (see Section 2.3). 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.

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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 extensively used on coral reefs, and may range from smallscale village-based operations involving four or five fishers to large commercial operations which 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 are often broken deliberately to scare closely reef-associated fishes such as groupers (Epinephelinae), snappers or emperors from their refuges. In deeper water, the Kayakas and muro-ami drive-netting techniques involve teams of swimmers which repeatedly drop weighted scarelines onto the reef in order to drive fishes towards a bag net. Carpenter and Alcala (1977) calculated the damage to 1ha of reef during a single muro-ami operation involving 50 fishers who each struck the bottom 50 times with a 4 kg weighted scareline. Six percent 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; Alcala and Gomez, 1987; 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 which 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 1.15m, and that a gallon-sized drum will have the same effect within a radius of 5 m. A “typical” charge will kill most marine organisms including invertebrates within a radius of 77m. 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 many reefs from 15-30’ either side of the equator, which are susceptible to hurricane damage, the effects of blast fishing are often localized and negligible in comparison with those of hurricanes (S. Jennings, pers. obs.). However, intensively blast-fished areas will be subjected to chronic rather than the acute damage associated with natural events. In areas such as the Philippines, damage attributable to blast fishing is an increasing cause of concern. Stupefacients are widely used by reef fishers. Traditionally, poisons extracted from plants were extensively used for reef fishing, but in the last few decades, synthetic chemicals such as sodium cyanide and chlorine have

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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 which have an acute effect are quickly dispersed, the chronic effects may be significant. The long-term direct effects of fishing on reefs are largely determined 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 regrow 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-’ for branching species and 0.5-1.9 cm year-’ 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, 1990; Done, 1987, 1988; Done et al., 1988; Endean et al., 1988; 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). 2.3. Indirect Effects on Habitat

The direct effects of fishing change the structure of fish and benthic communities and may lead to the resuspension of sediments. Changes in the structure of fish and benthic communities may affect the growth of those organisms which are responsible for structuring habitats. The resuspension, transport and subsequent deposition of sediment may affect the settlement and feeding of the biota in other areas. Trawling, in particular, can be responsible for resuspending a large proportion of the sediment load in some marine environments. Those parts of the trawl net that come into contact with the sea bed will cause bottom sediments to be resuspended but the turbulence created by the trawl doors suspends most material and plays a key role in herding fishes towards the net (Main and Sangster, 1981). The quantity of sediment resuspended by trawling depends on sediment grain size and the degree of compaction which is higher on mud and fine

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sand than on coarse sand. Sediment concentrations of 100-550 mg 1-' have been recorded lOOm astern of shrimp trawls in the muddy Corpus Christi Bay (Schubel et al., 1978) and Churchill (1989) reported that transmissometers, which had been employed to record turbid water parcels, frequently recorded the highest levels of turbidity during periods of trawling activity. In deeper areas where storm-related bottom stresses were generally weak, the quantity of sediment resuspended by otter trawling was significant. Churchill (1989) produced sediment budgets for parts of the midAtlantic Bight and concluded that trawling was the main factor initiating the offshore transport of sediment at depths of 100-140m (Figure 6). However, the transport of sediment resulting from fishing activities would not produce significant large scale erosion over a period of a few years. In deeper water, where currents are weak and sediment is rarely in suspension, the effects of resuspension and subsequent deposition are readily detected. Thiel and Schriever (1990) investigated the potential effects of mining polymetallic nodules at depths greater than 4000m by harrowing the sediment with an 8-m wide rake. Having harrowed 20% of the study area during 78 traverses, the remaining 80% of the area was affected by the redeposition of sediment. The potential effects of sediment resuspension include clogging of feeding apparatus or reduction of light availability (Rhoads, 1974) and

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Figure 6 Histograms indicating the mass (monthly mean) of sediment put into suspension by trawling on the muddy seabed of the Middle Atlantic Bight from 1 January 1985 to 31 March 1985 (a) and from 1982 to 1985 (b). Circles indicate the mass (monthly mean) of sediment resuspended by currents from 1 January 1984 to 23 March 1984. (Redrawn with modifications from Churchill, 1989.)

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sediment deposition has been shown to inhibit the settlement and growth of oysters and scallops (Moore, 1977; Jones, 1992). However, given the range of sediment types in the marine environment and the natural spatial and temporal variations in sediment load (Moore, 1977), it is unlikely that the population level consequences of sediment resuspension and deposition can be determined from small scale studies of siltation effects. The surface of marine sediments is an important site of benthic production. Brylinsky er al. (1994) demonstrated that the biomass of benthic diatoms (measured as chlorophyll a) was significantly less in trawl door furrows on a muddy substratum in shallow water. However, one month after the trawling had taken place there was a diatom bloom in the furrow, which Brylinsky et al. (1994) attributed to the release of nutrients from the sediment. Emerson (1989) considered the effects of sediment disturbance resulting from wind stress on production in the southern North Sea, and found a significant negative correlation between wind stress and total macro- and meiobenthic production. The intensive trawling of Posidoniu 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 resuspended sediment (GuillCn er 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 that large-scale changes in primary production could be reliably correlated with changes in fishing intensity using existing data. However, given that a large proportion of the continental shelf area is now trawled, and that tools such as stable isotope analysis can be used to trace the origins and transformations of organic matter in the marine environment (Owens, 1987; see Section 5.3), it is increasingly likely that the impacts of fishing on the relative roles of benthicand planktonic-based food chains could be investigated. The most convincing evidence for the indirect effects of fishing on habitat structure comes from the study of fishing effects on coral reefs. The direct effects of fishing (see Section 2.2.3) have been widely reported because popular reef fishing techniques often cause rapid and highly visible damage. However, the intensity and selectivity of fishing practices may have been responsible for initiating the transition of reef communities between relatively stable algal and coral-dominated phases (Levitan, 1992). Understanding the ways in which fishing can lead to shifts in ecosystem state is dependent on an understanding of the roles of herbivores and corallivores in the reef ecosystem. Herbivorous and corallivorous species both erode the reef matrix and the rate at which coral reefs grow is determined by the relative rates of coral accretion and erosion. Some species of parrotfishes (Scaridae) and urchins erode the reef matrix while feeding on algae (Birkeland, 1989; Bak, 1990, 1994; McClanahan, 1992, 1995a; Bellwood,

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1995) and may markedly reduce rates of reef accretion (Glynn et al., 1979; Birkeland, 1989; Macintyre et al., 1992). Corallivorous species such as the crown-of-thorns starfish Acanthaster planci (L.) and the gastropod Drupella spp. cause erosion by feeding directly on coral (Moran, 1986; Turner, 1994). Herbivorous fishes and invertebrates not only determine rates of reef accretion but may also have a substantial impact on the distribution and abundance of reef algae (Lawrence, 1975; Brock et al., 1979; Hay, 1984, 1985; Hay and Taylor, 1985; Lewis and Wainwright, 1985; Carpenter, 1986; Lewis, 1986). Herbivorous fishes may clear space for coral settlement and thereby enhance the survival and growth of young coral colonies but, unlike urchins, most herbivorous fishes do not damage these colonies once they are growing (Potts, 1977; Bak and Engel, 1979; Brock et al., 1979; Sammarco, 1980; Lessios et al., 1984; Hughes et al., 1987b). Herbivorous fishes are targeted by many reef fishers and their abundance may be significantly reduced on intensively fished reefs. Sea urchins, however, are rarely important target species and once the biomass of herbivorous fishes is reduced by fishing the urchins begin to dominate the grazing community. In addition, the biomass of those fishes which prey on urchins (Hiatt and Strasburg, 1960; Randall, 1967; Hoffman and Robertson, 1983; Reinthal et al., 1984; McClanahan, 1995b) is also reduced by fishing, since many of these species, in particular triggerfishes (Balistidae) and emperors, are targets of reef fishers or are easily caught because of their aggressive behaviour. McClanahan (1992) developed a biomass-based energetic model to describe algal grazing by sea urchins and herbivorous fishes which suggested that sea urchins would tolerate low algal biomass owing to their low consumption and respiration rates. This enables them to persist at low levels of algal biomass and productivity, out-competing herbivorous fishes and reaching maximum biomass levels an order of magnitude higher. As a result, once an urchin-dominated community is established it is unlikely that herbivorous fishes can re-establish themselves (McClanahan and Shafir, 1990; McClanahan, 1992). Relationships between predatory fishes and their urchin prey have been explored on Kenyan reefs by comparing herbivore communities at a series of sites subject to different fishing intensities (McClanahan and Muthiga, 1988; McClanahan and Shafir, 1990; McClanahan and Obura, 1995) where the triggerfish Balistapus undulatus (Park) and wrasse Chelinus trilobatus (L.) are the main urchin predators (McClanahan, 1995b). Predator removal through fishing appeared to result in the ecological release of sea urchins and the competitive exclusion of weaker competitors such as herbivorous fishes. Thus the more heavily exploited Kenyan reef lagoons were characterized by denser populations of larger sea urchins, fewer and smaller fishes and reduced coral cover (McClanahan and Muthiga, 1988). Changes in fishing pressure and urchin mortality and urchin recruitment are responsible

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for shifting these Kenyan reef ecosystems to different, and relatively stable, states (Figure 7). McClanahan (1992) suggested that the persistence of herbivorous fishes on many reefs may be dependent on the presence of sea urchin predators which maintain sea urchins at a level that prevents them becoming dominant. Similarly, the structure of urchin Paracentrotus lividus (Lmk) populations in the northwestern Mediterranean also appear to be controlled by fish predators which are affected by fishing. Sala and Zabala (1996) demonstrated that the density of urchin populations in a marine reserve where their potential fish predators were abundant was significantly lower and that the urchins tended to adopt a crevice dwelling behaviour rather than feeding on exposed surfaces. When the effects of intensive exploitation on herbivorous fish populations are coupled with a decrease in the abundance of invertebrate herbivores, the resulting increase in algal biomass may have a marked influence on the development of coral reefs. Thus the mass mortality of the algal feeding sea urchin Diadema antillarum, Philippi in those intensively fished regions of the Caribbean where herbivorous fishes were already scarce (Bak

Fish and coral dominance Urchins: High predation Low recruitment

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Figure 7 A conceptual model of the coral reef ecosystem. The model indicates those processes which cause shifts between three equilibrium states and which have been demonstrated by simulation and empirical studies. (Redrawn with modifications from McClanahan, 1995a.)

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et al., 1984; Lessios et al., 1984, 1985; Carpenter, 1985, 1988a, 1990; Hughes et al., 1987a,b; Lessios, 1988) was followed by significant increases in algal cover and significant decreases in coral cover (Hughes et al., 1987b; Done, 1992; Knowlton, 1992; Hughes, 1994). Crown-of-thorns starfish feed directly on living coral and outbreaks of this species have led to widespread decreases in live coral cover and reef structural complexity (Goreau et al., 1972; Endean and Stablum, 1973; Glynn, 1973; Nishihira and Yanmazato, 1974; Faure, 1989; Zann et al., 1990) reducing the availability of suitable habitat for reef fish communities (Sano et al., 1984; Bouchon-Navaro et al., 1985; Williams, 1986). The existence of an inverse relationship between the abundance of crown-of-thorns starfish and their fish predators suggests that starfish population outbreaks could have resulted from the removal of fishes such as emperors and triggerfishes that prey upon juvenile starfish (Ormond et al., 1991). Keesing and Halford (1992) documented mortality rates of over 6% d-' for crown-ofthorns starfish which had recently settled to the benthic habitat and attributed this to predation. Sweatman (1995) studied predation on juvenile crown-of-thorns starfish in one location on the Great Barrier Reef and his data suggested that the predation rates would be too low to regulate crownof-thorns starfish populations. At present, there is only weak inference to suggest that removal of predators is responsible for outbreaks of crown-ofthorns starfish. Further studies to test quantitative hypotheses on larger temporal and spatial scales are needed to determine the indirect impacts of fishing on crown-of-thorns starfish populations. If intensive fishing can lead to crown-of-thorns outbreaks then there is good evidence to suggest that this would reduce the potential fish production from a reef ecosystem. The muricid gastropod Drupella sp. is corallivorous and rapid increases in its population density have led to coral mortality approaching that caused by crown-of-thorns starfish (Turner, 1994). The removal of Drupella predators by fishers has been cited as a possible cause of these outbreaks (Turner, 1994) but the significance of fishing cannot reliably be determined using the limited data currently available (McClanahan, 1994b; Ayling and Ayling, unpublished). Degradation of reef habitats which results from the direct or indirect effects of fishing will affect fish yield, both by causing a redistribution of the exploitable fish biomass and, in severe cases, by reducing the potential production of that ecosystem. Russ and Alcala (1989) suggested that reduced butterflyfish (Chaetodontidae) abundance in a newly exploited Philippine reserve was a result of a reduction in live coral cover associated with destructive fishing techniques, although they commented on the difficulty of differentiating between effects resulting from direct removal of fishes and those resulting from habitat modification. Porter et al. (1977) noted a significantly lower biomass of zooplankton was associated with

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rubble rather than coral habitats and it might be expected that fish density would change in response to such changes in food supply. The abundance of many reef fishes is positively correlated with topographic complexity (Risk, 1972; Porter et al., 1977; de Boer, 1978; Luckhurst and Luckhurst, 1978; Carpenter et al., 1981; Thresher, 1983; Kaufman and Ebersole, 1984; Patton et al., 1985; Roberts and Ormond, 1987; Grigg, 1994; Jennings et al., 1996a) and habitat complexity will also influence the rates at which larval fish recruit to the reef from the plankton (Jones, 1988; Connell and Jones, 1991). Most of these studies demonstrate habitat effects by making comparisons between sites within regions of high habitat complexity. The differences are greater when large well-developed areas of reef are compared with areas that have been fished destructively until little topographic complexity remains (Pauly et al., 1989). Changes to kelp bed habitats in temperate waters have been attributed to the indirect effects of fishing, but subsequent re-examination of the evidence for these changes suggested that fishing was not the cause. On the Atlantic coast of Nova Scotia, the reduction in the biomass of American lobster was assumed to have led a reduction in the predation rates on the sea urchin Strongylocentrotus droebachiensis (O.F. Muller). As a result, urchin populations flourished, leading to the destruction of kelp beds (Mann and Breen, 1972; Mann, 1982). However, more detailed analysis of feeding rates, stomach contents and biomass of lobsters indicated that they could not have controlled the population structure of urchins (Miller, 1985) and that the increases in urchin populations which led to the destruction of kelp beds may have been stimulated by increased larval recruitment (Hart and Scheibling, 1988; Elner and Vadas, 1990). 2.4. Natural Versus Fishing Disturbance

To date, most studies have investigated the effects of fishing on benthic communities in shallow seas on the continental shelf at depths greater than 100m. This is not surprising as the majority of demersal fishing activity occurs in this depth range, and quantitative ecological studies become logistically complex at greater depths. Benthic communities in these environments experience continual disturbance at various scales (Hall, 1994). Large-scale natural disturbances, such as seasonal storms and strong tidal currents, form a background against which other smaller disturbances occur, such as those induced by predator feeding activities (von Blaricom, 1982; Oliver and Slattery, 1985; Hall et al., 1994). Hall et al. (1994) suggested that frequent small-scale predator disturbances may have a considerable additive effect on benthic communities, creating a long-term mosaic of patches in various states of climax or recolonization (Grassle and Saunders,

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1973; Connell, 1978). It was concluded, however, that while it was possible to detect short-term effects of predator disturbance, large-scale effects could not be inferred. This implies that small-scale disturbance events, even when frequent, are masked by a background of large-scale disturbances or that the small-scale of disturbance permits rapid recolonization such that largescale effects never become apparent. These observations are summarized in Figure 8. Clearly, the scale and frequency of disturbance events can increase until lasting ecological effects can be observed against a background of natural disturbance. The additive effects of an entire fishing fleet may reach such a threshold. Moreover, fishing effort in shelf seas is not homogeneously distributed. Fishers concentrate their effort in grounds that yield the best catches of commercial species and avoid areas with obstructions and rough ground that would damage their gear. In addition, fishing is severely restricted in some areas, such as shipping lanes and around oil rigs. Consequently, early estimates of area swept by bottom gears are unintentionally misleading as they imply physical disturbance spread homogeneously across large (> 100 km2) areas (Welleman, 1989). More recently, “black box” recorders have been fitted to a proportion of the Dutch beam trawl fleet allowing them to be tracked during fishing operations. The Dutch fleet accounts for 50-70% of the total beam trawling effort in the North Sea (Rijnsdorp et al., 1996a).These records indicate that beam trawling effort is very patchily distributed in the North Sea; while it is estimated that some

Figure 8 A conceptual model which illustrates the relative importance of fishing disturbance in habitats that are subject to different levels of natural disturbance. As levels of natural disturbance decline, fishing disturbance accounts for a greater proportion of the total disturbance experienced (d, d f ) and becomes increasingly important.

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areas are visited over 400 times per year, others are never fished (Rijnsdorp et al., 1996a, 1997; Plate 1). The distribution of bottom trawling disturbance can also be ascertained from the occurrence of physical damage in populations of animals that are able to withstand such injuries. Up to 55% of the population of the starfish, Astropecten irregularis Pennant had lost arms in a heavily beam-trawled area of the Irish Sea, compared with only 7% in a less intensively fished area (Kaiser, 1996). Within intensively fished grounds, the background levels of natural disturbance may have been exceeded leading to long-term changes in the local benthic community. However, as pointed out by many previous authors, communities observed at the present time may be the product of decades of continuous fishing disturbance (Bergman and Hup, 1992; de Groot and Lindeboom, 1994; Dayton et al., 1995). Detecting those long-term changes in benthic fauna which can be attributed to fishing activities has been problematic in all but the most obvious cases (Riesen and Riese, 1982; Sainsbury, 1987). However, a few long-term datasets which record by-catches of benthic species have revealed reductions in potentially vulnerable species or changes in epibenthic communities. Philippart (1998) examined a dataset of returns of epibenthic by-catch species from the southern North Sea dating back to the 1930s. Fishers were paid to retain examples of a selection of species and deliver them to the Netherlands Institute for Sea Research. Beam trawling superseded otter trawling as the main Dutch fishery from about 1970. Consequently, landings of benthic species might have been expected to increase as beam trawls catch a larger proportion of benthic species compared with otter trawls. However, the decrease in the incidence of species returned to the laboratory continued after 1970. Furthermore, Holtmann et al. (1996) reported a decrease in the abundance of the fragile burrowing heart urchin and the brittlestar Amphiura filiformis O.F. Muller in areas of the southern North Sea between 1990 and 1995. These trends suggest that fishing activity may have been the main cause of these changes. However, it is problematic to attribute these changes to fishing alone, as the southern North Sea has been influenced by eutrophication events leading to increases in the abundance of polychaete species and echinoderms such as A . filiformis (Pearson et al., 1985) and by oceanographic changes (Lindeboom et al., 1995). These observations emphasize the value of time-series data for identifying the factors which have had most influence on changes in community structure (see Section 5). 2.5. Conclusions

Fishing activities lead to changes in the structure of marine habitats and influence the diversity, composition, biomass and productivity of the as-

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sociated biota. The direct effects of fishing vary according to the gears used and the habitats fished, but they usually include the scraping, scouring and resuspension of substratum and occur against a background of natural disturbance. The relative impact of fishing on habitat and benthic community structure is determined by the magnitude of natural disturbance. The direct effects of a given fishing method on infaunal and epifaunal communities will tend to increase with depth and the stability of the substratum. In sheltered areas where complex habitats develop at minimal depth, such as coral reefs, the direct effects of fishing may be marked and have profound effects on the ability of the habitat to sustain fish production. The indirect effects of fishing on sea urchin populations and their subsequent effects on the rate of accretion and bioerosion in the reef habitat are one of the few well documented examples of top-down control in marine ecosystems (see Section 4). When a few species of predator, all of which may be fished, selectively feed upon one or two species of urchin which otherwise dominate the herbivore community on a reef, they have an unusual role of keystone species in a marine system. However, the tightly coupled relationships between urchins and their fished predators should not be regarded as ubiquitous and further work is needed to determine the relative roles of predator and environmental control in other ecosystems.

3. FISH COMMUNITY STRUCTURE

3.1. Introduction

Fishing has direct effects on fish community structure. Changes in the growth, mortality, production and recruitment of target fish populations provides the basis for conventional fisheries assessment and management (Beverton and Holt, 1957; Beverton, 1963; Cushing, 1968; Nikolskii, 1969; Gulland, 1977; Munro and Williams, 1985; Hilborn and Walters, 1992; Appeldoorn, 1996). However, the capture of target or by-catch species also has indirect effects on fish populations and the direct and indirect effects of fishing act in combination to determine the resulting biomass, size structure and diversity of communities. In this chapter we describe the effects of fishing on the diversity and size-spectra of fish communities and the life history traits of fishes. Fishing affects the diversity of fishes and other species and there is increasing concern that, together with pollution, fishing is a major cause of diversity loss in aquatic ecosystems (Hutchings, 1986; Ryman er al., 1995). One of the key reasons for justifying the maintenance of diversity in terrestrial and aquatic ecosystems has been that diversity may confer

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ecosystem stability in the face of anthropogenic impacts and human exploitation. In Section 3.2, we discuss the changes in species and intraspecific diversity which may be induced by fishing and their effects on the productivity and stability of ecosystems. Since many fishing techniques are size-selective (see Section 2.2), changes in the size structure of populations should be expected following fishing. Decreases in the mean size of target fishes and reductions in the abundance of larger fishes are one of the most widely reported and quickly observed changes when fishing effort increases (Russ, 1991). Since the size structure of the biota in marine ecosystems follows relatively regular patterns there may be general changes in these patterns as a result of fishing (Rice and Gislason, 1996). These may result from the direct effects of size-selective fishing and the indirect effects of fishing on predator-prey relationships (see Section 4). We consider these changes and their ecological significance in Section 3.3. Size-selective fishing will affect species with different life history traits in different ways, and these effects are reviewed in Section 3.4. Since species with late maturity and slow growth towards a large maximum size are typically affected more by size-selective fishing than small fast-growing species with early maturity, it might be expected that the species composition of fish communities will change in response to fishing and that smaller fastgrowing species will dominate the biomass. Moreover, within species, fishing is selective with respect to a number of life history traits such as growth, which are at least partially heritable, and exploited populations would be expected to evolve in response to harvesting. We do not reiterate much of the material describing the responses of target populations to fishing but we do describe changes in growth and reproduction, many of which have a genetic component. To date, such evolution of life histories has often been overlooked because it is slow in comparison with the periods in which managers, operating under contemporary socioeconomic constraints, have to act. However, given that ecosystem-based management may operate on longer time scales, the effects of fishing on life history traits need to be considered as part of any holistic management strategy.

3.2. Diversity 3.2.1. Local Extinctions and Redundancy

While it is increasingly clear that fishing has been responsible for the local extirpation of species such as giant sea bass Stereolepis gigas Ayres in California, skate Raja batis L. in the Irish Sea (Brander, 1981) and various grouper species in the tropical Indo-Pacific (Randall and Heemstra, 1991)

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fishing is relatively unlikely to cause species extinction (Beverton, 1990, 1992a). The probability of the extinction of marine species resulting from fishing is low in comparison with well-documented effects of hunting on terrestrial animals (Steadman, 1995) because economic extinction occurs first and may deter furthur exploitation. As such, the allocation of marine fishes to endangered categories has been rather arbitrary and may reflect the personal interests of those who compile the lists rather than the probability that the species will be lost. It is expected that the indirect effects of fishing and other human activities on habitat will lead to more extinctions than the attempts of fishers to remove fish. There is good evidence for local reduction in species richness following fishing. On Fijian reefs, the diversity (measured as species richness) of target groupers and snappers was significantly correlated with fishing intensity (Figure 9), as was the diversity of snappers and emperors on Seychelles reefs (Jennings et al., 1995; Jennings and Polunin, 1997). Comparisons between fished and unfished (marine reserve) areas on coral reefs have consistently shown that species richness is higher in the unfished area. Such studies have been conducted in Kenya (McClanahan and Muthiga, 1988;

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Samoilys, 1988; McClanahan, 1994a; Watson and Ormond, 1994), the Caribbean (Bohnsack, 1982; Roberts and Polunin, 1993), Seychelles (Jennings et al., 1996b) and the Red Sea (Roberts and Polunin, 1992). In the North Sea, Rice and Gislason (1996) looked at changes in the diversity (measured as Shannon-Weiner index, which incorporates both species richness and eveness elements (Magurran, 1988)) across length classes for the demersal fish assemblage as sampled by trawl surveys from 1977-1993: a period of increasing fishing intensity. The slope of the diversity spectra did not change consistently with time. Indeed, Greenstreet and Hall (1996) compared diversity in the North Sea groundfish assemblage using a longer term data set from the periods 1929-1953 and 1980-1993 and found that species diversity was only marginally greater for the whole assemblage in the early period. There are some key differences between the studies in tropical and temperate waters. First, the tropical studies have been conducted in small areas and the fishes which were studied are relatively site-attached and nonmigratory. Secondly, there is concern that, in the North Sea, large changes in diversity took place well before scientific investigations began, since this area has been fished intensively for over a century and the population level effects of fishing were apparent in the last century (Anon., 1885) (see Section 3.4 and Figure 4). Many local species losses in the North Sea were reported before scientific trawl surveys began (Knijn et al., 1993). Local reductions in the species richness of reef fishes following fishing are not necessarily associated with clear decreases in the yield from reef fisheries (Jennings and Polunin, 1995, 1996a, 1997) and this has prompted commentators to suggest that some species can be considered redundant. If fish production is considered to be the only ecosystem function of concern, then it is reasonable to suggest that the local loss of one or two snapper species, when 15 or more species are relatively abundant, is unlikely to lead to detectable changes in the total sustained yield from these fisheries. However, while the extirpated species may be functionally redundant at the present time, the value of ecological functions will change. Terrestrial ecologists increasingly recognize the weakness of the redundancy concept and thus there is greater emphasis on the idea of functional similarity (Ehrlich and Daily, 1993; Schulze and Mooney, 1993; Abrams, 1996; Collins and Benning, 1996; Martinez, 1996). The concept of functional similarity allows the level of similarity to be quantified (at a given place or time) whereas redundancy implies a non-constructive “valuable or worthless” “all or none” perspective (Collins and Benning, 1996). Furthermore, the extent of functional similarity is expected to vary with levels of resolution from physiological through population to ecosystem. Even if species richness does not play an important role in maintaining ecosystem processes at present, it may be necessary to ensure that there are a range of species to fulfil new roles when conditions change. Moreover, since the recruitment of

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related species rarely varies in unison, diversity ensures that some members of a functional group may be abundant at any given time. It is notable that those systems in which the most dramatic ecosystem effects of fishing are reported are those in which very few species fulfil key functional roles (see Section 4). Martinez (1996) concludes that functional redundance has anthropogenic and pejorative connotations, which implies an ability of ecologists to see into the future with absolute certainty. Conversely, functional equivalence or similarity at a place and time is acceptable. Most “longterm” ecological studies, especially in tropical waters, have been operative for periods rather shorter than the lifespan of a large skate, Atlantic halibut Hippoglossus hippoglossus (L.) or grouper (Jennings and Lock, 1996), and do not allow us to predict the role of species in future years. 3.2.2. Indirect Diversity Losses Fishing may lead to indirect changes in the species diversity of fishes or invertebrates by affecting predator-prey relationships or changing habitat structure. On tropical reefs it is well known that the relatively homogeneous algal-or urchin-dominated habitats which may develop in intensively fished areas (see Section 2.3) harbour a fish community with lower species richness (McClanahan, 1994a). However, if there are marked changes in the abundance of piscivorous fishes, without corresponding changes in habitat, then the available evidence suggests that changes in prey diversity are unlikely. On Fijian reefs, Jennings and Polunin (1997) recorded marked decreases in the biomass of piscivores with increasing fishing effort (Figure lo), but the species richness of all herbivorous and invertebrate feeding fishes which were the potential prey of target species, but which were not targeted by the fishery, was not correlated with the biomass of piscivores present (Figure 11). Similarly, on Seychelles reefs, where there were also significant decreases in the biomass of piscivorous and piscivorous/invertebrate feeding fishes, there were no increases in the species richness of all herbivorous and invertebrate feeding prey which were not targeted by the fishery (S. Jennings, unpublished data). In these reef ecosystems there are no indications that predator abundance and prey species richness are closely linked. This contrasts with the situation in limnetic systems such as the African lakes; but we would suggest that the structures of food webs in limnetic and reef ecosystems are not comparable (see Section 4.2 ). 3.2.3. Diversity Loss and Ecosystem Stability Species extinction is only one facet of diversity loss (Gaston, 1996). Many fished species have latitudinal ranges of thousands of kilometres and self-

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sustaining populations of the same species inhabit many different ecosystems. One concern is that the losses of some populations following overfishing will lead to increased instability in the ecosystem. The study of ecosystem stability, complexity and fragility and the impact of diversity loss on these processes has been an active field of ecological research. The extent to which a system is viewed as stable depends on the measure of stability and the spatial, temporal, taxonomic (species, trophic group, etc.) and numeric (presence/absence or abundance) scales considered. Thus an ecosystem which is highly fragile over a few years may be highly stable over centuries, or an ecosystem which is fragile at small spatial scales may be stable when viewed at larger scales. These arguments are developed by Pimm (1991) and Nilsson and Grelsson (1995). For the present purposes, we regard one of the key measures of stability to be total biomass and production of fish protein on temporal scales of years and decades and

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spatial scales of tens and hundreds of kilometres. Continued production of protein is effectively an ecosystem service, given that it is beneficial to humans (Westman, 1977). Individual species show little stability in their production over these time scales and longer (Soutar and Isaacs, 1974; De Vries and Pearcy, 1983; Shepherd and Cushing, 1990) but total fish production from those intensively exploited ecosystems which continue to support a relatively diverse assemblage of fishes, such as the North Sea, is remarkably stable, even though the component fishes pass through boom and bust cycles (Ursin, 1982; Figures 12 and 13). Relationships between diversity and stability have been widely explored by theorists (MacArthur, 1955; May, 1972, 1973; Walker, 1992; Ehrlich and Daily, 1993; Lawton and Brown, 1993; Lawton, 1994; Naeem et al., 1994; Johnson et al., 1996) but there remain few comprehensive and realistic empirical tests of their hypotheses in most habitats (Naeem et al., 1994, 1995; Tilman and Downing, 1994). Studies on terrestrial grassland, suggest

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that biodiversity begets stability (Tilman and Downing, 1994; Tilman, 1996; Tilman et al., 1996); diverse systems are more likely to contain some species which can thrive during a given environmental perturbation and thus compensate for the loss of species resulting from the disturbance. The diversity of species within an ecosystem will constrain the structure of food webs and may determine their stability. However, within food webs the number of trophic resources per species appears to be independent of the number of species in the community and the majority of links in food webs are weak. Those species which interact strongly tend to be the exception even though they may be the target of study (Pimm, 1982; Cohen and Briand, 1984). As such, the loss of a number of species may cause little short-term change in ecological function. However, as more species are removed, we approach a threshold where further removals will lead to a shift in ecosystem structure or function. The threshold may be reached rapidly if the species interact strongly. Thus the removal of a few species of urchin-eating fishes by fishers has led to the proliferation of urchins on Kenyan (McClanahan and Muthiga, 1988, 1989; McClanahan, 1990, 1995a) and Caribbean (Done,

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1992; Knowlton, 1992; Hughes, 1994) reefs and the subsequent bioerosion of reef habitats (see Section 2.3).

3.2.4. Intraspecijic Diversity While species extinctions are unlikely to result from the direct effects of fishing, losses of intraspecfic diversity are expected. Numerous fished stocks have collapsed, and many other stocks have been reduced to very low abundance before they recovered (Myers et al., 1995). Since large old fishes may be more heterozygous and some stock structures may have a genetic component (Smith et al., 1991; Carvahlo and Hauser, 1994), reduced intraspecific diversity would be expected following intensive exploitation. This reduction in intraspecific diversity has been of considerable concern in salmonid fisheries (Ryman et al., 1995) but has rarely been investigated for strictly marine species. For the contemporary marine fisheries manager the stock is an operational rather than genetic unit and stocks are regarded as breeding groups that have sufficient temporal and spatial integrity to be considered as self-perpetuating units, even if there is genetic exchange between them (Pawson and Jennings, 1996). However, there is a continued need to identify those population units which have defined genetic characteristics for the purposes of recording losses of intraspecific diversity and deciding how to protect it. Concern with species diversity has often masked concern for intraspecific diversity. This results from the poor communication between fisheries ecologists and geneticists and the requirements for many managers to operate on time scales that do not allow them to address genetic concerns (Carvahlo and Hauser, 1994; Pawson and Jennings, 1996). In addition, many fisheries were intensively exploited for decades or centuries before genetic studies were initiated and genetic changes will already have occurred. For example, the Newfoundland cod fishery was flourishing in the sixteenth century (Cushing, 1988b) and few areas of the North Sea were not trawled before 1900 (Figure 4). Indications as to the extent of changes which may have occurred when fisheries were first exploited are provided by Smith et al. (1991). They demonstrated that heavy exploitation of a previously unfished orange roughy stock over a 7-year period led to a significant decrease in heterozygosity. Their results suggested that losses of genetic diversity took place well before the stock would be considered endangered by those concerned with fish population dynamics. Fishing is selective with respect to a number of life history traits such as age and size at maturity, growth rate and reproductive output, which are at least partially heritable, and exploited populations would be expected to

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evolve in response to harvesting. Such evolution leads to changes in intraspecific diversity and is discussed in Section 3.4.2. 3.3. Size Structure

The size structure of the biota in a marine ecosystem follows regular patterns (Sheldon et al., 1972, 1973) and the primary linear relationship between size (as classes) and total normalized biomass (the width of the log size class interval divided by the log of total biomass in that interval) can be predicted by models of energy flow from prey to predators (Kerr, 1974; Platt and Denman, 1977, 1978; Dickie et al., 1987). Kerr (1974) suggested that deviations from such models could provide an index of the effects of exogenous stress on marine communities. It is reasonable to expect that fishing is one process which may cause the size distribution of biota within an ecosystem to differ from that which is predicted, since fishing will lead to the selective removal of larger fishes and their removal may, in turn, affect their predators or prey. Recent research has increasingly focused on the secondary scaling within the body size and biomass relationship since a series of domes corresponding to component trophic groups, such as phytoplankton, zooplankton or fishes, are effectively superimposed on the primary linear scaling (Thiebaux and Dickie, 1992, 1993; Duplisea and Kerr, 1995). Thiebaux and Dickie (1992, 1993) developed a model to describe the secondary scaling (domes) in biomass size spectra. This model allowed Duplisea and Kerr (1995) to examine temporal variation in the demersal fish assemblage of the Scotian shelf by assessing annual deviations from a theoretical curve. They suggested that interannual variations in the curvature of the fitted quadratic models reflected large changes in fisheries exploitation practices. Thus, the deviation of the actual curve from the fitted model in 1978 may have reflected the increased food resources available following a short-term decrease in exploitation rate (see Section 4) and an increase in the relative abundance of larger fishes within the population. Given the relatively weak curvature of the secondary dome over a small proportion of the total length range, aggregate logarithmic numbers of all fish species combined by length class decline linearly over the range of length classes which are fully sampled by many experimental gears such as those used on groundfish surveys. Comparison of such distributions in space and time in the North Sea and on the Georges Bank indicate that slope varies considerably between areas but much less through time within an area. Pope er al. (1988) considered that the slope of the line provided a broad indicator of exploitation regime, with more heavily exploited areas having a steeper decline with size. Indeed, the slopes of distributions on Georges Bank show

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more gradual declines in years after management regimes were imposed. Further study of North Sea data also suggested that the relationships between total biomass and body size were effectively linear over the selective size range considered (larger fishes sampled during groundfish surveys) and that the slopes and intercepts of these lines were linear functions of fishing intensity (Anon., 1996b; Greenstreet and Hall, 1996; Rice and Gislason, 1996; Figure 14). These effects were largely attributed to the selective removal of larger fishes by fishers (Pope et al., 1988; Anon., 1996b) and have also been demonstrated by Rijnsdorp et al. (1996b) who compared the length distributions of fishes caught in trawl surveys from 19061909 and 1990-1995. Analysis of size spectra as measures of fishing effects are at an early stage but if they are expanded to include all trophic groups they may provide a good basis for the study of fishing effects at the ecosystem level (see Section 5.3).

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3.4. Life History Traits

3.4.1. Changes in Multispecies Communities Fishes exhibit a range of life history tactics, which are presumably shaped by natural selection, to fit particular ecological demands (Stearns, 1976; Stearns and Crandall, 1984) and thus it is expected that fishing will affect fishes with different life history traits in different ways. Species with short lifespans and rapid population growth, which mature early and channel a large proportion of their resources into reproductive activities, are likely to respond rapidly to fishing, but, so long as fishing intensity and recruitment are in phase, they may be fished sustainably at younger ages and higher levels of mortality. Fisheries based on slower growing species, which mature later and at a larger size, are likely to be vulnerable to intensive exploitation despite naturally more stable population sizes which, in the unexploited state, are buffered by numerous age classes against the recruitment failure of individual cohorts (Adams, 1980; Beddington and Cooke, 1983; Roff, 1984; Kirkwood et al., 1994). A number of empirical studies have suggested that larger and late maturing species are more susceptible to exploitation (Brander, 1981; Bannerot et al., 1987; Trippel, 1995) and, since the maturation and growth parameters of fishes are closely interrelated (Am, 1959; Beverton, 1963, 1987, 1992b; Leggett and Carscadden, 1978; Jennings and Beverton, 1991; Charnov, 1993), a suite of other life history traits may also correlate with responses to exploitation. There are concerns, however, that these cross-species analyses treat species or stocks as independent data points when this may not be the case (Harvey and Pagel, 1991; Martins, 1996). For example, slow growing species such as skates and rays (Raja spp.) have often decreased in abundance following exploitation (Brander, 1981; Walker and Heessen, 1996), a response which has been attributed to their advanced ages at maturity and low fecundity. However, members of the genus Raja share other characteristics such as broad body shape and the laying of egg cases on the seabed which could also be responsible for their susceptibility. A comparison between members of this genus, or between stocks of one species, would eliminate the effect of other variables that they have in common (Harvey and Pagel, 1991; Martins, 1996)and this approach would provide a better basis for examining responses to exploitation (Jennings et al., 1998). Most fisheries are relatively unselective and many species experience high levels of mortality as by-catch even if they are not the primary targets of the fishery (Alverson et al., 1994). The susceptibility of late maturing and larger fishes to fishing suggests that small and early maturing species would increase in relative abundance in an intensively exploited multispecies fishery. However, while the life histories of smaller species may enable them to

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sustain higher instantaneous mortality rates than larger species, they may also suffer lower fishing mortality simply because they are less desirable and less accessible targets in a size-selective fishery. As a result, observed shifts in fish community structure result from the combined effects of differential fishing mortality and the variable susceptibility of species with different life histories. There have been changes in the structure of fish communities in most fished marine ecosystems which have been studied for a decade or more. Koslow et af. (1988) investigated the effects of fishing on the structure of Jamaican reef fish communities. The largest fish regularly caught in traps (the main fishing method), both piscivores such as large groupers and snappers, and herbivores, such as large parrotfishes, virtually disappeared from catches at all study sites. Deep-bodied fishes such as angelfishes (Pomacanthidae), surgeonfishes (Acanthuridae) and triggerfishes (Balistidae) also generally declined in heavily fished areas. There were few signs of change in the non-target communities apart from increases in squirrelfishes (Holocentridae) and non-balistid tetraodontiformes at some sites (Koslow et af., 1988). Pauly (1979) described the development of a demersal trawl fishery in the Gulf of Thailand. The larger fishes were selectively removed and the catches were increasingly dominated by flatfish, squids and crustaceans as the fishing effort continued to rise. Harris and Poiner (1991) also examined changes in a tropical demersal fish community between 1964 (prior to exploitation) and 1985-86 (after commercial trawling). The abundance of benthic associated species decreased and semi-pelagic species increased. Most changes occurred in the target and by-catch taxa and there was little evidence for indirect effects. Greenstreet and Hall (1996) compared community structure in the North Sea groundfish assemblage in the periods 1929-1953 and 1980-1993. Despite marked increases in fishing effort during this period there was little change in community structure of the non-target species and the changes in the target species were a result of the direct effects of fishing and recruitment changes that have been widely documented elsewhere (Ursin, 1982). These results were corroborated by Rijnsdorp et af. (1996b) who compared data from 19061909 and 1990-1995. However, even in the early part of this period the North Sea was already heavily fished (Cushing, 1988b; Figure 5 ) and many of the major changes in community structure may already have occurred (see Section 5.5). Studies of the relationships between fishing intensity and fish community structure in the tropics, where virtually unfished and yet accessible sites are often located, have indicated the rapidity with which the composition of target fish communities can change in response to fishing (Russ and Alcala, 1989; Russ, 1996b; Jennings and Polunin, 1996a). Low levels of fishing intensity, such as those which occur when a marine reserve is occasionally fished, are sufficient to cause dramatic changes in the trophic structure of the

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fish community (Figure 15). Similarly, biomass estimates for species at high trophic levels, in areas subject to a range of fishing intensities, demonstrated that marked decreases in biomass occur at relatively low fishing intensities (Figure 10) and that, at higher fishing intensities, the biomass is relatively stable. Jennings and Polunin (1996a) investigated the effects of fishing effort and catch rate on the structure and biomass of reef fish communities in seven Fijian reef fisheries. In the two least intensively fished grounds the estimated annual yields of invertebrate feeding and invertebrate feeding or piscivorous fishes did not exceed 4% of the biomass estimated by visual census. However, yields of these trophic groups approached 20% of biomass in the most intensively fished grounds where biomass was significantly lower. The fishing effects observed were primarily attributed to significant differences between the fish communities in the least intensively fished ground and all others. Thus at higher fishing intensities, the biomass of target species provided a poor index of relative fishing pressure. The results suggest that the annual removal of 5% of fish biomass may cause significant structural changes in target reef fish communities. Multivariate analysis of changes in the distribution of biomass between families in these fisheries (Figure 16) also demonstrated that the least intensively fished sites could readily be distinguished from others (the effect was significant; Jennings and Polunin (1996a)) but that at higher fishing intensities the structure did not vary consistently in response to fishing effort. The dramatic changes which occur when fish communities are fished for the first time suggest that fish communities in temperate systems have been in the “fished” state for a long time. Cousin Reserve (full protection)

Ste. Anne Reserve (some fishing occurs) invertebrate feeders

Figure 15 Composition (by weight) of the non-cryptic diurnally active reefassociated fish community in Seychelles’ marine reserves. Cousin Reserve is an effectively protected area whereas some fishing concessions are provided in Ste. Anne. (Data from Jennings et al, 1996b.)

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Figure 16 Multi-dimensional scaling ordination indicating the relationships between the structure of target fish communities in six Fijian fishing grounds subject to different fishing intensities (A represents the lowest fishing intensity and F the highest). Five randomly selected sites were censused within each fishing ground and each replicate is coded with the same letter. The greater the similarity between fish communities the closer the symbols. Stress = 0.16. (From Jennings and Polunin, I996a.)

3.4.2. Intraspecijic Changes in L$e Histories

Since fishing is selective with respect to a number of life history traits which are at least partially heritable, exploited populations would be expected to evolve in response to harvesting (Miller, 1957; Nelson and SoulC, 1987; Policansky, 1993). There are good precedents for such effects because a number of life history traits such as age and size at maturity, growth rate and reproductive output have been shown to have a genetic basis (Policansky, 1993) and selective predation on fishes by other fishes has

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been seen as a major cause of evolutionary change (Hixon, 1991). In practice it has proved difficult to detect heritable changes in exploited fish populations, since they are minor in comparison with phenotypic responses. For example, reduced population size in circumstances where food is limiting may result in large increases in growth and fecundity (Nikolshi, 1969; Borisov, 1978). Only in recent years have systematic attempts been made to elucidate the relative significance of phenotypic and genotypic responses to fishing (Stokes et al., 1993). Rijnsdorp (1993b) developed a model to examine the effects of size-selective fishing on the evolutionary fitness of North Sea plaice and to estimate the relative significance of genetic and phenotypic factors in explaining the changes in reproductive strategy that have occurred during the last century (Rijnsdorp et al., 1991a). Rijnsdorp (1993b) calculated the fitness profiles which resulted from allocating different proportions of the available resources to growth and reproduction. Given the current rates and selectivity of exploitation in the North Sea, the simulations suggested that fitness would improve when reproductive effort was increased. Rijnsdorp (1993b) discussed the results of the modelling exercise in relation to the changes in reproductive strategy that have been recorded since the late 1940s (Bannister, 1978; Rijnsdorp, 1989; Rijnsdorp et al., 1991a). He suggested that the observed increases in the fecundity of smaller females were in accordance with the changes suggested by the model and were unlikely to be a result of phenotypic plasticity, since improved conditions result in the faster somatic growth of plaice rather than increased size-specific fecundity (Rijnsdorp, 1990, 1993b). Rijnsdorp (1993a) also tried to determine the relative importance of genetic and phenotypic effects on changes in maturation and reproduction of North Sea plaice. Analysis of the phenotypic responses to an increase in juvenile growth suggested that only part of the decrease in length at maturity could be attributed to the observed increase in juvenile growth. Similarly, Rowell (1993) produced a model to assess the optimal age at maturation for North Sea cod if reproductive output was assessed under different mortality schedules. The model suggested that if there was a genetic component to the variation in age at maturity there could be a decrease in the average age at maturity as fishing pressure increases. Changes predicted by the model were fairly similar to those changes which were recorded in data from 1893, 1923 and the 1980s suggesting the importance of at least some genetic component. Since life history traits are quantitative, the evolutionary effects of harvesting can be investigated using quantitative genetics (but these approaches do not address the complex demographic features of fish populations where age structure is primarily determined by fluctuations in recruitment). Thus, Law and Rowell (1993) incorporated quantitative genetics into a model of population dynamics. Using this model they examined the role of selection

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for body length in North Sea cod and suggested a small selection response of around 1 cm at age 1 year after 40 years of exploitation. Law and Rowell (1993) emphasized that the potential errors in this analysis were large, but their assumptions were sound and there is little reason to indicate that they were not examining a real effect. In management terms their results indicate that evolutionary change will be very slow and of very small magnitude in comparison with phenotypic responses. This possibly explains why the evolutionary effects of exploitation have previously received so little attention. It is clear from the efforts of Law and Rowell (1993), Rijnsdorp (1993a) and others that it is extremely difficult to separate the effects of phenotypic and genetic changes in life history in wild populations of fishes subject to fluctuating environmental conditions, although there have been successes with salmon (Salmonidae) populations (Nelson and Soule, 1987). A number of investigators have also addressed these processes in experimental systems (Silliman, 1975; McKenzie et ul., 1983; Reznick et al., 1990). McKenzie et ul. (1983) attempted to separate the effects of genetic and environmental factors in influencing the growth and maturation of platies and Silliman (1975) has conducted selective harvesting experiments with the freshwater tilapia Oreochromis mossumbicus (Peters). In the population from which large fish were taken, Silliman (1975) demonstrated that growth rates of males decreased and the difference was heritable. Similarly, selective harvesting experiments with Duphnia mugnu populations have shown that the selective removal of larger individuals selects for slower growth (Edley and Law, 1988; Law and Grey, 1989; Law, 1991; Law and Rowell, 1993). Experimental studies have also shown how small evolutionary changes in life history traits will be masked by phenotypic responses. In guppies, Poeciliu reticulutu (Peters), which received a restricted total food supply (this may not consistently be the case in the natural environment), the increased resources available at lower population density led to increased size at maturity, decreased age at maturity, increased fecundity, increased frequency of reproduction and increased population growth rate (Reznick and Bryga, 1987; Reznick et al., 1990; Reznick, 1993). When mortality rates were artificially increased in order to simulate the effects of fishing, there was selection for early maturity at a smaller size and higher fecundity earlier in life. Reznick (1993) notes that the plastic response to the increased food supply was much greater than the evolved response to high mortality rates. Thus if the same relative effects were apparent in a wild fishery, the active selection for smaller body size would be masked by the increased growth at higher resource availability. Evolution is undoubtedly slow on time scales of interest to managers, but that does not reduce its long-term importance. The evolutionary-enlightened manager should adopt an evolutionarily stable optimal harvesting strategy that sacrifices yield in ecological time to maintain a larger adult prey size

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and higher yield in evolutionary time (Brown and Parman, 1993). A theoretical basis is now being developed to attempt this (Law and Grey, 1989; Blythe and Stokes, 1993; Stokes and Blythe, 1993), and if management moves from population to ecosystem-based approaches (see Section 6) it may be possible to institute long-term management in some areas. 3.4.3. Reproduction Size selective fishing can have a marked impact on the sex ratios of fish populations and artificially curtail reproductive lifespan. While the majority of temperate marine fishes are gonochoristic (individuals do not change sex), tropical families contain almost 50% hermaphrodite species in which most individuals function as both sexes - either simultaneously or sequentially. Protogynous hermaphroditism (female to male sex change) has been recorded in emperors, parrotfishes, wrasses (Labridae), porgies (Sparidae) and groupers and protandrous hermaphroditism (male to female) in porgies and snooks (Centropomidae) (Sadovy, 1996). Intensive and size-selective fishing of protogynous hermaphrodites can bias sex ratios. Thus Buxton (1993) demonstrated that the sex ratio of porgy populations in an exploited area was female-biased in comparison with a marine protected area. Similar changes have been reported in grouper populations (Bannerot et al., 1987) and Sadovy (1996) indicates that the proportion of male fish in gag grouper Mycteroperca microlepis (Goode and Bean) spawning aggregations in the Gulf of Mexico has fallen from 17% to 2% over a 10-year period. Thompson and Munro (1983) recorded male : female sex ratios of 1 : 0.72 and 1 : 0.85 for the groupers Epinephelus striatus and Mycteroperca venenosa (L.) on lightly exploited offshore Jamaican banks in comparison with ratios of 1 : 5.6 and 1 : 6.0 at the heavily fished Port Royal Cays. Many other grouper species mature as females and become males at sizes well in excess of those at first capture. However, information such as this only provides a very general indication of complex changes in the spawning aggregations. Shapiro et al. (1993) noted that the male :female sex ratio within a spawning aggregation of red hind grouper Epinephelus guttatus (L.) was 1 : 4.9, whereas it was 1 : 10.9 in inshore areas outside the aggregation period. Moreover, the fishes within the aggregation were larger than those inshore. Smith and Jamieson (1991) reported that size limits in the trap fishery for Dungeness crab in British Columbia effectively bar females from the catch because they rarely grow to a size in excess of the limit. Consequently, in heavily exploited areas, males greater than the size limit are rare and the sex ratio favours adult females. Based on an assessment of male mating activity and the sizes of mating pairs Smith and Jamieson (1991) proposed that mature females would have difficulties finding a sexual partner in intensively

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exploited fisheries and that the mating of fewer larger females would adversely affect egg production. Evolutionary theory suggests that the ratio of ma1e:female fishes in unfished populations will attain an equilibrium that maximizes reproductive output (Werren and Charnov, 1978; Charnov, 1982). A reduction in the number of male fish owing to fishing will disturb this equilibrium and may reduce the proportion of eggs that can be fertilized. There appear to be no direct tests to indicate when sperm production might be limiting but Shapiro et al. (1994b) concluded that sperm production in the male of the wrasse Thalassoma bifasciatum (Bloch) is sufficiently costly that males allocate sperm carefully amongst their frequent daily spawnings and for the males studied, the number of sperm released per spawn did not decline throughout the daily spawning period. Fishing will have more impact on reproductive output if sex change is determined by local demographic conditions rather than by absolute size or age. An increase in female bias with increasing fishing effort is consistent with the idea that protogynous sex change occurs at a characteristic size. In contrast, if sex change were regulated by a purely behavioural mechanism then this would compensate for the loss of males. In those species which have been studied sex ratios and relative size have been a cue for sex change (Ross, 1990), but these are not commercial species. Sadovy (1996) considered why biased sex ratios might persist if these mechanisms applied to fished species. For example, the removal of multiple males through fishing could produce delays in response when numerous females have to change sex or sex change may take longer to initiate or complete in progressively smaller females. If sex change is not closely tied to length, the decision by a female protogynous fish to change sex should be influenced by the sex ratio during spawning periods and/or by factors which vary directly with the spawning sex ratio such as relative rates of behavioural interaction with males or females outside spawning periods. Shapiro et al. (1994a) suggested that female red hind grouper were not able to make these decisions outside the spawning period as they did not encounter male fish. Fishing can affect the mating systems of fishes. During the rapid expansion of the tilefish fishery Lopholatilus chamaeleonticeps Goode and Bean from 1978 to 1982, population size fell by 50% or more and males spawned 2-2.5 years earlier and lOcm smaller (Grimes et al., 1988). These authors suggested that the change in population density may have allowed smaller and younger males to claim mating territories in the spawning grounds of the mid-Atlantic Bight. Harmelin et al. (1995) also reported that fishing pressure modified the social conditions in a Mediterranean wrasse Coris julis (L.) population and induced earlier sex change. Shifts in the size and age distributions of fish populations can also have profound influences on their reproductive output. The relative fecundity

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(number of eggs per unit of body mass) of fishes increases as they grow and thus a population of a given biomass will have a greater potential fecundity when composed of larger rather than smaller individuals (Bohnsack, 1990). In addition, when the reproductive lifespan of fishes is artificially curtailed by fishing, their potential reproductive output will not be realised. Evolutionary logic suggest that fishes lower their annual reproductive output only when it will ensure that their lifetime reproductive output is increased. In unfished populations, changes in reproductive output will be governed by changes in other life history traits or physical and biological characteristics of the environment (Jennings and Beverton, 1991). However, when a reduction in reproductive output is a direct consequence of fishing mortality, and therefore of no evolutionary benefit, the population will only maintain evolutionary fitness by rapid changes in reproductive strategy. These changes should involve a compensatory increase in reproductive output at a given size or age. There is some evidence for such changes in the spiny lobster where there was a 16% increase in size-specific fecundity after exploitation (De Martini et al., 1992). In North Sea plaice, younger fishes of a given length had a higher absolute fecundity (Horwood et al., 1986). The responses of individuals can have a marked effect on the response of the population to exploitation. Koslow et al. (1995) studied changes in the east Tasmanian orange roughy stock from 1987-1992 when the size of the stock fell by 50% following intensive fishing. Mean individual fecundity increased 20% in same period. The compensatory increase in individual fecundity, coupled with an apparent increase in the proportion of females spawning annually, limited the decline in egg production over this 5-year period to 15%. However, Leaman’s (1991) study of stocks of ocean perch Sebastes alutus Gilbert compared life history characteristics in lightly and heavily exploited stocks. At low stock abundance there was an increase in growth rate and decrease in age at maturity, but size-specific fecundity was lower in the faster growing fish. The reasons for this response are not clear. Rijnsdorp et al. (1991a) reported that the observed changes in growth, maturation and fecundity appeared to have compensated for about 25% of the losses in total egg production owing to increased exploitation for North Sea plaice, cod and sole. 3.5. Conclusions

Most of the marked effects of fishing on diversity and community structure occur at relatively low levels of fishing intensity. However, once systems enter a fished state, diversity and overall production may often remain relatively stable despite further changes in fishing intensity. As a result, studies conducted in systems where fishing preceded scientific investigations

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suggest that the systems are not markedly affected by fishing. Top-down (predator controlled) effects on prey fish diversity do not appear to be strong and the direct effects of fishing on fish community structure dominate those that result from the depletion of predators. On longer time scales (decades and centuries) than those that are important to contemporary managers (years) the role of species in a community is likely to change. Thus species that are functionally similar at the present time may perform different, and significant, roles when conditions change. As a result, local losses of diversity owing to fishing are likely to be significant in the longer term. In many systems that have undergone the most dramatic shifts it is notable that few species perform a given trophic function (see Section 4). Within a fished community, those species which mature late, grow slowly and have low reproductive output tend to be more susceptible to the direct effects of fishing than faster-growing species with early maturity. Fishing has selective effects on the life history strategies and genetic structure of many exploited stocks, but such effects have not been widely investigated because they are small in relation to short-term plastic responses in the life history. In the longer term, it is likely that the genetic effects of fishing will become increasingly marked and new approaches to management will be required if fishing is not to act as the overriding evolutionary force for many fish populations (see Section 6).

4. TROPHIC INTERACTIONS 4.1. Introduction

One of the most widely expressed concerns about the intensive and selective fishing activities of humans is that they will lead to imbalances in ecosystem function which have ramifications for non-target species. Thus fishers who capture small “forage fishes” such as sardines or pilchards Surdina spp., anchovies Engruulis spp., sandeels Ammodytes spp., capelin Mullotus villosus (Miiller) or Norway pout Trisopterus esmarki (Nilsson) will compete with other predators in the marine ecosystem. Industrial fisheries in the North Sea, for example, accounted for over half the total catch by the late 1980s (Anon., 1993a). Many forage fishes provide food for bird and marine mammal populations and in many cases the birds or mammals are species of considerable conservation concern (Bax, 1991). There is increasing pressure to manage marine ecosystems with a view to ensuring the well-being of birds and marine mammals rather than maximizing fish production for humans. Moreover, some fishery biologists have also expressed concern about the

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intensive fishing of forage fishes since these may provide food for more valuable fished species. Fishing has been cited as the cause of changes in community structure in ecosystems ranging from Antarctica (McKenna and Saila, 1991) to the tropics (Pauly, 1979; Harris and Poiner, 1991). Fishing causes direct changes in the structure of fish communities simply by reducing the abundance of target or by-catch species (see Section 3.4. l), but these reductions may lead to responses in non-target species through changes in competitive interactions and predator-prey relationships. The indirect effects of fishing on trophic interactions in marine ecosystems have become a major concern of the conservation movement (Anon., 1996a) and a good scientific basis for management decisions is essential. In addition, it is in this field that improved links between fish population biology and ecology would have particular benefits. The current debates amongst fisheries ecologists invoke comparison with debates about the relative roles of “top down” (predator) or “bottom up” (environmental and prey) control in ecosystems and the relative significance of “donor controlled” dynamics (in which victim populations influence enemy dynamics but enemies have no significant effect on victim populations) in food webs. Predation is recognized as a key structuring process in aquatic ecosystems (Kerfoot and Sih, 1987) but empirical evidence suggests it is wrong to assume that most predator-prey relationships are tightly coupled and that the removal or proliferation of one species which eats another will result in detectable changes in ecological processes. In particular, simplistic models of predator-prey interactions often take no account of prey switching, ontogenetic shifts in diet, cannibalism or the diversity of species in marine ecosystems and thus they often fail to provide valid predictions of changes in abundance. In section 4.2 we review the empirical evidence for and against the proliferation of prey species following the removal of their predators. This is of particular significance because it has been suggested that the deliberate removal of predatory fishes (or other top predators) may allow fishers to harvest more of their prey (Jones, 1982; Grigg et al., 1984; Munro and Williams, 1985). Birds and mammals may prey on marine fishes and humans compete with them as top predators. In Section 4.3 we consider whether the removal or proliferation of prey has significant impacts on predator populations, discussing the relative roles of fisheries and environmental change in determining the availability of prey and how these processes affect the abundance and life histories of predators. Fish population biology has a short history (see Section 1) and yet there have been marked changes in the species that dominate the biomass and yield from many of the most productive marine fisheries. There has been much debate as to whether these changes are natural fluctuations in marine

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ecosystems or are induced by fishing. In Section 4.4 we consider the significance of species replacements and whether these have been stimulated by fisheries which remove biomass and therefore reduce interactions between species which formerly competed with, or ate, one-another. In Section 4.5 we describe how discarded fish, and animals killed in the path of trawls, have benefited some species of scavenging birds and fishes and finally, in Section 4.6, we discuss whether the effects of fishing on trophic interactions are reversible. 4.2. Predator Removal

Decreases in the abundance of large fished predators, such as the bluefin tuna Thunnus thynnus (L.) or the blue whale Baluenopteru rnusculus (L.) have been amongst the most widespread and thoroughly documented direct effects of fishing. Given that many of these species were very abundant and that their prey consumption rates were high, there have been many attempts to correlate their decline with the proliferation of their prey (e.g. Laws, 1977; Tiews, 1978). On coral reefs, where the functional and species diversity of fishes is relatively high, the indirect effects of fishing on the abundance of unfished prey species appears to be minor. While Jones (1982), Grigg et al. (1984) and Munro and Williams (1985) have suggested that a fishing strategy that selectively targets piscivorous fishes may lead to increases in the production or biomass of their prey, empirical evidence suggests that any changes in the abundance of prey species following the capture of their predators may not be detectable. On the scale of kilometres to tens of kilometres Bohnsack (1982), Russ (1985), Jennings et al. (1995) and Jennings and Polunin (1997) 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 (Figure 17). This result is apparently contradicted by those studies that demonstrate increases in the abundance of small fast-growing species from low trophic levels in catches from intensively fished areas. However, the changes in catch composition which have been reported may result from fishing activities altering the fished habitat, from fishers shifting their attention to the only remaining resources or from fishers reducing the amount of catch discarded. These factors have rarely been treated explicitly in previous assessments but would not have affected the results of Jennings et al. (1995) and Jennings and Polunin (1997). They used fishery independent visual census methods to estimate abundance, compared habitats at sites subject to different fishing intensities, and worked in areas where destructive fishing practices were not in use. It is clear that the power to detect minor changes in prey populations at the large spatial scales

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Biomass of piscivores >30cm (g m-’) Figure 17 Relationships between the mean biomass (*95% CL, n = 7) of fishes > 30cm length and the mean biomass of their herbivorous and invertebrate feeding prey (& 95% CL, n = 7) < 15cm length which are not caught by any fishing method used in these fisheries. Data from 10 Fijian fishing grounds subject to different fishing

intensities (Jennings and Polunin, 1997).

examined in the studies of Bohnsack (1982), Russ (1985), Jennings et al. (1 995) and Jennings and Polunin (1 997) was relatively low, since there were large errors associated with fish biomass estimates from heterogeneous habitats. However, even if some prey release did follow the exploitation of piscivores, and could not be detected, such small increases in prey production would not have compensated for the long-term losses in yield that result from reducing piscivore populations (Jennings and Lock, 1996). It has been argued that an increase in the abundance of prey fishes on reefs is expected to follow a reduction in the abundance of their predators because such effects have been observed in freshwater ecosystems where more comprehensive and carefully controlled studies have been conducted. We consider, however, that the studies of predator-prey relationships on reefs are not fundamentally flawed and that the indirect effects of fishing are smaller and less predictable than in freshwater lakes. Within lake ecosystems the majority of biomass is aggregated into a few relatively distinct size classes that loosely correspond to phylogenetic groupings (Sprules et al.,

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1983). Phylogenetic diversity is considerably lower in freshwater lakes than in reef ecosystems, and the lake organisms within size groups tend to have a relatively limited range of life history traits and morphologies. Size has a key role in determining the potential predators or prey of the organisms within food webs (Peters, 1983; Cohen et al., 1993) and thus the organization of limnetic communities places strong constraints on community structure that do not occur in many marine environments (Neill, 1994). As a result, the abundance of piscivores in lakes often has a profound and predictable effect on the dynamics of their prey (Zaret and Paine, 1973; LeCren et al., 1977; Eggers et al., 1978; Marten, 1979; Bare1 et al., 1991; Wanink, 1991) and the most convincing evidence for the presence of “top down” control of aquatic ecosystems has been obtained from studies in lakes (Carpenter and Kitchell, 1984; Shapiro and Wright, 1984; Carpenter et al., 1985; Carpenter, 1988b; Carpenter and Leavitt, 1991). In many of the more diverse (species-rich) marine systems the biomass spectrum is extended and there is more variance in size within the main phylogenetic groupings (see Section 3.3). This is particularly apparent on reefs. Moreover, the phylogenetic groupings tend to contain more species, with a wider range of life history traits, behavioural differences and feeding strategies. Thus 8 to 53% of species in reef fish communities eat other fishes (Parrish et al., 1986; Hixon, 1991) and many of these are generalists (Hiatt and Strasburg, 1960; Vivien, 1973; Vivien and Peyrot-Clausade, 1974; Blaber et al., 1988). As a result, the overall effect of all piscivores on their prey can be substantial although the impact of any individual species, or small group of species, is minor. This situation has been termed “diffuse predation” by Hixon (1991) and differs markedly from the situation in lake ecosystems, and some marine systems such as the Barents Sea, where a few keystone species dominate the biomass within a trophic group. The trophic relationships between most species on reefs are poorly understood, despite many recent advances in elucidating the general structure and function of food webs (Pimm, 1982; Warren, 1990; Pimm et al., 1991; Paine, 1992; Martinez, 1993a,b, 1994; Warren, 1994; Leibold, 1996). On the basis of diet analysis (Hiatt and Strasburg, 1960; Vivien, 1973; Vivien and PeyrotClausade, 1974; Parrish et al., 1985, 1986; Blaber et al., 1988; Norris and Parrish, 1988) and stable isotope studies of trophic relationships (Jennings et al., 1997; N.V.C. Polunin, University of Newcastle, unpublished data) reef food webs appear to share more characteristics with many terrestrial webs than those in lakes. Thus multichannel omnivory is common and most links between predators and their prey are very weak (Paine, 1992; Polis and Strong, 1996). As such, and in the apparent absence of keystone piscivores, the intensive fishing of some piscivores has little indirect impact on reef fish community structure. Lawton (1989) has also made the intriguing suggestion that the more variable structure, greater connectance and more exten-

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sive omnivory shown by webs from constant environments may also be a product of more stringent dynamic constraints for persistence and stability in fluctuating environments. This may help to explain why there is more evidence for predator-prey coupling in some low diversity systems than on reefs. The lack of evidence for prey release following the exploitation of piscivores (Bohnsack, 1982; Russ, 1985; Jennings et al., 1995; Jennings and Polunin, 1997) contrasts with the strength of the evidence for the structuring role of predation at much smaller scales (metres and tens of metres). Thus Caley (1993), Hixon and Beets (1993) and Carr and Hixon (1995) have conducted elegant and replicated studies that demonstrated that experimental reductions in piscivore abundance lead to detectable decreases in the abundance and diversity of prey fishes. As with other processes which structure fish communities on reefs, the relative significance of predation depends upon the scale at which it is investigated. In temperate systems, there is some evidence for the predator-based control of prey species, but this is largely confined to the relationships between humans and their target fishes. Thus, when humans stop fishing, the biomass of target species tends to increase. There are numerous studies indicating that the population of target fishes decreases with increasing fishing effort and expands following the cessation of fishing (see Section 4.6). The strength of this relationship is likely to result from the conservative fishing strategies employed by humans who, in the majority of fisheries, are unwilling to be flexible in their aims and target a relatively small proportion of the total fish fauna (see Section 6.4). Most predatory fish, conversely, are very generalist feeders, often switching to invertebrate prey or cannibalism and eating many species of fishes at different stages in their life history. Predatory fishes rarely show the energetically inefficient human trait of pursuing a single species and are more likely to switch prey. The feeding strategies adopted by predatory fish are rarely used by human fishers unless they are entirely reliant on their fishery as a food source (Jennings and Polunin, 1996b). In temperate waters, studies of predator-prey relationships are often confounded by the direct effects of fishing since many prey species are found in the by-catch or may be the targets of directed reduction (industrial) fisheries. Thus models used to investigate predator-prey relationships need to incorporate fishing mortality and rapidly become very complex (Pope, 1979; Sparre, 1991). Pope and Macer (1996) investigated whether predation effects accounted for trends in the recruitment of the North Sea cod and whiting Merlangius merlangus (L.). These species primarily predate their own young, haddock Melanogrammus aeglefinus (L.),Norway pout and sandeel. Pope and Macer (1996) estimated mutual predation involving cod, haddock and whiting from 1921 to 1992, although for the period 1921 to 1974 they made the assumption that predation mortality rate was

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a function of predator biomass (and unaffected by prey abundance) rather than relying on the multispecies virtual population analysis that was developed in later years (Anon., 1994). The model of Pope and Macer (1996) also assumed that feeding behaviour is fixed and that there is no prey substitution. Despite these assumptions, their gross conclusion that predation effects did not account for trends in recruitment of cod and whiting species seems robust. In the case of cod, Pope and Macer (1996) tentatively suggested that overfishing had the overwhelming impact on population history, a result more generally supported by Myers et al. (1996) in their analysis of population trends in a range of cod stocks. The abundance of fishes that prey on sandeels has been markedly reduced in the northwest Atlantic and long-term changes in the abundance of sandeel populations with time were consistent with regulatory control by herring and mackerel (Fogarty et al., 1991; Figure 18). However, Fogarty et al. (1991) did not address the role of other predators or environmental change. Sherman et al. (198 1) showed that population explosions of sandeels in the northeast and west Atlantic coincided with depletions of larger tertiary

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predators such as Atlantic herring Clupea harengus L. and mackerel Scomber scombrus L., but there were other factors that could have explained the change. These debates about the role of predator or environmental control have proved remarkably difficult to test using conventional correlative approaches (see Section 5.2) but it is clear that the population structure of pelagic forage fishes is highly variable in the absence of changes in predator populations. Since the majority of fish biomass in most intensively fished systems is consumed by other fishes it is widely assumed that the removal of predators must affect the abundance of prey species. The models of Bax (1991) suggest that the majority of fish biomass in the intensively fished Benguela, Georges Bank, Balsfjorden, Eastern Bering Sea, North Sea and Norwegian/Bering Sea ecosystems was consumed by other fishes, but there is little evidence to suggest that changes in the abundance of piscivores would have cascading impacts on other parts of the system. High fish consumption rates do not consistently imply that predation is a structuring force when most species can act as predators and prey in the course of their life history and when adult predators are still capable of switching diet and feeding strategy in response to prey availability. There has been much interest in the effects of removing marine mammals such as whales on their prey populations (Beverton, 1985; Kock, 1994; Kock and Shimadzu, 1994). Baleen whale populations in the Antarctic were drastically reduced by whalers and it has been estimated that some 147 million t year-' less krill Euphausiu superba Dana production was consumed as a result (Laws, 1977). Increases in a number of species such as penguins (Sladen, 1964) and seals (Payne, 1977) were attributed to krill release following decline of the whale stock. There were also reports of increased growth, earlier maturity and increased reproductive rates of the remaining whales and other species (Laws, 1977; Payne, 1977; Bengston and Laws, 1985). However, a subsequent re-examination of much of this information suggested that many of these reports and the associated analyses were in error (Beddington and de la Mare, 1985; Cooke, 1985; Kock and Shimadzu, 1994) and that the trends were not convincing. Horwood (1987) summarizes much of this debate in relation to the population biology of the sei whale Balaenopteru borealis (Lesson). Contemporary studies of krill dynamics suggest that reduced krill consumption would have limited impacts on the dynamics or production of krill populations. Rather, changes in the breeding biology and growth rates of species which may compete with whales are the result of environmental changes and the effects of these changes on krill abundance. On the Antarctic Peninsula, for example, krill recruitment is influenced significantly by the extent and duration of sea-ice cover there during the previous season (Siege1 and Loeb, 1995), the environment beneath sea-ice apparently providing a favourable habitat for

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the development of larval krill (Daly, 1990). Krill abundance is thus closely linked to environmental conditions, and it now seems likely that poor recruitment following years of reduced ice cover will have significantly greater effects on krill abundance than mortality owing to consumption by cetaceans. At South Georgia, where much of the early whaling activity was centred, krill abundance fluctuates by a factor of 20 between years (Brierley et al., 1997), and in years of low abundance krill predators suffer catastrophic breeding failures (Croxall et al., 1988; Brierley and Watkins, 1996). Such fluctuations in krill abundance were evident at South Georgia before the whaling industry began to affect whale stocks there (Hardy and Gunther, 1935; Priddle et al., 1988), and at that time slumps in krill abundance also coincided with seasons characterized by warmer sea-surface temperatures (Mackintosh, 1972). 4.3. Prey Removal

4.3.1. Fishes

In a number of temperate ecosystems, such as the heavily fished Pacific and Atlantic upwellings and coastal shelves, relatively few planktivorous species are responsible for linking zooplankton production to fish production. Planktivores such as sandeels, sardines, horse mackerel Trachurus spp., capelin and anchovy are often targeted by reduction (fish meal) fisheries which frequently remove 20% or more of their total biomass on an annual basis (Anon., 1993a). These planktivores are an important component of the diet of many piscivorous fishes such as cod and hake. Birds and mammals also prey extensively on planktivorous fishes and humans compete with them as top predators. Walters et al. (1986) studied interaction between Pacific cod Gadus macrocephalus Tilesius (piscivore) and Pacific herring Clupea pallasi Valenciennes (planktivore) in the Hecate Strait, British Columbia. Estimates of herring stock size and cod landings per unit effort suggested that years of high herring abundance were followed by good cod catches which were in turn followed by good herring catches. Walters et al. (1986) modelled the interactions between herring and cod using mathematical descriptions of species interactions developed by Andersen and Ursin (1977) and applied their model to a time series of cod and herring abundance data for the period 1955-1975. Walters et al. (1986) made two assumptions when applying an equation in this form, notably that predation loss is independent of loss owing to intraspecific interactions such as food competition, and that the predators response to prey density is linear such that satiation would not lead to decreases in consumption rate when prey

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density is high. The model indicated that increases in cod abundance have a negative effect on juvenile herring survival while increases in herring abundance have a positive effect on cod recruitment. By accounting for the effects of cod predation, much of the general change in herring recruitment rates could be explained. This was consistent with an interaction, but could also have been consistent with environmental forcing which could not be demonstrated by the individual environmental indices which they tested. It is also possible that the recruitment variations were not fully explained because they failed to identify the delay in the timing of herring consumption and how this translates into cod recruitment and survival. As with many of these studies, several interpretations are consistent with the available data. Atlantic herring were the main prey fish of Atlantic cod and marine mammals in the Norwegian Sea-Barents Sea ecosystem. The massive reduction in herring biomass following intensive fishing and falling sea temperatures in the 1960s was thought to reduce herring predation on larval and juvenile capelin which, coupled with lower sea temperatures, led to stronger capelin recruitment (Blindheim and Skjoldal, 1993). In the early 1980s, however, sea temperature rose again (Dickson et al., 1988) and cod recruitment increased. The warmer conditions did not favour capelin and their recruitment failed in 1984 and 1985. This poor recruitment, coupled with the effects of intensive fishing, led to the collapse of the capelin stock. Since herring biomass was already very low, those species which fed on pelagic fishes had few alternative prey, and cod, seals and seabirds began to starve (Dragesund and Gjosaeter, 1988; Hamre, 1988; Figure 19). The cod stock rapidly reduced populations of prey fish and started to cannibalize its own young (Mehl, 1987). Mehl and Sunnand (1991) reported that cod weight-atage (by year class) had decreased between 20 and 70% from 1984-88 and individual food consumption had fallen by 40-70%. During this period the cod stock had increased in biomass from 1 to 1.5 million t because recruitment was good. Capelin formed an increasingly small proportion of the cod diet in this period and total prey consumption by cod generally declined. There seemed to be little evidence for prey switching as capelin became a less significant proportion of the diet (Mehl, 1986, 1987), possibly because an abundant alternative prey source was not available. Magnhson and Palsson (1991) have also suggested that cod growth is reduced during periods of low capelin biomass in Icelandic waters. In the northwest Atlantic, capelin are also an important component of cod diets but there is less evidence for the tight coupling of the cod and capelin predator-prey relationship. Capelin abundance declined rapidly in the northwest Atlantic during the late 1970s, and although this was commonly assumed to be a result of overfishing it may have been attributable to environmental factors (Leggett et al., 1984). Lilly (1991) showed that the

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quantity of capelin in the stomachs of cod varied in accordance with the abundance of capelin in the northwest Atlantic but that during periods of low capelin abundance the cod did not appear to switch to other diets. This suggested that a severe reduction in capelin biomass could influence cod growth. However, there is no empirical evidence for such an effect: Millar et al. (1990) found no relationship between capelin abundance and cod growth and Akenhead et al. (1982) found no relationship between cod growth and capelin or cod biomass and water temperature in northeast Newfoundland and Labrador. The marked decreases in the population sizes of large marine mammals, tunas and sharks will have reduced the numbers of these species that die from natural processes and sink to the sea floor. There are few studies of the role of nekton falls in the deep ocean (Dayton and Hessler, 1972) but one study in the Santa Catalina basin suggests that approximately 11% of benthic community respiratory requirements are met by nekton carcasses reaching the basin floor (Smith, 1985). Given the effects of fishery discards on the distribution, abundance and behaviour of scavengers (see Section 4.5) it might be expected that the reduced frequency of nekton falls in the deep ocean may have significant effects on the biota. 4.3.2. Birds The reproductive success of many seabirds is positively related to the availability of their favoured prey fishes and prolonged periods of low prey biomass may lead to significant decreases in seabird population size (Furness, 1982; Powers and Brown, 1987). For example, the reproductive success of the brown pelican and other seabirds in California (Anderson et al., 1982; MacCall, 1986; Ainley et al., 1995), various seabirds feeding in the upwellings on the South American (Duffy, 1983) and South African (Burger and Cooper, 1984) coasts, kittiwakes Rissa tridactyla (L.) (Murphy et al., 1991) northern gannets Sula bassana (L.) and guillemots Uria spp. in Alaska (Montevecchi and Myers, 1995, 1996), Atlantic puffins Fratercula arctica L. in Norway (Barrett et al., 1987), guillemots in the Barents Sea (Vader et al., 1990), Atlantic puffins in the north-west Atlantic (Brown and Nettleship, 1984), shag Phalacrocorax aristotelis (L.) (Harris and Wanless, 1991; Aebischer and Wanless, 1992) kittiwake (Wanless and Harris, 1992; Hamer et al., 1993) great skua Stercorarius skua (Briinnich) (Hamer et al., 1991; Klomp and Furness, 1992) Arctic terns Sterna paradisaea Pontoppidan (Monaghan et al., 1992), Arctic skua Stercorarius parasiticus (L.) (Phillips et al., 1996) and other species (Bailey, 1991; Bailey et al., 1991) in Scotland is related to the availability of prey fishes. An example of the relationship between the breeding success of Arctic terns and food supply is

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presented in Figure 20. The continued absence of prey in consecutive years may lead to reductions in bird population size. Thus there are strong correlations between guano production and sardine abundance off southwest Africa (Crawford and Shelton, 1978; Siegfried and Crawford, 1978; Newman and Crawford, 1980; Crawford et al., 1987, 1992; Shelton, 1992) and Peruvian anchovy and guano production off western South America (Duffy, 1983; Muck, 1989). Indeed such tight coupling has led some commentators to suggest that seabirds could be used as crude indicators of the health of fish populations (Monaghan et al., 1989; Furness and Barrett, 1991; Anon., 1996c; Anker-Nilssen et al., 1997; Furness and Camphuysen, 1997). Given the unequivocal evidence for the effects of fish availability on seabird populations, the evidence for the indirect effects of fishing on seabird populations is dependent on the evidence for the effects of fishing on pelagic fish populations. There is no doubt that populations of short-lived pelagic species are highly variable in space and time, because the total stock size is not buffered against individual recruitment failures when only one or two other age classes contribute to the total stock biomass and because the range of these schooling pelagic species is often closely linked to their 3500 0

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abundance. Much of the population variability can be related to changes in the environment, for example during periods of lower productivity associated with the cessation of intense upwelling or as a result of the mismatch of reproduction and the timing of secondary production (Wright and Bailey, 1996). However, fishing undoubtedly increases the susceptibility of pelagic fish populations to environmental effects by further reducing the age structure of the population and increasing dependency on the success of individual recruitment events. Thus fishing has been indirectly responsible for some of the reduction in seabird populations in the upwelling ecosystems off Peru and southwest Africa (Crawford and Shelton, 1978; Siegfried and Crawford, 1978; Newman and Crawford, 1980; Duffy, 1983; Crawford et al., 1987, 1992; Shelton, 1992). Some collapses of capelin stocks, for example in the Barents Sea, suggest that overfishing has contributed to reductions in seabird populations. Anker-Nilssen et al. (1997) reported that guillemot populations on the Norwegian coast decreased by 80% in 1985-1987 following the collapse of the Barents Sea capelin stock (see Section 4.3.1 and Figure 19). Piatt (1987) and others investigated the effects of changes in capelin abundance in more detail. The reproductive success of seabirds can be influenced by capelin abundance and capelin had to reach a threshold density if birds were to feed effectively (Piatt, 1990). Above the threshold, fluctuations in capelin density were compensated for by behavioural responses (Burger and Piatt, 1990), but below the threshold density, no amount of foraging would compensate for the lack of prey. Interestingly, the analysis suggested that there was little competition for prey between birds, whales and cod in this ecosystem because the different feeding strategies they adopted led to their spatial separation. The key factor determining seabird reproductive success was whether the total size of suitable prey stocks, as determined by fishing and environmental effects, exceeded threshold levels (John Piatt, Alaska Biological Science Center, pers. comm.). The differential responses of birds and mammals with different feeding strategies to changes in prey abundance were effectively demonstrated by a comparative study in Prince William Sound, Alaska. Hayes and Kuletz (1997) had demonstrated that the guillemot population had declined in abundance following reductions in the Pacific sandeel Ammodytes hexapterus Pallas stock, but in a wider analysis of the same ecosystem Kuletz et al. (1997) indicated that those species such as marbled murrelets Brachyramphus marmoratus (Gmel.) Brandt, pigeon guillemots Cepphus columba Pallas, tufted puffins Lunda cirrhata Pallas, arctic terns, harbour seals Phoca vitulina (Gray) and minke whales Balaenoptera acutorostrata L. which fed on schooling forage fishes had all declined in abundance while harlequin ducks Histrionicus histrionicus (L.), goldeneyes Bucephala clangula (L.) and sea-otters Enhydra lutris (Fleming) which prey on benthic invert-

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ebrates have remained stable. A comparison of the diets of piscivorous birds from 1972-1981 and 1989-1995 demonstrates a shift in prey from Pacific sandeels to gadoids: gadoids are energy-poor species which can only be captured with more foraging effort. The rapid declines in reproductive success of seabirds inhabiting the Shetland Islands in the northern North Sea during the 1980s coincided with the development of an industrial fishery for sandeels. For example, the territorial attendance, chick growth rate and breeding success of skua was lowest in the late 1980s when recruitment of sandeels was poor (Phillips et al., 1996) and Klomp and Furness (1992) demonstrated that the number of breeding territories apparently occupied by great skua had fallen, foraging effort had increased and that there was a 75% fall in breeding success. The northern coast of Scotland and associated islands are used by the largest seabird colonies feeding in the North Sea and there was considerable concern that the industrial fishing of sandeels in other parts of the North Sea was responsible for these declines. Sandeels were known to be a key component of the diet for many seabirds in this region, since other small schooling prey are scarce (Wright, 1996). Despite strong evidence for the indirect effects of fishing on seabird populations in some other ecosystems, changes in the sandeel stocks around Shetland appeared to reflect variability in the pre-recruit immigration and survival of sandeels rather than changes in the total stock size (Monaghan, 1992; Wright, 1996). In addition, prey availability to seabirds was found to be influenced by the size of young-of-the-year sandeels and the time when they became available to seabirds at chick rearing. It was notable that many of the bird species which were most reliant on young-of-the-year sandeels suffered the greatest declines in reproductive success (Wright, 1996). These declines coincided with the development of the industrial fishery but given that young-of-the-year sandeels are first subject to fishing mortality towards the end of the chick rearing period, any fishing effects in the northern North Sea depended on the fact that recruitment was influenced by reductions in stock size. There is no evidence for a relationship between recruitment and parent stock in sandeels or most other small pelagic fishes, and, during a year of good recruitment in 1991, the reproductive success of the seabirds improved as some of those bird feeding grounds which previously held low stocks of sandeels were recolonized (Wright, 1996). In conclusion, the available evidence suggested that the sandeel reduction fishery was unlikely to have been responsible for changes in bird populations in this region. On many coasts, it is widely recognized that seabirds and fishers exploit the same shellfish stocks. As a result, concern had been expressed that the overexploitation of shellfish was affecting the quantity of food available for birds such as the eider duck Somateria mollissima (L.) and oystercatcher Haematopus ostralegus L. and that fishers were disturbing feeding birds.

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Conversely, fishers have complained that the birds were eating large quantities of shellfish which they could otherwise harvest (Anon., 1996~).The most comprehensive studies of the relationships between birds and shellfisheries have been conducted on the Exe estuary in southern England and the Loughor estuary (Burry Inlet) in south Wales (Goss-Custard et al., 1995a,b; Stillman et af., 1996). Stillman et al. (1996) developed models which determine the effects of mussel harvesting on overwintering oystercatcher populations and validated the models using empirical data. The models use foraging decisions, interactions and the energetic requirements of individual birds as a basis for predicting the population level consequences of fishing on shellfish beds (Figure 21). When fishers are present on the mussel beds, the oystercatchers have to relocate in order to feed, and each bird has to decide between feeding in areas with many mussels where many competitors will be encountered, or feeding in areas with few mussels where few competitors will be encountered. As a result, competitively dominant birds will tend to feed in areas where mussels are abundant and subdominants in other areas. However, the decision made by each bird depends on the decisions already made by other birds. The model accounts for such decisions and for the

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metabolism (hence food requirements) of the birds, tidal cycles (which affect the availability of mussels), gut processing rates and the availability of other feeding grounds. Gut processing rates of overwintering oystercatchers on the Exe estuary are such that birds that are regularly disturbed by fishing activities (and also to a significant extent by people walking themselves and their dogs) may not be able to meet their energetic demands and may switch to feeding on earthworms in fields surrounding the estuary. Stillman et af. (1996) initially ran the model using the known population of overwintering oystercatchers and the density of mussels on the beds. Every day, each bird would adjust its feeding strategy with respect to all other birds in the population. The oystercatchers were required to meet their energetic demands on a daily basis, and if they could not meet these demands by feeding in the estuary or surrounding fields they would die. A comparison of the results with empirical data (counts of the birds feeding on adjacent fields) indicated that it provided a good description of population processes and the model was run to investigate the effects of different mussel fishing strategies on the population. Three fishing strategies were compared, low-tide thinning, low-tide stripping and high-tide stripping. Low-tide thinning involves the removal of adult mussels at low-tide and leaves bed size unchanged, low-tide stripping involves the removal of all mussels for sorting after collection and can potentially reduce bed area to zero, and high-tide stripping is equivalent to low-tide stripping, but bed area can only be reduced to 25% of the original size because all the beds are not accessible at high tide. With all fishing strategies, the fishers’ collection rate depended upon the density of mussels, so this affects the time they stay on the beds and disturb the birds. Of the three fishing methods considered, low-tide stripping had the most damaging effects on bird populations. The Stillman et al. (1996) model has practical management implications because it may alert managers to the impending starvation of overwintering oystercatcher populations by the numbers of birds feeding upshore or leaving to feed on fields adjacent to the estuary. When such changes in behaviour are observed, fisheries could be regulated in order to minimize their effects on bird populations. This form of adaptive management is already in use on wildfowl reserves where shooting is banned during cold weather. The model has also been parameterized for the Burry Inlet where oystercatchers feed on cockle beds which are exploited by fishers. Within the Burry Inlet, there was no indication of the birds feeding in fields and, on average, they only spent 40% of the day feeding as opposed to 85% in the Exe. These preliminary analyses suggest that fishing is likely to have more effects on overwintering oystercatcher populations on the Exe estuary (Figure 22). At present, the model does not account for redistribution of birds between estuaries, and has not been parameterized for other larger sites where the shellfish populations are

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Fishing effort Figure 22 Hypothetical relationship between fishing effort in shell-fisheries and the mortality or emigration of shorebirds. Arrows indicate the status of the relationship between fishing effort and mortality or emigration in two British estuaries. (Based on a model developed by Stillman et al., 1996.)

far less stable and fishing activities are highly sporadic in space and time. However, both the model and associated empirical data suggest that fishing activities have the potential to impact populations of shellfish-feeding birds in confined estuaries.

4.3.3. Mammals Many marine mammals forage on fished species (Harwood and Croxall, 1988). Within some ecosystems, the collapse of prey stocks has resulted in changes in the diet or behaviour of marine mammals and, in some cases, has led to starvation. For example, the reproductive success and behaviour of fur seals Arctocephalus australis australis Zimmerman in the Peruvian upwelling ecosystem is correlated with Peruvian anchovy biomass (Majluf, 1989) and the collapse of the shrimp and capelin fishery in the western Gulf of Alaska and eastern Aleutian islands from 1976-1981 coincided with the decline of harbour seal and the northern sealion Eumetiopias jubatus (Schreber) populations (Hansen, 1997). The collapse of the Norwegian capelin stock not only had implications for cod and seabirds (see Sections 4.3.1 and 4.3.2), but also for grey seals

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Phoca vitulina vitulina L. (Hamre, 1988, 1991, 1994; Figure 19). When the capelin stock collapsed, and given that herring biomass was already low, there were no abundant and alternative sources of prey for the seals (Dragesund and Gjosaeter, 1988; Hamre, 1988). Some seals died of starvation, but others made an atypical migration out of the Barents Sea, presumably in search of food (Haug et al., 1991). During the course of this migration a large proportion of the seal population died of starvation or was trapped in fishing nets (Haug et al., 1991). The collapse of the capelin fishery may also have led to a shift in the distribution of humpback whales during the late 1980s (Christensen, 1990). The redistribution of marine mammals is frequently a response to shifts in the density of their prey. For example, changes in the species of marine mammal which were recorded in the Gulf of Maine appeared to be associated with the proliferation of sandeels (Payne et al., 1990). Thus the humpback Megaptera novaeangliae (Borowski) and fin Balaenoptera physalus (L.) whales which fed on sandeels entered the area whereas those right Balaena glacialis (Muller) and sei whales which formerly fed on copepods, a prey-species of the sandeel, tended to leave (Payne et al., 1990). The intensive exploitation of many fishes which appear in the diet of the common porpoise Phocoena phocoena (L.) in the northeast Atlantic has led to concern that fishing may have indirect effects on its population status (review: Hutchinson, 1996). The coincidence of porpoise distributions with those of Atlantic herring had led to the conclusion that the demise of the North Sea herring stocks was responsible for decreases in the abundance of porpoises. However, harbour porpoises appear to feed on both demersal and pelagic fishes, and a number of gadoids increased in abundance as the herring declined (Figures 12 and 13). Moreover, in some areas, porpoise remained abundant following the collapse of herring stocks. Hutchinson (1996) concluded that the incidental catch of porpoises in nets was more likely to affect porpoise populations than any impact of fishing on their prey species. The porpoises in the northeast Atlantic, in common with other marine mammal species which have relatively varied diets and which feed in ecosystems where the choice of prey is varied, are not dramatically affected by the overfishing of some of their prey species. Grey seals feeding in the northern North Sea also show considerable plasticity in their diets and feed on a wide range of pelagic and demersal species as they become seasonally available (Hammond et al., 1994a,b). In general, only those marine mammal populations with a limited choice of prey (such as the seals feeding on capelin and herring in the Barents Sea) will be affected by fluctuations in the abundance of these prey. If it can be demonstrated that fishing has caused the declines in prey populations then it

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is reasonable to infer that fishing is indirectly responsible for the reduced abundance of marine mammals. 4.4. Species Replacement

There have been dramatic shifts in the composition of fish catches from many locations (reviews: Kawasaki et al., 1991; Sherman et al., 1993) and it is often suggested that the depletion of one species by fishing has allowed another species to proliferate as a result of reduced competition or predation. We refer to species replacements if they involve large absolute changes in the biomass of species within the system and if the changes appear to be compensatory: i.e. one species proliferates as another declines. True species replacements should not be confused with the changes in catch composition which occur as a fishery is developed and ultimately overexploited. These changes usually reflect the behaviour of fishers who, in many circumstances, will tend to target smaller fishes from lower trophic groups as the primary target species are overfished (see Section 3.4). Thus the increased catches of smaller species are an indication of the dynamic behaviour of fishers who are striving to maintain yield and do not necessarily indicate that the biomass of smaller species is increasing. Some of the most dramatic shifts in fish community structure have occurred in the lughly productive fisheries for clupeoids in upwelling ecosystems. For example, following the collapse of the Californian sardine Sardinops sagax (Jenyns) fishery in the California current, the biomass of northern anchovy Engraulis mordax Girard another fast-growing planktivorous species, began to increase (MacCall, 1986). The northern anchovy is commonly regarded as a competitor with the sardine because they have similar feeding habits and life history traits. The collapse of the sardine stocks was purported to have released additional food for the anchovy and had led to reduced predation on anchovy larvae. As a result, the anchovy had proliferated. While there was little doubt that the effects of intensive fishing and environmental fluctuation had led to the collapse of the sardine stock (MacCall, 1986) the evidence for competitive replacement was not convincing. In particular, the theory of competitive interaction was also not supported by an examination of long-term changes in the relative abundance of fish scales within sediment cores, since there were several periods when sardine and anchovy were both abundant (Soutar and Isaacs, 1974; Baumgartner et al., 1992; Figure 23). Moreover, the increase in anchovy abundance lagged the decline of sardine by almost a decade (MacCall, 1986) and other arguments for competitive replacement which are based on changes in growth rates were dubious, since they failed to allow for environmental changes (MacCall, 1986). In general, there is little convincing

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evidence to suggest that intensive fishing of the sardine led to the proliferation in anchovy stocks, and in common with the changes in many upwelling systems it seems that all species do well in times of high productivity and low fishing effort (De Vries and Pearcy, 1983; Shackelton, 1987). Since the early 1980s the stock of Pacific sardine has started to recover and brown pelicans (see Section 4.3.2) have switched from feeding upon anchovies to feeding upon sardines. Anchovies have also replaced pilchard in catches from the upwellings on the southwest coast of Africa. Once again, there is little convincing evidence to suggest that fishing was responsible for the shift which was observed. The examination of anchovy and pilchard scales in sediment cores (Shackelton, 1987) suggested that there was no historical competition between the species. Moreover, an attempt to shift the system back to its former state by overfishing the anchovy in the hope that the depleted but more valuable

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pilchard would return was not successful (Crawford et al., 1987) although it was suggested that subsequent increases in other planktivores may have followed the reduction in anchovy and pilchard biomass (Crawford et al., 1987). In the pelagic fisheries of the Peruvian upwelling system the anchovy was partially replaced by sardine following the collapse of the anchovy stock in the early 1970s. However, the change has not been truly compensatory and the combined biomass of these small planktonic fishes has decreased dramatically (Figure 24). Anchovy and sardine were thought to interact strongly since anchovies and sardines prey on one another’s eggs (Santander et al., 1983). However, this appears to be another situation in which predation is an important source of energy transfer but not a structuring force. In upwelling systems it seems that the total primary production, which is driven by upwelling, is more likely to be responsible for longterm changes in fish populations. Such effects are suggested by the relative abundance of anchovy and pilchard scales in sediment cores from Peru (De Vries and Pearcy, 1983).

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There have been dramatic changes in the composition of pelagic fish populations in the north Atlantic. The rapid decline in herring on Georges Bank and in the Gulf of St. Lawrence in the late 1960s was coincident with the strong recruitment to the Atlantic mackerel stocks. Lett and Kohler (1976) suggested that competition between Atlantic herring and mackerel explained changes in herring growth and recruitment, but their results can reasonably be interpreted in other ways (Shelton, 1992). It was also notable that the relatively strong 1970 year class of herring occurred at about the peak of mackerel biomass. By the end of the 1970s both species were at a low level probably owing to high fishing pressure in conjunction with environmental factors (Grosslein et af., 1980). Similarly, the change from an Atlantic herring to a pilchard-dominated pelagic community in the English Channel took place without evidence of a fishing effect (Cushing, 1961; Southward, 1980). Russell et af. (1971), Steele (1974), Southward (1980) and Southward et af. (1988) have argued convincingly that climatic changes were largely responsible for these changes, though there is a suggestion that unrecorded heavy trawling on the spawning fish may have hastened the collapse of the herring stock (Southward, 1963). In the North Sea, changes in the abundance of herring and mackerel have been implicated with the shift to a gadoid-dominated community in the 1970s. Following the reduction in North Sea herring and Atlantic mackerel stocks in the late 1960s and early 1970s, and the subsequent increases in recruitment of many demersal species, it was suggested that the removal of herring and mackerel had led to increases in copepod populations and increases in the food supply for larvae of demersal fish (Jones, 1973). Pelagic species such as mackerel and herring may interact with pelagic stages of gadoids (Hislop, 1996) because they prey on larvae or juveniles and compete for food with gadoid larvae. However, the first exceptional year class of haddock occurred in 1962 when herring biomass was still similar to that in the previous decade and mackerel biomass was also high (Hislop, 1996). The evidence for reductions in the biomass of herring and mackerel providing a window of opportunity for the gadoid outburst is not convincing and the improved recruitment of gadoids appears to have resulted from environmental change. Ursin (1982) discussed the effects of the exceptionally abundant year classes of gadoids on the structure of North Sea fish communities and noted that they may persist for many years. He suggested that this provided evidence for considerable “spare capacity” in the ecosystem. Thus the strong year classes of haddock, whiting and Norway pout in the North Sea in 1967 would have led to increases in the biomass of the gadoid stock from 0.3 to around 1.5 million t by 1968-1969 in a North Sea fish community with a total biomass of around 8-9 million t. The haddock alone would have effectively doubled the stock of demersal feeding fishes for many years. When the

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changes in gadoid biomass were fed into the holistic North Sea ecosystem model of Andersen and Ursin (1977) the large haddock year class soon grazed down the benthos, and this would have been expected to lead to increased food competition. However, there was no empirical evidence to suggest that the growth of demersal stage haddock was reduced by competitive interaction as it was little different from that in preceding years (Ursin, 1982). Furthermore, the zooplankton bloom which was predicted to follow the massive reduction in pelagic stocks in 1964-1967 did not appear to occur in nature (Ursin, 1982). It appeared that the simple linked predator-prey relationships used in the model did not mimic the more complex processes in the natural environment where there was little evidence for top-down control. Interestingly, the addition of other species to the model started to increase the stability of the model system, in accordance with empirical observation. The effects of competition and predation are also assumed to have determined the interactions between fish communities on Georges Bank. From the early 1960s to the late 1970s there were marked declines in the abundance of both pelagic and demersal fishes on the Georges Bank. The declines in pelagic species were most marked: in the late 1960s the biomass of the principal pelagics was more than double that of other finfish and squid, whereas by 1975 these pelagic and demersal communities were approximately equal in size. Heavy fishing was undoubtedly the main cause of the decline in biomass of the demersal community, but there were also questions about the effects of secondary changes due to species interactions (Grosslein et al., 1980; Overholtz and Tyler, 1985). By the mid-1970s Grosslein et al. (1980) noted that the demersal (groundfish) community was shifting back to its former composition, with haddock and spiny dogfish (or spurdog) Squalus acanthius L. recovering relatively quickly. However, the increase in haddock recruitment in the early 1970s had little long-term impact as the stock was subsequently reduced by intensive fishing (Rothschild, 1991; Figure 25). In the period from 1963 to 1986, as the total biomass of the demersal community fell (Gabriel, 1992; Mayo et al., 1992), the proportion of spiny dogfish and skates increased from 24% to 74% (Sherman, 1991; Figure 26). The proliferation of spiny dogfish appeared to be associated with regulations which prevented distant-water fleets from fishing on Georges Bank. These fleets had formerly fished spiny dogfish and winter skate which were used as food fishes in Europe and could be converted into fish meal. However, once the “distant-water” fisheries closed, the local coastal fleets selectively targeted valuable food fishes such as haddock, and while their biomass fell, the biomass of dogfish and skates increased (Rothschild, 1991). It is notable that the biomass of spurdog in the North Sea has decreased dramatically as fishing effort has increased

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Figure 25 Mean annual catch rates of haddock (open circles), spiny dogfish (filled squares) and winter skate (open triangles) during groundfish surveys on the Georges Bank. (Redrawn from Rothschild, 1991.)

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Figure 26 Percentage composition (by weight) of the Georges Bank demersal fish community in 1963 and 1986. (Redrawn from Sherman, 1991.)

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(Greenstreet and Hall, 1996), since it continues to be a valuable species in this region (M. Vince, pers. comm.). Total fish consumption resulting from predation is considerably higher than fish production in the Georges Bank system (Sissenwine et al., 1984; Sissenwine, 1986) and following changes in the fish community during the late 1970s and early 1980s (Figures 25 and 26), Overholtz et al. (1991) demonstrated that fish consumption by spiny dogfish was greater than that attributable to all other fishes, birds and marine mammals. As a result it was suggested that the continued proliferation of spiny dogfish resulted in their consuming large numbers of haddock and capping haddock recruitment (Sherman, 1991). However, the empirical evidence for strong and consistent ’interspecific interactions between spiny dogfish and haddock was not good. Rothschild (1991) looked at the multispecies interactions between 16 of the main stocks on Georges Bank. His analysis suggested that the majority of interactions, including those between haddock, skate and dogfish, were weak, in accordance with current food web theories (see Section 4.2) and the results of a study of multispecies interactions prior to the proliferation of dogfish (Sissenwine et al., 1982). While Rothschild (199 1) does not preclude the existence of interactions that are transient in space and time, and not readily detected using conventional statistical techniques (see Section 5), his results tend to suggest that the lack of fishing pressure on the dogfish and continued overfishing of the haddock are largely responsible for the changes observed, and that the recovery of the haddock would be encouraged by reducing fishing pressure on the haddock rather than simply attempting “pest control” with the dogfish. The willingness with which investigators attribute species replacements to changes in competitive or predator-prey relationships is a little surprising given that long-term data series describing the fluctuations of pelagic stocks in many systems are now available and that the environment, rather than intraspecific competition or predation, is usually shown to govern cycles in fish populations. Indeed, it is a combination of fishing and environmental changes which has been responsible for many of the collapses observed. Thus the abundance of herring (Jenkins, 1927; Hodgson, 1957) and pilchards (Culley, 1971; Southward, 1980) in European fisheries fluctuated for hundreds of years when levels of fishing mortality were relatively low. Similarly, studies of the abundance of scales within sediment cores have shown that there were large fluctuations in the populations of anchovies and pilchards in upwelling ecosystems (Soutar and Isaacs, 1974; De Vries and Pearcy, 1983; Shackelton, 1987; Baumgartner et al., 1992). In the English Channel, changes in climate largely explain the variation in pilchard and herring abundance (Russell et al., 1971; Southward, 1980; Southward et al., 1988). In upwelling ecosystems, the intensity of upwelling, and hence primary production, determines the biomass of fishes in the system (De

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Vries and Pearcy, 1983; Shackelton, 1987) and in the California current system for which the longest time series of abundance estimates exists, fluctuations in the biomass of anchovies and sardines have followed similar trajectories in fished (recent) and unfished periods (Baumgartner et al., 1992; Figure 23). Numerous reports of species shifts come from systems where so little study has been conducted that it is impossible to ascribe the changes to any specific influence (Kawasaki et al., 1991; Sherman et a[., 1993). However, it is often notable that the fluctuations were occurring long before the intensive industrial fisheries of the 20th century were in operation (Kawasaki er al., 1991; Sherman et al., 1993). In general, fishing has acted in synergy with environmental factors to magnify and accelerate the collapse of many pelagic stocks, but the collapse typically occurred at a time of marked environmental change and the replacing species would have been likely to proliferate in any case. It is probable that the environment, rather than any indirect effect of fishing, creates the window of opportunity for the “replacement” species. There is little convincing evidence to suggest that fishing has caused compensatory replacement of one fish stock by another even though such replacements are often cited as indirect effects of fishing. In reality, the main effects of fishing on community structure result from changes in the abundance of target species as fishers strive to maintain yields. It is relatively unusual for target species to fulfil keystone roles and for changes in their abundance to have substantial or consistent effects on other fishes which are prey or competitors. Daan (1980) in a review of the evidence for and against replacement also concluded that clear cases of the replacement of a depleted stock by another were difficult to find and not conclusive; his findings have been largely overlooked by many popular commentators. The marine ecosystem is often well adapted to fluctuations in component species and predator-prey relationships are often weak and plastic. 4.5. Scavengers and Discards

Fishing activities result in the capture of non-target species and undersized individuals of target species. A proportion of this by-catch will be discarded dead or dying because it is illegal to land it or because there is little or no economic gain associated with sorting or retaining it in relation to the catch of other groups. In addition, pelagic fishes such as mackerel may be “slipped” and returned into the sea when fishers have underestimated the size of the target school or overestimated the tow length and their catch is too large to be landed. The majority of these fishes are damaged during capture and confinement and die shortly afterward (Lockwood et al., 1983).

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Moreover, in fisheries which are managed using quotas, target fish above the minimum legal landing size may be rejected in favour of larger, more valuable, specimens. Alverson et al. (1994) have estimated that 27 million t of bycatch are discarded every year. This is approximately 27% of the current global fish catch. Discards are preyed upon by a range of scavengers whose ecology was extensively reviewed by Britton and Morton (1994). It is estimated that 475000t of fish, offal and benthic invertebrates are discarded into the North Sea annually (Camphuysen et al., 1993). Seabirds are a ubiquitous feature off the stern of every fishing boat. Camphuysen et al. (1993) undertook field experiments to calculate the percentage of each component of discards eaten by seabirds (Figure 27). They estimated that seabirds consumed approximately 90% of offal, 80% of roundfish, 20% of flatfish and 10% of the invertebrates discarded annually in the North Sea. This was estimated to be enough food to maintain c. 2.2 million seabirds; more than the total estimated population of scavenging seabirds in the North Sea. The effect of this additional supply of food was reflected in population changes, with a tenfold increase in the number of breeding seabirds from 1900-1990 (Lloyd et al., 1991; Furness, 1996). Examples of

Figure 27 The estimated annual consumption of different types of discards from trawlers in the North Sea by avian and non-avian scavengers and predators. (Data from Camphuysen et al., 1993.)

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changes in kittiwake and fulmar populations are shown in Figure 28. These patterns of change in seabird populations are similar to those which have been observed in Australia, Canada and the Falkland Islands (Blaber and Wassenberg, 1989; Thompson, 1992; Blaber et al., 1995; Chapdelaine and Rail, 1997). Field studies have demonstrated that there is intense competition for offal and discards between scavenging species (Hudson and Furness, 1988; Furness et al., 1992; Camphuysen et al., 1993; Garthe et al., 1996). Some species are more adept than others at utilising certain components of the discards. Fulmars, Fulmarus glacialis (L.) and gulls Larus spp. are the main consumers of offal in the northern and southern North Sea respectively, and their feeding success is positively correlated with their numerical dominance at fishing boats. Feeding success is also related to bird size and handling ability; thus kittiwakes consume smaller-sized fish whereas gannets take the largest components of the discards. The inability of smaller gull species to swallow large fish whole makes them vulnerable to kleptoparasitism by larger gull species, great skuas and gannets (Furness, 1996). Since the large quantities of fish discarded in the northeast Atlantic support populations of scavenging seabirds (Furness, 1992, 1996; Camphuysen et al., 1995; Garthe et al., 1996), Furness et al. (1992) have cautioned that plans to reduce discarding could have profound effects on their populations. Moreover, Camphuysen et al. (1993) advocated that any changes in mesh-

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Year Figure 28 Changes in the number of pairs of kittiwakes (open circles) and fulmars (filled circles) breeding on the east and northeast coasts of Britain. (Data from Furness, 1992.)

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size regulations which reduce the proportion of catch discarded should be introduced gradually to avoid adverse effects for competitively inferior seabird species. Oro et al. (1996) investigated the effects of a trawling moratorium on the reproductive and population biology of Audouin’s Gull Larus audouinii Payraudeau in northeast Spain. The larger of these colonies, on the coast, holds 70% of the total world population, and the smaller colony, on an island 600 km from the shore, holds 4%. The larger colony shifted diet during the moratorium but maintained reproductive success by feeding at inland sites such as rice fields, coastal lagoons, dunes and beaches. Such feeding areas were not accessible to the smaller colony and they were forced to fish actively at greater energy costs thereby leading to a decline in their reproductive success. With loss of additional feeding grounds a rare bird species could be threatened by the cessation of trawling. While conservation groups and some fisheries biologists are pressing for a reduction in the quantity of catch discarded, the practice of discarding has become so ubiquitous that such a reduction could have adverse effects on seabird populations. According to estimates by Camphuysen et al. (1993), seabirds consume approximately 50% of the material discarded into the sea. The remainder sinks to the seabed whereupon it becomes available to midwater and benthic predators and scavengers (Figure 27). Few studies have recorded the consumption of discarded material in midwater. Cetaceans and sharks feed on material discarded from shrimp trawlers in the Torres Strait, Australia (Hill and Wassenberg, 1990; Wassenberg and Hill, 1990) and killer whales, Orcinus orca (L.) were observed feeding on fish slipping through the meshes of nets at freezer trawlers off the Shetland Islands (Couperus, 1994). Whether this food source has significant consequences for populations of marine mammals remains unknown. The paucity of studies relating to midwater scavenging behaviour probably reflects sampling difficulties (Britton and Morton, 1994). Fishing activities provide two main sources of food for benthic scavengers: first, as food falls that originate from discards and by-catch that are not consumed by seabirds and midwater predators and scavengers (Figure 27); secondly, as demersal trawls and dredges are dragged across the seabed they dig up, displace, damage or kill a proportion of the epi- and infaunal animals in the path of the gear (see Section 2.2). In addition, some of the animals caught in the net may escape from the codend, but subsequently die. These latter sources of carrion have been termed “non-catch” mortality (Bergman and Santbrink, 1994). Falls of carrion are regarded as perturbations in deep-sea benthic communities, as they promote diversity by providing pulses of organic matter to localized areas of the seabed (Dayton and Hessler, 1972; Stockton and DeLaca, 1982). In this environment, scavengers move large distances to consume carrion, demonstrating the importance of

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this food source (Dayton and Hessler, 1972). Surprisingly, there have been few studies of the influence of carrion generated from fishing activities on the benthic communities of shelf seas (Ramsay et al., 1997b). The behaviour of scavenging fish species in response to trawling disturbance is frequently exploited by North Sea trawlers which have been observed lining up to fish the same tow (M. Fonds pers. comm.). Caddy (1973) noted that the density of predatory fishes in recently dredged areas was 3-30 times higher than in the area outside the dredge tracks. Kaiser and Spencer (1994) observed 35 times as many fish shoals over a recently beam trawled line compared with adjacent unfished areas. These studies implied that fish moved into areas of disturbance. Similarly, gadoids were observed to aggregate around newly disturbed pits in sandy sediments (Hall et al., 1994). Dietary analyses of gurnards (Triglidae) and whiting caught on recently beam trawled and undisturbed areas revealed that both species consumed significantly greater numbers of the amphipod, Ampelisca spinipes Boeck within the fished area. This amphipod constructs a tube that protrudes from the surface of the seabed which makes it vulnerable to contact with bottom fishing gear. Interestingly, gurnards normally eat large prey items such as shrimps, Crangon spp., and swimming crabs, Liocarcinus spp., but preferentially selected A . spinipes when feeding within the trawl tracks. This switch in diet implied that large numbers of amphipods were made available to predatory fish as a result of trawling (Hughes and Croy, 1993). Adult queen scallops, Aequipecten opercularis L. do not occur in the diet of whiting under normal circumstances. However, after trawling the distinctive orange gonads of these bivalves were recorded in whiting stomach contents, indicating that the molluscs had been damaged by the trawl (Kaiser and Spencer, 1994). Similar responses to fishing disturbance were also recorded for dab, Limanda limanda (L.) which were attracted to animals damaged by the trawl within 20 min, and increased to three times their former abundance after 24 h (Kaiser and Spencer, 1996a). However, these responses varied between different habitats. Although dabs aggregated in trawled areas in shallow ( < 20 m) sandy habitats (Kaiser and Spencer, 1996a) dab abundance was reduced 24 h after fishing in deeper (40 m) muddy areas (Kaiser and Ramsay, 1997). However, the diet composition of those dab captured in the trawled area differed significantly from those captured in adjacent undisturbed areas. Dab from the undisturbed areas mainly consumed the arms of the brittlestar Amphiura spp. which lie on the surface of the sediment, whereas those feeding in the disturbed area greatly increased their intake of brittlestars and fed predominantly on the oral discs of brittlestars. This suggests a more localized reponse to fishing disturbance in the deeper muddy habitat, and emphasises the influence of local environmental conditions on predator behaviour (Kaiser and Ramsay, 1997). It is clear that fish consume damaged or exposed animals in the trawl path, but there is no clear

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evidence, as yet, that they consume discards in the form of food falls from the surface (but see Olaso et al., 1996). Many of the mobile epibenthic invertebrate fauna are facultative scavengers (Britton and Morton, 1994). Many of these animals have physiological features or behavioural adaptations that enable them to survive the capture and discarding processes. For example, starfish are able to regenerate extensive damage to body tissues and hermit crab inhabit gastropod shells that protect them from mechanical damage. Not surprisingly, invertebrate scavengers are indicative of areas of trawl disturbance (Collie et d., 1997). Recent studies have demonstrated that scavenging invertebrates consume both discards and damaged fauna left on the seabed in the path of the trawl (Berghahn, 1990; Kaiser and Spencer, 1996a; Ramsay et al., 1997b). After the initial disturbance, or the arrival of a food fall on the seabed, a succession of scavenger aggregation occurs. Succession is governed by scavenger abundance, speed of movement and the prevailing current conditions (Sainte-Marie, 1986; Sainte-Marie and Hargrave, 1987; Nickel1 and Moore, 1992; Ramsay et al., 1996). Ramsay et al. (1997a) used time-lapse and video cameras to observe variation in the behavioural responses of scavenging species attracted to fish discarded from beam trawlers in different habitats. Starfish, Asterias rubens L. and hermit crabs, Pugurus bernardus L. were the two most abundant scavengers in offshore gravel and inshore sand habitats. The ratio of the background abundance of starfish to hermit crabs was 3:2 and 5:2 at the offshore and inshore habitat respectively. In all habitats, hermit crabs were the first scavenger to arrive at the carrion, rapidly increasing to a density of 330 m-’ in the offshore habitat. At such high densities, hermit crabs competitively excluded most species with a ratio of one starfish to ten hermit crabs feeding on the carrion. This contrasted with the inshore habitat where at least five species fed on the carrion with an equal ratio of starfish to hermit crabs. Far fewer hermit crabs were attracted to the dead fish, despite a similar background abundance compared with the offshore habitat. These observations are similar to the seabird interactions observed at trawlers, where the feeding success of fulmars is directly related to their numerical dominance in relation to other larger species (Furness et al., 1988; Camphuysen et al., 1993). Ramsay et al. (1997a) suggested that hermit crabs were attracted to fish carrion in greater numbers in the offshore habitat because (i) stronger tidal currents offshore produced a larger odour plume or (ii) competition for food was greatest in the offshore habitat. Experiments using traps baited with either dead fish or swimming crabs (animals typically killed by trawling) revealed distinct feeding preferences among different scavenger species. Hermit crabs preferred dead fish and avoided pots baited with dead crabs, whereas whelks Buccinum undatum (L.) showed the opposite responses (Ramsay et al., 1997a). Moore and

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Howarth (1997) demonstrated that scavengers avoided food in the presence of dead conspecifics or closely related genera. T h s does not explain the aversion of whelks to dead fish which is most probably a response to the presence of hermit crabs in the fish-baited traps. As demersal fishing gears are dragged across the seabed they dig up or damage a variety of fauna (see Section 2.2) which produces a trail of patchily distributed food resources. These different food resources are probably partitioned between each scavenging species by mechanisms of food preference and competitor/predator avoidance (Ramsay et al., 1997a). Although intense inter- and intraspecific competition occurs over foodfalls, there is evidence that group foraging enhances food acquisition (Brown and Alexander, 1994). Berghahn (1990) found that shore crabs, Carcinus maenas L. consumed entire fish within 3 h. In contrast, hermit crabs have relatively weak chelae that are poorly adapted for cutting flesh. The observations of Ramsay et al. (1997b) revealed that the majority of fish flesh was consumed within 24 h of arrival on the seabed. The feeding activities of increasing numbers of hermit crabs, enzymes secreted by feeding starfish and bacterial decay, all facilitated this process. Whereas populations of seabirds have shown clear responses to the extra food resources made available by discarding (Furness, 1996) the consequences for fish and invertebrate scavenger populations are not clear. In the period from 1970 to 1995 there have been increases in the biomass of several non-target species in the North Sea while the biomass of gadoids and species fished industrially decreased. Those species which have increased in abundance, such as the dab and long rough dab Hippoglossoides platessoides (Fabricus) (Heesen, 1996) are scavengers and may benefit from damaged benthic fauna which is released when beam trawls are dragged across the seabed. There have also been growth responses in plaice and sole which appear to be related to the development of the beam trawl fishery (de Veen, 1976; Houghton, 1979; Millner and Whiting, 1996). A recent study suggests that the biomass of organisms killed by trawling on the seabed is similar to the biomass of organisms discarded at the surface, i.e. 1.5 times the biomass consumed by seabirds is available for fish and invertebrate scavengers (Lindeboom and de Groot, 1998). However, there are many alternative explanations for the proliferation of dabs and long rough dabs and the increased growth of plaice and sole. For example, dabs and long rough dabs are small in size, grow rapidly and mature early (see Section 3.4.1) and eutrophication appears to have enhanced populations of polychaetes and brittlestars in coastal waters thereby increasing the food supply for juvenile flatfishes (Duineveld et al., 1987; Heessen and Daan, 1996; Rijnsdorp and van Leeuwen, 1996). So far, evidence for the expansion of populations of benthic invertebrate scavengers in response to carrion generated by fishing activities is weak,

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while increases in populations of scavenging seabirds are well documented. One of the main differences between seabirds and benthic invertebrate scavenging behaviour is the mechanism of carrion aquisition. Seabirds actively seek out and target fishing vessels as a source of food, migrating from distances in excess of 10 km (Camphuysen et al., 1993; Oro and Ruiz, 1997). Consequently, seabirds are able to obtain a relatively constant food source by feeding at trawlers. Conversely, benthic invertebrates rely on the chance occurrence of food-falls of fisheries carrion, or a trawling disturbance occurring on the seabed within the range of their sensory and ambulatory capabilities. Previous studies have shown that they migrate over relatively small distances of between 5 and 50m (Sainte-Marie and Hargrave, 1987; Nickell and Moore, 1992; Ramsay et al., 1997a), hence the chances of repeated encounters with fisheries carrion are low in all but the most intensely fished areas. Scavengers, such as crustaceans and starfish, may be better adapted to withstand the effects of repeated trawling disturbance, which, coupled with the removal of predators and competitors, has maintained their populations at a fairly constant level (Ramsay et al., 1997b). 4.6. Reversibility of Fishing Effects

The effects of fishing on the abundance of target fish populations are typically reversible over the short time periods (years rather than decades or centuries) which matter to managers operating within contemporary socioeconomic constraints. Indeed, the reversibility of fishing effects is an assumption of many conventional fisheries yield models. If small levels of fishing effort consistently shifted ecosystems to alternative states then species-specific fisheries would be rather more transient than they are at present. Thus heavily exploited plaice, sole and haddock stocks have continued to support fisheries in the northeast Atlantic for many decades despite yearto-year variations in biomass associated with fluctuations in recruitment and high levels of fishing mortality. When fishing mortality is reduced, the biomass of fish populations will characteristically increase once again (Myers et al., 1995) albeit after a few years if conditions for egg and larval survival are not immediately favourable (Corten, 1986, 1990). There are many examples of such resurgence in tropical and temperate ecosystems (Stephenson and Kornfield, 1990; Harmelin et al., 1995; McClanahan and Kaundaarara, 1996; Russ and Alcala, 1996b). In the few situations where one group of fishes has apparently replaced another in a fished system, and where fishing mortality has been reduced, environmental factors, and not competition or predation, are more likely to prevent the population resurgence of the overfished species (see Section 4.4).

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On longer timescales, the reversibility of fishing effects is not expected and it is wrong to regard system fluctuations as management failures. The widely overlooked modelling work of Shepherd and Cushing (1990) demonstrates that long-term fluctuations in fish stocks can be entirely self-generated by recruitment variation. Shepherd and Cushing (1990) modified a catch forecast model in which recruitment was controlled by a stock recruitment relationship with pseudorandom noise superimposed. The model assumed that other factors such as climatic variation, predator abundance and fishing mortality were constant. The modelled stock exhibited periods of abundance and decline which were entirely self-generated and the stock recovered from near extinction intermittently, purely by chance events. The behaviour of the model was not dissimilar from that of long-term changes which are usually attributed to climatic changes or interaction between stocks. Indeed, theoretical studies of long-term changes in fish stocks have widely questioned the assumptions of some fishery managers that stocks are naturally persistent (Steele and Henderson, 1984). Species are a useful marketing division, but ecologically we cannot expect stability in species composition. Steele and Henderson (1984) suggested that the rapid succession of dominant species every few years in the waters of New England and the North Sea may be considered as an acceleration of the longer period ecological jumps observed in the historical record rather than inadequate control of naturally persistent populations. 4.7. Conclusions

Since fishing increases the mortality of target and by-catch species it is widely assumed that it will lead to indirect changes in trophic and competitive interactions. Thus fishing may change the number of prey available to predators, the number of predators which consume prey, and affect the level of competition between species. The empirical evidence reviewed here suggests that such effects are by no means ubiquitous and, as in terrestrial food webs, the consequences of changes in species abundance, or even local species deletion, are rarely dramatic: those species which occupy keystone roles are the exception rather than the rule. However, while a tiny proportion of the species found in marine ecosystems perform a keystone role at any given time, their influence on the ecosystem can still be widespread and dramatic. The indirect effects of fishing on fish communities appear to be minor (unless they are mediated by habitat change; see Section 2), most changes in community structure result from changes to the fished habitat or the removal of target species. Those cases in which fishing affects a keystone species, such as sea urchins on some tropical reefs, are relatively rare. As

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Lawton (1989) suggests, the population dynamics literature has been biased towards significant interactions. The conclusion that most predator-prey relationships in marine food webs are weak is in keeping with current understanding of terrestrial and estuarine food webs studies which have often been overlooked in fisheries science (Raffaelli and Milne, 1987; Raffaelli et al., 1989; Paine, 1992; Polis and Strong, 1996). While some food webs do contain strong interactions involving “keystone” species there are also many neutral feeding links of limited consequence for the dynamics of the consumer or consumed (Lawton, 1989). Between these lie the possibilities: “donor-controlled” dynamics in which victim populations influence enemy dynamics but enemies have no significance on victim populations and the alternative case in which enemies influence victim populations but not vice versa. The evidence for donor-controlled dynamics between fish predators and prey is limited to a few instances where the fishes at higher trophic levels do not have access to alternative prey when the abundance of their main prey species declines. Thus the Barents Sea cod stock was severely affected by the loss of capelin and herring in a system where other forage fishes were absent. Many marine mammals and birds have markedly less plastic feeding strategies than predatory fishes, and thus the relationships between forage fishes and many species of marine mammals and birds are strongly donor-controlled. If fishing affects prey populations then fishing will indirectly affect their abundance and reproductive success. In these circumstances the debate about fishing effects largely centres around the role of environment and fishing in determining the dynamics of the prey fishes.

5. STUDY OF FISHING EFFECTS 5.1. Introduction

Despite a long history of the scientific investigation of marine fish populations and production processes in marine ecosystems (see Section 1) there is no widely accepted procedure for investigating the effects of fishing, and many other anthropogenic activities, in the marine environment. Those approaches which have provided the investigators with the greatest statistical power to detect change have been conducted at small scales. Thus the carefully replicated studies of the effects of beam trawling on benthic communities (Auster et al., 1996; Kaiser and Spencer, 1996b) have not, thus far, been possible at larger scales. Moreover, studies of fishery-induced changes in the trophic interactions between fishes have been largely dependent on correlative techniques (Jennings and Polunin, 1997). These approaches often

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lack the power to determine the significance of predator-prey relationships which are transient and plastic. In operational terms, a poor understanding of feeding interactions within the marine ecosystem is one of the factors that precludes active management of the ecosystem effects of fishing. We are increasingly concerned that scientists lack a good framework for testing interactions in the marine environment and are now claiming that fishing has many wider and deleterious impacts on ecosystem function because they intuitively believe this to be the case and not because they can convincingly demonstrate an effect (but see Section 5.2). If these claims are subsequently shown to be in error then there is a danger that the role of scientists in offering management advice will be further discredited. There is almost unlimited scope for research into the effects of fishing and for obtaining more site-specific information on the impacts attributable to a given gear or fishing strategy. Possibilities for research of this type have been recently discussed by Jennings and Lock (1996) for tropical reefs and Anon. (1995b) for the North Sea. However, the history of science suggests that marked improvements in our general understanding of fishing effects will take place by adopting novel research techniques and changing our philosophical approach to problems we encounter, rather than simply “filling in the gaps” using existing techniques. The results of existing studies, in particular those which relate to the direct effects of fishing on benthic organisms and their habitats (see Section 2) or the effects of fishing on relatively siteattached reef fish communities (see Section 3) already provide a basis for managing the ecosystem rather than target fish populations. However, in relation to effects of this type there are still concerns as to whether fishing effects in one region have marked implications for ecological processes in adjacent unfished regions, and whether changes in the productivity of communities impacted by trawls will have detectable impacts on fish production processes. The understanding of the indirect effects of fishing on trophic relationships in the marine ecosystem remains poor, not least because it has been constrained by the absence of an adequate experimental and analytical framework for testing fishing effects. For management purposes it is clearly necessary to know how resilience, stability and productivity are affected by fishing at different trophic levels. In this section we consider how the study of fishing effects could be improved. In Section 5.2 we discuss the reasons why conventional statistics may not help us to detect fishing effects, and consider the alternatives. In Section 5.3 we ask whether current approaches for the investigation of marine food webs, and the effects of fishing on their structure, can be improved. Section 5.4 is concerned with the modelling of ecosystem processes in fished and unfished ecosystems and the utility of approaches which aggregate across species. Lastly, in Sections 5.5-5.7 we discuss the selection of research sites for the study of effects, the temporal and spatial scales on

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which such studies should be conducted and the advantages of controlled manipulative studies. 5.2. A Statistical Basis for Correlative Studies

Most attempts to detect the effects of fishing have been based on a test of the null hypothesis that there is no fishing effect (Dayton et al., 1995). In such ecological investigations, particularly those on a large scale, the errors associated with the data are extremely high, and the investigations often have little power to detect change. As a result, the null hypothesis may be accepted even when it should be rejected (type I1 statistical error). Subsequently, when the results of studies are passed from scientist to politician to manager the idea that there is “no power to detect a fishing effect” is replaced with the assertion that there is “no fishing effect”. Such concerns prompted Peterman (1990) to advocate the routine use of statistical power analysis, in order to ensure that those who interpreted the results were aware of the probability of committing a type I1 error. Existing approaches to testing for fishing effects have been based on minimizing the effects of type I errors (stating that a fishing effect exists when there is no such effect) and usually state that the frequency of such errors should not exceed 5%. However, the probability of making a type I1 error and not detecting a fishing effect when one exists is rarely considered (Dayton et al., 1995; Thrush et al. 1995). Dayton et al. (1995) suggested that policy makers are largely unaware of the basis on which most statistical tests are made, and should be informed of the consequences associated with making either type of error. These concerns are further discussed by Peterman (1990), Peterman and M’Gonigle (1992), Shelton (1992) and Dayton et al. (1995). Parametric statistics are widely used for conducting statistical tests in fishing effects studies and yet they are often inappropriate. In particular, much ecological data, even after transformation, do not meet the requirements of parametric tests, multidimensional problems involving biotic and abiotic variables are difficult to address on a holistic basis and many parametric tests cannot be used with data describing community attributes such as trophic structure and diversity. Since the multidimensional properties of ecosystems are of particular concern to those wanting to study the wider effects of fishing, statisticians must start to develop a new framework for hypothesis testing and detecting the effects of fishing. Such work has recently revolutionized approaches to investigating the effects of pollution (Clarke and Green, 1988; Clarke, 1993; Clarke and Ainsworth, 1993) and these approaches are increasingly used to investigate fishing effects (Thrush et al., 1995; Greenstreet and Hall, 1996; Jennings and Polunin, 1996a; Kaiser and Spencer, 1996b).

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The overwhelming problem associated with detecting the indirect effects of fishing is that species may only interact occasionally and yet these interactions have long-term impacts on community structure. Such interactions might occur when distributions coincide at one stage in the life history, when a transient temperature change leads to redistribution or when one species is unusually abundant following a successful recruitment event. Although the interactions are short-lived in space and time, they may have powerful structuring effects, and these effects can persist for many years even if no further interaction occurs. There are no recognized procedures for estimating the strength, frequency or consequences of transient interactions. Even the most effectively funded of contemporary monitoring programmes lack the resources to monitor the many aspects of individual behaviour that will determine the frequency and strength of interaction.

5.3. Investigating Marine Food Webs

There is considerable scope for improving our empirical understanding of predator-prey relationships in fished systems and, at present, some of the most useful approaches have not been used to their full potential. The two areas of particular interest relate to the effects of fishing on trophic relationships and how ontogenetic shifts in diet, and plasticity in feeding strategies at all stages in the life history, affect the response of species to changes in the ecosystem. Stable isotope analysis could provide a guide to relationships in marine food webs but the full potential of this approach is yet to be utilized for the study of fishing effects. In particular, stable isotope analysis would allow long-term investigations (using historical collections of material such as otoliths, scales or feathers) of the effects of fishing on trophic relationships. Studies of stable isotope composition provide indications of the origins and transformations of organic matter (Fry and Sherr, 1984; Owens, 1987; Preston, 1992) and thus provides a useful tool for investigating trophic relationships (Fry, 1988; Yoshioka et al., 1994; Peterson et al., 1985; Peterson and Fry, 1987). Historically, sampling constraints have prevented the description of stomach contents over ecologically meaningful periods (Hobson and Welch, 1992, 1994) and the assumption of fixed diets in many predator-prey models is bound to suggest tighter predator-prey interaction than occurs in a system where many species probably switch diet quite freely. The I5N and 13C composition of fish tissue will reflect the composition of assimilated food and provide a long-term indication of feeding strategy by integrating the differences in assimilated food over time (Hobson and Welch, 1992, 1994; Thomas and Cahoon, 1993).

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In the few circumstances where stable isotope analysis has been used to investigate trophic relationships in the marine environment the results have been highly relevant to our understanding of fishing effects and the role of plastic feeding strategies. Thus Wainwright et al. (1993) examined scales which had been collected over many years from fishes on the Georges Bank and the results suggested that the feeding habits of many species underwent considerable year-to-year variations. Jennings et al. (1997) measured 15N in the tissues of Mediterranean reef fishes and found that the fishes adopted different feeding strategies at different sites. Consequently, the same species could sometimes be assigned to different trophic groups at different sites. These data provide empirical support for current theories of food web dynamics which suggest that fixed trophic levels are rarely a feature of food webs (Polis and Strong, 1996) and support the ideas developed in other empirical studies that the removal of predatory fishes will have limited impacts on the productivity of their prey (see Section 4.2). It would be particularly interesting, and feasible, to use historical material in order to compare shifts in the trophic structure of fish communities where there is evidence for donor control (such as in the Barents Sea ecosystem, see Section 4.3 and Figure 19). Stable isotope studies have also provided insights into changes in the feeding strategies of seabirds during the decline and development of fisheries. Thus Thompson et al. (1995) examined stable isotope ratios in northern fulmar Fulmaris glacialis feathers collected in the early 19th century and 1993. The changes in I5N and 13C composition were consistent with a change in diet from prey of high to low trophic position and suggested that the fulmars formerly fed on offal generated by whaling activities but with the cessation of whaling they increasingly fed on fishes. Given that large-scale changes in many marine ecosystems have been well documented since the late 19th and early 20th centuries, and that collections of scales, otoliths and other body parts throughout this period are available in fisheries laboratories and museums, there is considerable scope for using stable isotope analysis to examine shifts in community structure as fish, seabird, whale and seal populations proliferated and declined. Moreover, stable isotope analysis could be used to examine spatial and temporal shifts in trophic interactions in relation to fishing intensity and provide a better insight into the responses of marine ecosystems to changes in fishing pressure. 5.4. Modelling Ecosystem Processes

Historically, there has been considerable interest in food chains and their role in driving fish production (Ryther, 1969; Gulland, 1970; Steele, 1974).

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Given the logistical difficulties associated with conducting studies of production processes at the ecosystem level, and given that the results of many uncontrolled experiments are confounded by the dynamic behaviour of fishers and managers, one option is to model the ecosystem and test the effects of different fishing strategies. Two types of models have been most widely investigated, dynamic models (Andersen and Ursin, 1977) and steady-state models which assume that trophic groups have a constant biomass (Walsh, 1975). Models of the latter type have been increasingly developed with a view to providing information for fishery management (Larkin and Gazey, 1982; Polunin and Klumpp, 1992; Christensen and Pauly, 1993). These models may help to quantify and predict the changes in production of systems which are subject to fishing pressure and the modellers accept that there is little possibility of doing this on a species-by-species basis in trophically complex systems. Moreover, incorporating additional complexity into models is not always desirable: a stage is often reached where further complexity contributes more to the variance in output than to realism. Production models based on an analysis of the size distribution of organisms within ecosystems may provide a basis for assessing the effects of fishing. The size distribution of the biota in aquatic ecosystems follows regular patterns because size constrains predator-prey relationships and the transfer of energy. Production, respiration and production to biomass ratios are closely related to the body size of individual organisms and are reflected in relationships between biomass density and body size (Boudreau et al., 1991; Boudreau and Dickie, 1992; Thiebaux and Dickie, 1992, 1993). Dickie et al. (1987) suggested that energy flow within ecosystems could be studied through analysis of the body size spectrum of biomass density and the spectral distribution of production with body size. Specific production and biomass density can then be expressed as functions of individual body size, without referring the observations to taxonomic units (Boudreau et al., 1991). Such models are particularly appealing when a single species of fish may be a prey organism, a predator on other fishes or a cannibal in the course of its life history and when its feeding strategy will often depend on the relative abundance of prey species which vary in space and time. Size-based production techniques have been seen as a means by which to reduce the huge investments of time and energy required for assessing the production of fish stocks (Sprules and Stockwell, 1995). While there are good reasons to employ the species as the principal taxonomic unit in fished systems, notably because species or intraspecific stocks were the targets of fishers and the categories favoured by buyers or consumers (see Section l), species-based research often fails to provide useful information on the general properties of ecosystems. Individual species undergo marked variations in biomass owing to massive and largely unpredictable variations in larval

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survival (Hjort, 1914; Zijlstra, 1963; Cushing, 1988a; Doherty, 1991; Doherty and Fowler, 1994), while the yield from, or biomass of, whole communities often remains relatively stable (Sutcliffe et al., 1977; Holden, 1978; Pauly, 1979; Ursin, 1982; Sissenwine, 1986). Such stability provides compelling reasons for developing indices of production which aggregate over species (Dickie et al., 1987; Sprules and Stockwell, 1995). At these scales, a chaotic system appears to be relatively stable and small-scale noise blends into large-scale pattern. The size distribution of normalized biomass in aquatic ecosystems can be modelled using primary linear scaling (Platt and Denman, 1977, 1978) (see Section 3.3) and a series of repeated domes corresponding to component trophic groups, such as zooplankton and fishes (Thiebaux and Dickie, 1992, 1993; Duplisea and Kerr, 1995). Such structural regularities suggest that the size distribution of any trophic group can be predicted from the study of size distributions in other trophic groups. Furthermore, since production to biomass ratios are dependent on body size, a biomass size distribution can be scaled to provide a production estimate. Sprules and Goyke (1994) and Sprules and Stockwell (1995) have reported zooplankton size distributions in Lakes Ontario and Erie and used these to estimate potential fish production. It should be possible to apply such methodology to marine systems in order to assess the effects of fishing on production processes. The methods would be particularly appropriate in upwelling and pelagic ecosystems. Steady-state models are increasingly used to model trophic interactions in marine ecosystems and recent developments have allowed these models to be used for predicting the effects of fishing on trophic interactions. The steadystate model of Polovina (1984) was used to estimate the biomass of component trophic groups in a reef ecosystem, but Christensen and Pauly (1992) incorporated a theory for analysing energy flows within the system (Ulanowicz, 1986) to produce the ECOPATH models which include routines for estimating production: biomass within trophic groups and the energy flows between groups (Christensen and Pauly, 1992, 1993). Pauly and Christensen (1995) used the results of ECOPATH models for a number of aquatic ecosystems (the quality of input data was very variable (Christensen and Pauly, 1993)) to estimate the proportion of primary production required to sustain global fisheries (Figure 1). In this calculation, Pauly and Christensen (1995) assigned fractional trophic levels (effectively a weighted average of the trophic levels at which a group feeds) to different harvested groups, and assumed an energy transfer efficiency of 10% between trophic levels. While the ECOPATH approach provided a static description of trophic structure in exploited ecosystems, it did not allow the prediction of changes which would result from specific management actions or changes in the

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ecosystem. Recent developments of these models (Walters et al., 1997) have replaced the linear equations that describe fluxes with coupled differential equations that can be used for dynamic simulation and the analysis of changing equilibria. There are currently a number of problems with the predictions made by these models (Walters et al., 1997), but with refinement they are likely to provide a valuable tool for assessing the effects of fishing on ecosystem processes and developing remedial management measures. Moreover, looking at the effects of fishing using models of this type will reorient attention from population to ecosystem-based issues (e.g. JarreTeichmann, 1998). In the few marine ecosystems where some species have keystone roles, conventional models in which individual predator-prey relationships are tightly coupled can provide effective predictions of the effects of fishing on community structure. McClanahan (1995a) developed an aggregated energy-based ecosystem model for a Kenyan coral reef and looked at the effects of changing fishing intensity and catch selectivity on rates of reef coral accretion or degradation, algal and urchin biomass and on the biomass of herbivorous, piscivorous and invertebrate feeding fishes. The r2sults indicated that fishing changed many ecological processes on the reef and that the benefits of short-term fishery yield had to be weighed against impacts on reef structure and processes. If all fish groups were targeted by the fishers, then the reef system became dominated by sea urchins which reduced algal and coral biomass and productivity (see Section 2.3). Fishing solely for piscivorous species resulted in low fisheries yields but high reef accretion. This was the result of herbivorous fishes grazing the algae and releasing coral from competition. A management strategy based on fishing piscivorous and herbivorous resulted in the highest and most stable yields, but at high levels of fishing the lack of herbivorous fishes may cause algae to competitively exclude coral. The theoretical predictions of this model were supported by McClanahan’s empirical work (McClanahan and Muthiga, 1988; McClanahan, 1990 1992, 1994a, 1995b; McClanahan and Shafir, 1990; McClanahan et al., 1996). This was one of the most convincing studies of the ecosystem effects of fishing activities and emphasized the potential for using models in the management of some marine ecosystems. 5.5. Selection of Research Sites

The empirical study of fishing effects is hampered by a lack of unfished control sites. The most intensively studied marine ecosystems are also the most heavily fished and, in most cases, were fished before scientific study began. Moreover, these ecosystems are affected by other anthropogenic activities such as waste disposal, oil, gas and gravel extraction and shipping

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which may lead to eutrophication, chemical pollution, sediment deposition, hypoxia and anoxia (Anon., 1995b). The effects of fishing and other anthropogenic activities may interact but such interactions are virtually unstudied and poorly understood. It is clear that the most dramatic changes in fish community structure occur when fishing begins or when fishing pressure first increases from a very low level (see Section 3.4.1). Once fishing has begun, and the system has entered the “fished” state, any changes in response to increasing fishing effort are often smaller and more difficult to detect (Figure 10). In many parts of the north Atlantic, ecosystems entered the fished state several decades or centuries ago, and we have no means of knowing their condition prior to this time (Cushing, 1988b). Moreover, the dynamic behaviour and multiple impacts of fishers means that it is increasingly difficult to determine the indirect effects of fishing and in many parts of the North Sea, for example, the types of gear used by fishers have so many effects on habitat and non-target species that it is questionable whether any species are not directly affected. Furthermore, the allocation of effort to catching different species has changed dramatically over time as fleets respond to changes in recruitment rates or stop fishing as regulations are imposed. Better opportunities for testing the indirect effects of fishing are provided by those sites where the fishing methods used are not destructive to habitat, where dead by-catch is not discarded and where fishers do not alter their fishing patterns dramatically in response to changes in the abundance of target species. Some pelagic fisheries and some fisheries on tropical reefs meet these criteria, but the ubiquity of intensive fishing is such that they are increasingly scarce. For scientists working in heavily exploited regions, finding control sites is increasingly difficult. Thus studies of the effects of trawling have compared communities in the vicinity of wrecks or oil rigs with those in adjacent fished regions (Hall et al., 1993). Ironically, such structures have also been used in studies which investigate the effects of the structure on benthic communities. The lack of suitable control sites is one reason why protected areas are needed in many heavily fished ecosystems (see Section 6 ) . 5.6. Spatial and Temporal Scales of Study

In marine ecology, as in terrestrial ecology, research at the largest spatial scales (e.g. planktonic production processes) focuses on organisms with fast dynamics and the studies often aggregate across species and individuals. At these spatial scales, species-level processes are viewed as unimportant or intractable to study. With decreasing spatial scale and increasing trophic level, the properties of individuals are regarded as more important. Thus

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zooplankton are a homogeneous mass but a blue whale has intrinsic value. Bar the few exceptions described in preceding sections, studies of fishing effects do not aggregate across species and are often conducted at small spatial scales. Ecologists probably receive more funding if they are observing fishing effects rather than obtaining “negative” results. Thus studies of the effects of fishing are increasingly oriented towards finding keystone species, even though a dispassionate examination of existing research suggests that they do not exist in many systems, their role may be transient, and the majority of interactions within food webs are weak. Moreover, given the framework for statistical testing which is currently in use (see Section 5.2), science is increasingly pushed towards smaller spatial scales, longer temporal scales and very homogeneous study sites in order to increase statistical power. Such emphasis is further driven by many reviewers and referees who increasingly demand statistical elegance even though scientists may no longer be investigating process on scales of relevance to managers and conservation bodies. Some ecosystem-based studies may lack statistical elegance, but at least the practitioners seek effects at the appropriate spatial and temporal scales. One feature of small-scale fishing effects studies with benthic and relatively site-attached fishes is that they have not indicated the spatial scales at which fishing effects will be significant in the longer term. The biomass of benthic or fish species at many sites is likely to be maintained by the settlement and subsequent growth of larvae which are the progeny of adults spawning outside the study area. Clearly, we need to further our understanding of larval transport in relation to hydrographic phenomena to determine the extent to which an area is self-recruiting or reliant on larvae from other sources. Such studies are particularly relevant to management issues because they will demonstrate the extent to which the effects of fishing are collective (Jennings and Lock, 1996). While we do not know the scale on which the collective effects of fishing are apparent we do not know the scale on which to conduct long-term investigations or the scale on which to impose management strategies (see Section 6). Given that many fishing effects studies have been conducted on small scales (often km to tens of km on tropical reefs), it is difficult to determine the cumulative effects of fishing on a global scale. On reefs, for example, there has been much interest in the effects of fishing on ecosystem processes such as accretion and bioerosion (McClanahan, 1989, 1995a; Done, 1992; Hughes, 1994; Roberts, 1995; Jennings and Lock, 1996; Jennings and Polunin, 1996b). However, there remains little scientific evidence (though some circumstantial) for ecosystem shifts outside the Caribbean and East African regions, and some reef fisheries in the Pacific islands appear to support relatively high yields with little evidence for adverse ecosystem shifts

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(Dalzell, 1996; Dalzell et al., 1996). Until techniques such as remote sensing can be more widely applied to catalogue large-scale changes on many reefs (Green et al., 1996) it is unclear whether the effects of fishing reported from a few sites can be used as a basis for useful generalizations on a global scale.

5.7. Conclusions

There is increasing interest in the study of fishing effects in marine ecosystems and yet there remains poor communication between fish population biologists and many others with interests in marine ecology. In order to develop a broad-based framework for the study of fishing effects it would be advantageous to strengthen, once again, the links between fish population biology, marine ecology and oceanography. Many pioneers in the field worked freely across these disciplines (Ryther, 1969; Steele, 1974; Cushing, 1975, 1982; Steele and Henderson, 1984; Sharp, 1988; Southward et al., 1988). We need to develop a good statistical basis for assessing the indirect effects of fishing at larger scales, since studies that provided investigators with the greatest statistical power to detect change have been conducted at small scales. In particular, studies of fishing-induced changes in the trophic interactions between fishes were often dependent on correlative techniques, short-term dietary studies and models based on tightly coupled predatorprey relationships. These approaches are inappropriate when predator-prey relationships are transient and plastic. A poor understanding of feeding interactions within the marine ecosystem is one of the factors which precludes active management of the ecosystem effects of fishing. Such interactions need to be further investigated. In order to assess the status of ecosystems and to devise management plans it is necessary to know which types and levels of yield are sustainable, which cause long-term shifts in the ecosystem and whether judicious management of the fishery can alter the probability of attaining favoured states. Control sites, unaffected by fishing, are an essential prerequisite for understanding the effects of fishing on marine ecosystems, since they provide an indication of the fluctuations which occur in the absence of a human predator. At the ecosystem level, all marine environments are affected by fishing. Even at smaller scales (tens of km), it is increasingly difficult to find sites where habitats or fish populations are not directly affected. Such sites are largely confined to isolated reefs in the tropics and parts of the Arctic and Antarctic.

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6. MANAGEMENT

6.1. Introduction

The well-documented economic collapse of fisheries and evidence for the wide-ranging indirect effects of fishing (see Sections 2-4) lead us to the expected conclusion that many marine ecosystems are overfished and that better management is needed. We do not wish to discuss overfishing, which has been widely treated elsewhere (Anon., 1997); suffice to say that wellmanaged fisheries would yield more than most fisheries currently do and that the world’s fishing fleets are massively overcapitalized. In 1993, for example, $US 54 billion was estimated to have been lost through overcapitalization of the world’s fishing fleets and $US 25 billion through the depletion and overexploitation of stocks. A number of properties will be common to any effective management strategy: the management goals must be defined and clearly understood by all those whom they will affect, the strategy must be operationally simple and easy to enforce, and it must be possible to determine, in quantitative terms, whether or not management has been successful. Few existing strategies have these properties (Pauly, 1997). Population-based management has become a complex and inaccessible subject, the increasing complexity often favouring the short-term interests of a few fishers rather than the longterm supply of protein and the preservation of habitat, function and diversity in marine ecosystems. In this chapter we discuss the advantages and disadvantages of population-based management, approaches to management which minimize the direct and indirect effects of fishing, and the case for marine reserves as an adjunct to other management methods.

6.2. The Role of Fisheries in Marine Ecosystems

Waste disposal, mineral exploitation, shipping and fisheries dominate activities in the marine environment, and the marine environment is widely treated as common property; albeit within national exclusive economic zones. The primary aims of management range from maximizing fisheries yield to maintaining diversity in areas of intrinsic conservation value. Such aims are rarely compatible. Thus the introduction of Nile perch to the African lakes, and the associated loss of cichlids, have led to marked reductions in diversity but the “function” of protein production has improved (Bare1 et al., 1991). In the North Sea, the activities of beam trawlers (see Sections 2.2 and 2.3) have marked effects on some benthic communities. Nevertheless, the highly productive fauna which dominates

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many of the shallow areas subject to continual natural disturbance, is probably promoted through the regular scouring by beam trawls, which maintains the food supply for target flatfishes. There are anecdotal reports of fishermen deliberately breaking up large sponge reefs off the Fseroe Islands, not as an act of vandalism, but with the aim of “preparing the ground” to provide a more productive feeding area for target species (A. Nsrrevang quoted in Anon., 1993a). While removing the sponge reef had the desired effect, scientists were able to convince the fishermen that the reef habitat promoted the survival of juvenile commercial species and that this outweighed the benefits of destruction. In many developing countries, small-scale artisanal fishing is the main source of food and income for a large proportion of the coastal population. In these circumstances, the immediate value of fishes may well dictate that the main goal of ecosystem management is to maximize fish production. Provided that the management strategy does not cause habitat degradation, it may be acceptable to tolerate shifts in community structure. Some insurance against the wider adverse effects of fishing can be provided with networks of marine reserves (see Section 6.5). Marine reserves can also bring localized, but in some cases substantial, benefits through encouraging tourism. Attempts to manage marine ecosystems are most likely to be successful in those relatively wealthy areas where socio-economic analyses suggest that society is willing to police the fishery and has sufficient resources to do so. Clearly the high human population density on many tropical coasts, coupled with the limited area of accessible reef, will mean that the fishery could not support the population even if it were perfectly managed (Jennings and Polunin, 1996b). At best, effective fisheries management will ensure that the maximum total yield will be taken from reefs on a sustainable basis but it will never be a panacea for poverty in coastal communities. Other sources of food and money must be provided to allow the recovery of fisheries and to allow the recovery of degraded reefs which also perform key functions such as coastal protection. In some developed countries, the majority of the population may gain minimal benefit from the exploitation of the marine environment. In these circumstances, conservation, recreation and other uses may take precedence over extracting the maximum yield of fishes, and it will be necessary to consider whether intensive, and often unprofitable, fishing is ecologically acceptable. Notwithstanding, there is little reason why relatively large yields cannot be taken without adverse effects on the marine ecosystem, birds and marine mammals if fishing effort and fish production can be synchronized more effectively (see Section 6.4) and marine reserves are used to provide insurance against widespread habitat degradation and population collapse (see Section 6.5).

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6.3. From Population to Ecosystem Management

When the first mathematical models which described fish population dynamics and the effects of exploitation were developed (see Section 1) it was assumed that they would provide a basis for introducing effective management. This assumption has often proved unfounded and management advice, at least in relatively wealthy countries, is now provided using less formulaic approaches such as virtual population analysis (VPA) (Pope, 1979). VPA has subsequently been modified to account for multispecies interactions (MSVPA), an approach which started to bridge the divide between population and ecosystem-based studies and stimulated new interest in species interactions and predation mortality (Sparholt, 1990; Gulland, 1991). While these models do not account for interactions between nontarget species they still provide good short-term forecasts of the population structure of exploited stocks. From the managers’ viewpoint, interactions between non-target species in these temperate systems were viewed as unimportant because their effects could not be directly related to changes in the ratio of fishing to natural mortality. Since “pest control” in large temperate marine ecosystems is still considered unfeasible by many managers (unless the pest itself has economic value), interactions with unexploited predators or prey, which may come and go uncontrolled, are of less interest (Daan, 1987). The collapses of fish stocks in temperate regions currently have little to do with inadequacies of population-based fishery science, and most scientists are well aware when stocks are endangered even if their warnings are not heeded by policy makers (Myers et al., 1996; Cook et al., 1997). While Individual Transferable Quotas and similarly powerful management tools can be used in those fisheries where assessment and management are well funded, they are not globally applicable. Most management strategies are simple variations on a theme and they do little to remedy the overriding problem that human fishing effort must be accurately synchronized with fish production. Fishing fleets develop rapidly when fishes are abundant, but do not shrink with equal rapidity when fish production decreases following poor recruitment. Moreover, scientific advice is not translated into regulations which are operationally simple and easy to enforce (Holden, 1994) and managers, fishers and fish populations do not operate on similar timescales. On those rare occasions when simple and decisive regulations are enforced, such as the closures of fisheries after collapse, they are usually effective. Unfortunately, such measures are a reaction to a crisis rather than a proactive approach to ensuring the sustainable exploitation of the marine environment. Much has been written about fisheries management and the effects of anthropogenic activities in the marine environment. Fisheries cannot be

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managed in isolation, and ecosystem management should integrate the activities of many regulatory bodies which control the marine environment. Even a perfect reef fisheries management strategy will not work if corals are being mined on the fishing grounds or if sediment is washed on to the reef from coastal farms. 6.4. Managing Trophic Interactions and Ecosystem Processes

There have been a number of attempts to manage the relationships between predator and prey species. Thus capelin in the northwest Atlantic are managed on a conservative basis to preserve them as forage species for cod, and anchovy were intensively fished off Namibia in an attempt to shift the ecosystem to a phase dominated by the more valuable pilchard (Shelton, 1992). Sugihara et al. (1984) reported that hake, chimaera and sharks were being cleared from a portion of the Spanish continental slope with the hope of producing a shrimp fishery and Rothschild (1991) suggested that the targeted fishing of spiny dogfish and winter skate, which had come to dominate the demersal community on Georges Bank, would be a worthwhile experiment as it might lead to increases in the stocks of more valuable groundfish. In addition, concern about the negative effects of marine mammals on fisheries and vice versa have stimulated a range of approaches to management. In Alaska, for example, there have been pleas to fishery managers to reconcile the need to address serious declines in populations of seals with the production-oriented goals of fishery management (Huppert, 1991) and there have been calls to limit fishing for krill as this might slow the recovery of whale populations (Lyster, 1985). Conversely, concerns that mammals were competing with fishers have led to suggestions that seabirds and fur seals in southwest Africa should be controlled to protect fisheries (Crawford et al., 1992). Earle (1996) and Hutchinson (1996) review other examples of where cetacean control has been suggested to boost fisheries yields. In reality, most of these rather crude approaches to managing the system do not have the predicted result. Intuition is not a good substitute for scientific understanding and the idea that some of the most valuable species at the highest trophic levels should be cleared to boost fisheries for their prey may result in many fishers trying to do their best for the ecosystem! In the few ecosystems where strong relationships between predator and prey dynamics have been identified, such as the Barents Sea and East African coral reefs, there is the possibility of managing ecosystem function. For example, McClanahan et al. (1996) attempted to modify interactions in Kenyan reef ecosystems by removing those sea urchins which would have suppressed the recovery of algal, fish and coral communities once fishing

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effort was reduced (see Section 2.3). After removing sea urchins from unfished experimental plots there were significant increases in fish density and diversity (species richness), and algal cover, within I year. On fished reefs, where the biomass of herbivorous fishes was low, algal cover increased as corals were overgrown by algae. On the unfished reefs, where the algae were more intensively grazed, coral cover increased. The authors cautioned that urchin removal would not be a good strategy on reefs where no attempt was made to prevent fishing. It is notable that the shifts observed in their manipulative studies were in accordance with those observed following the mass mortality of Caribbean urchins (see Section 2.3). In other ecosystems, such tight predator-prey interactions are atypical and the management strategies which have been proposed tend to reflect our understanding of ecosystem function. Thus in the North Sea, where strong links between predator and prey dynamics are rarely observed, there has been greater emphasis on managing species individually and in small groups. In those ecosystems that are susceptible to undesireable shifts, the manager can try to prevent them by adopting a number of approaches to harvesting. The potential benefits of some of these approaches are discussed in more detail by Christensen (1996) and Jennings and Polunin (1996b). They include the removal of predatory fishes to increase the biomass of their prey and fishing at lower trophic levels where potential production may be higher. In tropical reef systems, we believe that the understanding of fishing effects, coupled with data on yields that have empirically been sustained, provides a basis for improved management. In order to promote recovery in damaged fisheries, and to prevent adverse shifts in existing fisheries, habitatdestructive fishing techniques must not be used and fishers should crop many species of fishes and invertebrates from many trophic groups (Jennings and Polunin, 1996b). Widespread cropping increases the probability that a given total yield can be sustained in the long term and reduces the risk of deleterious shifts in ecosystem function. In addition, if fishers are used to catching, eating and marketing many types of fish, they are more likely to accept the changes in catch composition which follow changes in recruitment success and are less likely to target specific species until they reach economic extinction. Reef fishery managers should also seek to identify and protect the few species of fishes, such as urchin predators, which have keystone roles. A broad-based cropping strategy is comparable with that used by predatory fishes. It is notable that there is little evidence of top-down predator control of prey fish abundance in marine ecosystems, unless humans are the predators. The majority of fish biomass in the most intensively fished ecosystems is still consumed by other fishes (Bax, 1991; Overholtz et al., 1991) and yet humans have a capacity to fish species to economic extinction. This is not because the system does not have the capacity to accept the

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increased “predation” caused by fishing (a large year class of a piscivorous fish does not lead to the extirpation of prey species), but reflects the fishing strategies used by humans. It is possible that we can learn something about ecosystem management from observing the feeding strategies of predators. The key difference between many human fishing strategies and those of predatory fishes is that the fishers often target a few species of fishes very intensively (there are exceptions in tropical and temperate fisheries). Most predatory fish, conversely, are very generalist feeders, often switching to invertebrate prey or cannibalism and eating many species of fishes at different stages in their life history. Predatory fishes rarely show the energetically inefficient human trait of pursuing the same species and are more likely to switch prey (Hughes and Croy, 1993 and references therein). The feeding strategies adopted by predatory fish are rarely used by human fishers unless they are entirely reliant on their fishery as a food source (Hughes, 1993; Jennings and Polunin, 1996b). This reflects the different currencies that impinge upon foraging decisions taken by humans involved in commercial fisheries (financial gains) and predatory fishes (energetic gains). Such observations would suggest that humans should crop fishes less selectively and allocate their effort in proportion to the abundance of the different fished species. Unfortunately, increased targeting of a ‘wider range of species tends to be a reactive response to failure of a fishery for favoured species rather than a proactive approach to management of a sustainable fishery (Jennings and Polunin, 1996b). Many recorded collapses of fisheries for highly productive and short-lived species such as sardines, anchovies or capelin would have occurred in the absence of fishing pressure and there would have been knock-on effects on bird and mammal populations. If fishing can be more effectively synchronized with the fluctuating production rates of these fishes then there is little reason why humans should not benefit from the vast protein resource that is available following years of good recruitment. The important factor is that fishing effort must be rapidly reduced to very low levels when strong year classes are not entering the fished stock. 6.5. Marine Protected Areas

Levels of fishing effort, well below those which are sustainable, have significant effects on the diversity and structure of fish communities. The greatest effects of fishing on habitat, genetic diversity, age structure, life history traits and total stock size are most commonly observed when previously unfished areas are fished for the first time (e.g. Figures 10 and 16). Since all marine ecosystems are now fished (with the exception of deep-sea abyssal systems), marine ecologists do not have access to large areas where the

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structure and natural variability of systems can be studied and provide controls for fishing effects studies. Contemporary and large-scale (100s and 1000s of km) studies of fishing effects inevitably examine the effects of recent changes in fishing effort rather than the changes which occur when unfished areas are fished for the first time. Marine reserves would allow the comparison of ecological processes in fished and unfished marine environments while providing conservation benefits for some species which are impacted by fishing. Marine reserves offer many potential benefits as a management tool, as reviewed by Bohnsack (1990, 1996): 1. they simplify management and enforcement; 2. they protect habitat and benthic biota from direct (physical disturbance) effects of fishing; 3. they provide control sites for “fishing effects” studies and increase awareness of the effects of fishing; 4. they create areas with intrinsic conservation value which are unaffected by discarding and other activities; 5. they can protect the genetic structure of fish stocks and provide insurance against management failure. Reserves also have the potential to enhance fishery yields, but such effects are highly dependent on the size of the reserve and the mobility of the fish species under consideration. It would be unwise to advocate the introduction of reserves as a way of enhancing exploitable fish yields at the present time. There has been limited study of this effect in a few small reserves and the results are equivocal. Indeed, there is still some debate as to whether overall fish production in an area is higher if part of that area is closed to fishing (McClanahan and Kaundaarara, 1996; Russ and Alcala, 1996a). When an area is designated as a reserve and effectively managed, it is reasonable to assume that the biomass of some fishes in the reserve will increase. Numerous small marine reserves (1 to 10s km2) have now been established in the tropics and the biomass of relatively site-attached species is consistently higher in these reserves than in adjacent fished areas (Russ, 1985; Alcala, 1988; Alcala and Russ, 1988; Clark et al., 1989; Bohnsack, 1990, 1996; Roberts and Polunin, 1991, 1992, 1993; Polunin and Roberts, 1993; Watson and Ormond, 1994; Harmelin et al., 1995; Jennings et al., 1996b; Russ and Alcala, 1996b). Many of the species that increase in biomass are grouper and snapper species which are particularly susceptible to fishing. In temperate waters, a larger proportion of the fished species tend to be migratory and increases in the biomass of these fishes are likely to be less marked, even though reserves provide effective protection for some rockfishes and other non-migratory species that are very vulnerable to exploita-

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tion (Rowley, 1994; Dugan and Davis, 1993; Shackell and Martin Willison, 1995). If marine reserves are to provide insurance against stock collapse, preserve ecosystem structure and maintain genetic diversity in many fished temperate ecosystems then they will have to enclose a large proportion of the ecosystem and be used in conjunction with other fisheries management tools. For example, in the northeast Atlantic, many of the most important commercial species such as plaice (Houghton and Harding, 1976), sole (Greer-Walker and Emerson, 1990), spurdog (Vince, 199l), herring (Parrish and Saville, 1965), mackerel (Lockwood, 1988) and bass Dicentrarchus labrax (L.) (Pawson et al., 1987) make seasonal migrations over distances of 200 km, to in excess of 1000 km. Commercially fished species such as thornback ray Raja clavata L. (Anon., 1993b) and lemon sole Microstomus kitt (Walbaum) (Jennings et al., 1993) which do not undertake such extensive migrations are unusual. The problems of protecting highly migratory species with closed area management were recognized long ago and are still with us (Beverton and Holt, 1957; Daan, 1993). For migratory species, powerful and effective control of fishing effort (and hence fishing mortality) will be needed unless very large areas of the ecosystem can be closed. Moreover, even when large areas are designated as ‘protected’ there is still the problem of patrolling them. Smaller marine reserves do, however, provide an excellent means of protecting habitat from destructive fishing gears (Auster and Malatesta, 1995; Auster and Shackell, 1997) and protecting fishes at vulnerable stages in their lifecycle (e.g. Rowley, 1994; Dugan and Davis, 1993; Brown et al., 1995; Hutchings, 1995; Shackell and Martin Willison, 1995). These small reserves may not operate as desired if fishing effort cannot be effectively controlled outside the protected area or if the “protected” habitat is impacted by pollution and other anthropogenic activities. For example, small closed areas are used to protect those British estuaries which are used as nurseries by juvenile bass, but the regulations do not prevent pollution, habitat reclamation or development. An ecosystem-based approach to management is known to be necessary, but integrated protection for fishes and their habitat is almost impossible because the interests of regulatory bodies often conflict (Jennings, 1992). Thus 18 Government Departments, 88 British Parliamentary Acts and a series of new European regulations exercise authority in the marine environment around the UK (Davidson et al., 1991). 6.6. Conclusions

Low levels of fishing effort have significant effects on the diversity and structure of fish communities and fishing has most impact on habitat, genetic

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diversity, age structure, life history traits and total stock size when a previously unfished area is fished for the first time. All marine ecosystems except those in the deep sea are now fished and ecologists no longer have access to large areas where the natural structure and variability of systems can be studied. There are compelling arguments for establishing marine protected areas in order to describe the natural changes which occur in unfished marine communities and to provide insurance against management failure. In many ecosystems, exisiting understanding of the relationships between species is often too poor to permit the manipulation of ecosystem function. However, in those ecosystems where strong relationships between predator and prey dynamics have been identified, such as the Barents Sea and East African coral reefs, there is the possibility of managing ecosystem function. In practice, the management strategies which have been proposed in many marine ecosystems reflect our understanding of ecosystem function. Thus in the North Sea, where strong links between predator and prey dynamics are rarely observed, there has been greater emphasis on managing species individually and in small groups.

7. SUMMARY

Fishing, often in conjunction with environmental change, has reduced the abundance of many fish populations and driven others to economic extinction. Local populations of site-attached species, which are particularly accessible to fishers, may be extirpated by fishing. Many fishing techniques have direct effects on marine habitats and benthic fauna, but these effects occur against a background of natural disturbance and, as a result, the significance of fishing effects increases markedly with depth and environmental stability. Fishing effects are most dramatic on equatorial reefs, hard and stable substrata in temperate waters, deeper areas of continental shelves and in the deep sea. Conversely, on sandy seabeds in shallow shelf seas, wave and tidal stress is often high, and fishing may have little impact on the structure of communities that are already adapted to continued movement and the resuspension of substratum. Low levels of fishing effort may have significant effects on the diversity and structure of fish communities and the greatest effects and are most commonly observed when a previously unfished area is fished for the first time. However, once an ecosystem enters the fished state, diversity, structure and fish production tend to remain relatively stable across a wide range of fishing intensities, despite fluctuations in component species which are driven by environmental changes (recruitment) and targeted overfishing. Consequently, studies of fishing effects which begin in fished systems, and

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seek to detect changes with increasing fishing effort, will often suggest that fishing has limited effects on community structure and that the system is remarkably resilient. Only when fishing effort is so high that numerous species are depleted, or when fishers resort to habitat destructive fishing techniques, will further changes in community structure become apparent. Donor-controlled effects (prey abundance affects predator populations) are observed in those marine ecosystems where the predators are reliant upon relatively few prey species. Thus the reproductive success and abundance of birds and marine mammals is tightly coupled with fluctuations in the abundance of their preferred prey. Predatory fishes typically have plastic diets, exhibit ontogenetic shifts in diet and resort to cannibalism. As a result, fluctuations in the abundance of one or two prey species rarely have a marked effect on populations of predatory fishes. Their populations are primarily structured by recruitment variation and the direct effects of fishing. For birds and mammals, it is reasonable to state that they are indirectly affected by fishing if fishing is the cause of changes in their prey populations. There is good evidence to suggest that fishing has accelerated and magnified natural population declines in some prey fishes. Top-down control is not an important structuring force in marine ecosystems and the removal of predatory fishes rarely has significant effects on the dynamics of their prey. Rather, prey species are predominantly affected by the environment and the direct impacts of fishing. The exceptions to these generalisations occur when a few species of predator (most or all of which are fished) selectively feed upon a few species which perform a key trophic function. In these cases, the indirect effects of fishing can be dramatic. For example, the capture of sea urchin predators often leads to the proliferation of sea urchins and associated changes in habitat structure. Humans and some marine mammals do exert top-down control on the marine ecosystem but their fishing strategies may differ markedly from those of other predators. In many cases, humans employ relatively conservative fishing strategies since they are often unwilling to be flexible in their aims and target a relatively small proportion of their potential prey. Most predatory fish, conversely, are very generalist feeders, often switching to invertebrate prey or cannibalism and eating many species of fishes at different stages in their life history. Predatory fishes do not exhibit the energetically inefficient human (in most developed countries) trait of pursuing the same species and are more likely to switch prey. Fishing has direct impacts on stock abundance and accelerates and magnifies natural declines in many fish stocks following environmental change and poor recruitment. The tendency of fishers to target species in sequence as a fishery develops leads to changes in the composition of fished communities with time, but the dramatic and apparently compensatory shifts in the biomass of different species in many fished ecosystems have been driven

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predominantly by environmental change rather than the indirect effects of fishing. In most pelagic systems species replacements would have occurred, albeit less rapidly, in the absence of fishing. Fishing has been responsible for shifting coral reef ecosystems to alternate stable states, as a result of the tight predator-prey coupling between a few species of invertebrate feeding fishes and sea urchins. Discards from commercial fisheries are now an important source of food for many seabirds, and discarding has led to changes in their behaviour and population size. It is probable that significant reductions in the level of discarding would have adverse effects on some seabird populations. There is still no widely accepted procedure for investigating the effects of fishing, and many other anthropogenic activities, in the marine environment. In particular, marine scientists largely lack unfished control sites, models for the analysis of predator-prey interactions which may be transient and plastic, and a good statistical framework for hypothesis testing. There is considerable scope for making better use of novel techniques for investigating the flow of matter within ecosystems and for developing predictive models which aggregate across species. However, our existing understanding of fishing effects, while scientifically primitive, undoubtedly provides the basis for making improved management decisions. Evidence for the indirect effects of fishing and the collapse of some fisheries lead us to the usual conclusion that better management and less fishing is needed. Management at the ecosystem level is necessary but it is not necessarily compatible with attempts to maximize the annual yield from the component fish stocks. Rather, conventional fisheries management would be part of an ecosystem-based approach, where the benefits of fishing would be balanced against the need to maintain diversity, seabird and mammal populations, ecosystem service functions and the intrinsic conservation value of the ecosystem.

ACKNOWLEDGEMENTS We are very grateful to Peter Auster, Steve Blaber, Joe Horwood, Steve Lockwood, Tim McClanahan, Stuart Rogers and two anonymous referees for their helpful comments on the manuscript. We wish to thank Andrew Brierley for comments on krill ecology, John Piatt for providing information on seabird and capelin interactions, Jeremy Collie and Adriaan Rijnsdorp for providing figures and John Cotter, John Reynolds and Nick Polunin for support. The senior author is funded by the European Community.

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by juvenile fishes. Canadian Journal of Fisheries and Aquatic Sciences 50, 20582070. Walters, C.J., Stocker, M., Tyler, A.V. and Westrheim, S.J. (1986). Interaction between Pacific cod (Gadus macrocephalus) and herring (Clupea harengus pallasz] in the Hecate Strait, British Columbia. Canadian Journal of Fisheries and Aquatic Sciences 43, 830-837. 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 , 139-1 72. Wanink, J.H. (1991). Survival in a perturbed environment: the effects of Nile perch introduction on the zooplanktivorous fish community of Lake Victoria. In “Terrestrial and Aquatic Ecosystems: Perturbation and Recovery” (0. Ravera, ed), pp. 269-275. Ellis-Horwood, New York. Wanless, S. and Harris, M.P. (1992). Activity budgets, diet and breeding success of kittiwakes Rissa tridactyla on the Isle of May. Bird Study 39, 145-154. Warren, P.H. (1990). Variation in food web structure - the determinants of connectance. The American Naturalist 136, 689-700. Warren, P.H. (1994). Making connections in food webs. Trends in Ecology and Evolution 9, 136141. Wassenberg, T.J. and Hill, B.J. (1990). Partitioning of material discarded by prawn trawlers in Moreton Bay. Australian Journal of Marine and Freshwater Research 41, 27-36. Watson, M. and Ormond, R.F.G. (1994). Effects of an artisanal fishery on the fish and urchin populations of a Kenyan coral reef. Marine Ecology Progress Series 109, 115-129. Welleman, H. (1989). “De verspreiding van een aantal cacrobenthos soorten in de Noordzee”. NIOZ, Den Burg, Texel. Werren, J.H. and Charnov, E.L. (1978). Facultative sex ratios and population dynamics. Nature 171, 349-350. Westman, W.E. (1977). How much are nature’s services worth? Science, N. Y. 197, 960-964. Williams, D.M. (1986). Temporal variation in the structure of reef slope fish communities (central Great Barrier Reef): short term effects of Acanthaster planci infestation. Marine Ecology Progress Series 28. Witbaard, R. and Klein, R. (1994). Long-term trends on the effects of the southern North Sea beamtrawl fishery on the bivalve mollusc Arctica islandica L. (Mollusca, bivalvia). ICES Journal of Marine Science 51, 99-105. Witman, J.D. (1988). Effects of predation by the fireworm Hermodice carunculata on milleporid corals. Bulletin of Marine Science 42, 4 4 M 5 8 . Wright, P.J. (1996). Is there a conflict between sandeel fisheries and seabirds? A case study at Shetland. In “Aquatic Predators and their Prey” (S.P.R. Greenstreet and M. Tasker, eds), pp. 154-165. Blackwell Scientific Publications, Oxford. Wright, P.J. and Bailey, M.C. (1996). Timing of hatching in Ammodytes marinus from Shetland waters and its significance to early growth and survivorship. Marine Biology 126, 143-152. Yoshioka, Y., Wada, E. and Hayashi, H. (1994). A stable isotope study on seasonal food web dynamics in a eutrophic lake. Ecology 75, 835-846. Zann, L., Brodie, J. and Vuki, V. (1990). History and dynamics of the crown-ofthorns starfish Acanthaster planci (L.)in the Suva area, Fiji. Coral Reefs 9, 135-144.

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Zaret, T.M.and Paine, R.T.(1973). Species introduction in a tropical lake. Science, N.Y. 182, 449455. Zijlstra, J.J. (1963). Effects of recruitment fluctuations and trends in herring fisheries. Rapports et Procks- Verbaux des Rdunions, Conseil International pour I’Exploration de la Mer 154, 11-16.

A Biogeographical Perspective of the Deep-sea Hydrothermal Vent Fauna

’

Verena Tunnicliffe. Andrew

G . McArthur’*2and Damhnait McHugh3

’Department of Biology. University of Victoria. Victoria. B.C., Canada V8 W 3N5 ‘Present address: Josephine Buy Paul Center for Comparative Molecular Biology and Evolution. Marine Biological Laboratory. Woods Hole. M A 02543.1015. USA 3Depurtment of Organismic and Evolutionary Biology. Museum of Comparative Zoology. Harvard University. Cambridge. M A 02138. USA 1 . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Approach . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.2. Problems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2. Hydrothermal Vents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3.

4.

5.

6.

7.

355 356 357 358 2.1. Ridge-crest Venting .......................................... 359 359 2.2. Habitat Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 361 2.3. Habitat Character . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 363 2.4. Composition of the Vent Fauna . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Other Related Faunas . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 379 382 3.1. Seeps . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 384 3.2.Organic Remains . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 385 3.3. Other Sites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Biogeography of Faunas ...................................... 385 385 4.1. Distribution of Vents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Regional Studies on the Relations of Vent Faunas . . . . . . . . . . . . . . . . . . .386 392 4.3. SeepNhaleNent Relations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Local to Regional-Scale Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 394 395 5.1. Reproductive Strategies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 395 5.2. Dispersal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 396 5.3. Gene Flow . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Regional t o Global-Scale Processes ................................. 398 398 6.1. Ridge Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 399 6.2. Vicariance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Distribution Patterns of Taxa ................................... 402 402 7.1.Generidspecific Distributions .................................. 404 7.2. Pogonophorans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

ADVANCES IN MARINE BIOLOGY VOL . 34 ISBN 0-12-026134-0

Copyright 0 1998 Academic Press Limited A / / rights ojreproduction in any form reserved

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7.3. Polychaete Family Alvinellidae. ................................. 7.4. Polychaete Family Ampharetidae. 7.5. Gastropoda ................................................ 7.6. Bivalves.. ................................................. 7.7. Copepods 8. Patterns in Diversity 8.1. Gradients.. ................................................ 8.2. Global Differences.. ......................................... 8.3. Causes .................................................... 9. Summary ..................................................... Acknowledgements. ............................................. References

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410 414 415 418 419 420 421 422 423 426 426 426

ABSTRACT Biogeography seeks to distinguish patterns in the distribution of species and to determine causal processes. Hydrothermal vent habitats have several properties that invite biogeographic studies: constrained to active deep-sea ocean ridges, known in most oceans and anticipated in the rest, patchy in distribution, extreme conditions and a limited group of inhabitants. Biologists have studied 30 vent sites mostly in the Pacific and Atlantic. Currently, 443 invertebrate species are known to generic level although many more are under study. Additionally, 32 octopus and fish species are observed in and around vents. The faunas of other sulphide-rich deep-water habitats such as margin cold seeps and organic masses (wood, carcasses) do not show great affinity at the species level to the vent fauna but the higher taxonomic affiliations suggest close evolutionary ties for many groups. Many studies address the formation of regional faunas using the firstknown sites on the Galapagos Rift and northern East Pacific Rise as the major sites of comparison. Physical disjunction of ridge crests is a likely factor in promoting the extensive provinciality that currently exists. Nonetheless, faunas on two sides of the Pacific and in the Atlantic have closer relations to each other than to the nearby “normal” deep-sea fauna. At the individual ridge scale, extensive gene flow among separated populations occurs in many species and serves to maintain the regional species pool. However, major discontinuities between major ridges reduce or eliminate gene flow; vicariant processes appear to be important. The role of differing rates of spreading in different provinces and the concomitant effects on vent habitats and faunas need further investigation. Over 75% of vent species occur at only one site and none occur at all sites. Examination of the reproductive characters of some widespread species reveals no special dispersal strategy nor does reproductive strategy predict the extent of distribution. Vestimentiferan tubeworm species are highly

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endemic and found only at Pacific vents; their limited spread may be a result of recent entry into the habitat. Alvinellid polychaete cladogenetic pattern does not match geographic regions, indicating independent penetration by numerous lineages. Endemicity among vent gastropods is high with over 60% of genera limited to this habitat; many affiliations with other sulphide-rich habitats can be identified. Among the vent gastropods there are some lineages that may have entered the vent habitat in the Mesozoic. Hydrothermal vents provide a good testing ground for processes that control patterns in diversity. Ecological and historical controls at both local and regional scales can be discerned with further study.

1. INTRODUCTION

Biogeography is an old science. Natural history works dating back to Buffon in the 18th century focus on the distribution patterns of organisms. Initial observations invoked the concept that different areas of the world held different species despite apparent similarities in environmental conditions. Many insights into evolutionary puzzles on our planet, including continental drift, sprang from such documentation. Hallam (1994) sketches the tortuous constructs of natural historians trying to explain disjunct distributions using a pre-drift “fixed” world. In retrospect, it seems amazing that Wegener’s (1924) deductions of shifting continents met with such summary rejection from geologists, while gaining acceptance from biologists. The papers by Wegener were included in many Old World zoology courses from 1928 to 1948, but the shift toward compartmentalized science hampered the integration of most lines of evidence for plate tectonics until the 1960s. Biogeography seeks to determine and explain the forces that shape biotic patterning: an end in itself. These explanations have sought several paths in the history of the science from small-scale community-level processes to large-scale historical processes such as shifting plates and changing climates (Figure 1). There are many approaches to the discipline: using only extant organism distributions or incorporating fossil information, using select phylogenetic lines or whole assemblages, using a multitude of faunas or only one type; Myers and Giller (1988), Cox and Moore (1993) and Morrone and Crisci (1995) present many of these aspects. However, biogeographic studies also provide information for other work in phylogenetic relations, reproductive ecology and community patterning. The biogeographic approach can often maintain an important overall perspective for detailed work on component parts of an ecosystem or biome.

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/

\

Figure 1 Types of biogeographical approaches. There are two principal approaches: one explains distributions through ecological processes that operate on small scales; and the other focuses on historical processes over larger scales. Studies using the latter approach usually invoke either the role of individuals or the shifting of barriers that affect whole faunas.

Work at hydrothermal vents is at a young stage due to the short history of exploration and to the relative inaccessibility of the habitat. The database to formulate a biogeographic treatment is small, so why treat the subject at all - especially in a review? The main reason is that the specialized nature of the hydrothermal vent habitat and the geographic isolation on oceanic spreading ridges has focused attention on distribution patterns since the early 1980s. Another is that a review is a useful step on the path of biological research of hydrothermal vents. It is our intent to present the present state of information along with some novel analyses to highlight intriguing issues in current research at hydrothermal vents. 1.1. Approach

This review is an unabashed mixture of a variety of approaches to biogeography. Such a presentation reflects the mix of practitioners and types of information available. One objective of the paper is to present known information in assemblies that extend previous presentations or reorganize for a new look at old data. Morrone and Crisci (1995) outline five differing methodologies used in biogeographic studies. They underscore the need to reach beyond the narrative approach (extensively used here) and apply analytical methods that allow concepts to be tested in a predefined fashion rather than resting on belief structures. To date, comparisons of taxon distributions at

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vents have taken a descriptive approach with some similarity analyses of whole assemblages. A major challenge for the science is to adopt a cladistic and/or phylogenetic approach. However, much of the work on systematic and phylogenetic relationships among (and beyond) vent organisms is either underway or not yet attempted. While some of the groundwork has been laid in global relations of vent faunas, little attempt has been made to “read the stories” embedded in the distribution of individual taxonomic groups. Each could have a rather different tale to tell. A previous paper (Tunnicliffe et al., 1996) examines modern faunal distribution in relation to plate tectonic histories. Here, we present biological details that may deserve further exploration.

1.2. Problems

Vent biology is difficult to study: an in situ presence is required. Collections are made by submersible, manned or remotely operated. Much of the global ridge crest is unexplored as many areas are remote from the agencies engaged in vent research. Research effort in the studied areas is hghly uneven. All we have so far is a very patchy picture. Exciting discoveries surely await, in the southern oceans in particular. Significant opportunities still exist, even within many extant collections. The resources required just to sort, catalogue and disseminate material are non-trivial. There is much interest in most groups from vents, but some receive far more attention than others: vestimentiferans, molluscs, larger crustaceans and some polychaetes for example. For other taxonomic groups it is difficult even to find a specialist. Although systematics is one of the major foundations of biology, it is no longer attracting numerous new practitioners. Molecular biology is providing solutions for several taxonomic problems, but until the practitioners are willing to sequence all new finds, and the allied taxa from other habitats, comparative morphology will remain the mainstay of systematics. Inconsistent collection techniques plague attempts at comparative vent studies. It is extremely difficult to take quantitative samples that capture all animals given the time available, room in collection baskets, the large size ranges of the animals involved and their attachment to hard substrata (but see Van Dover, 1995, for one approach). Thus many diversity measures are not appropriate and even species lists may be unrepresentative. Table 1 presents the location of the sites for which biological observations are available. However, these observations range from a single species description to intensive investigation by dozens of biologists for 15 years.

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Table 1 Primary sites of hydrothermal vent biological collections.

Site

Location

Ridge

Region"

Amsterdam-St. Paul Plateau Izena Hole and Iheva Minami-Ensai Bonin Seamount Mariana fields Manus field Edison Seamount Fiji fields

41" 15's 79" 06'E

Southeast Indian

IND

27" 30" 127" 05'E 28" 24' N 127" 38'E 26" 42" 141" 05'E 18" 12" 144" 43'E 3" 10's 150" 17'E 3" 19's 152" 35'E 16" 49's 179" 29'E 16" 59's 179" 55'E 22" 13-22's 176" 37'E 22" 22's 176" 43'E 49" 46" 130" 16'W 48" 27" 128" 42'W 47" 57" 129" 06'W 46" O'N 130" O'W 44" 39-57" 130" 14-22'W 42" 45" 125" 43'W 41" OO'N 127" 30'W 27" 01" 111" 24'W 20" 51" 109" 04'W 12" 4650" 103" 57'W 11" 26" 103" 47'W 9" 50" 104" 17'W 00" 48" 86" 08'W 17" 25's 113" IO'W 18" 36's 113" 24'W 37" 51" 30 02'W 37" 18" 32" 16'W 36" 11'N 33" 57'W 29" 10" 43" 1O'W 26" 08" 44" 49'W 23" 22" 44" 57'W 14" 45" 44" 58'W

Okinawa Trough Okinawa Trough Bonin Trough Mariana Trough Manus Basin New Ireland Forearc North Fiji Basin

MBJ MBJ MBJ MBJ LFM LFM LFM

Lau Basin

LFM

Explorer Juan de Fuca Juan de Fuca Juan de Fuca Juan de Fuca

NEP NEP NEP NEP NEP

Gorda Gorda East Pacific Rise East Pacific Rise East Pacific Rise East Pacific Rise East Pacific Rise Galapagos Rift East Pacific Rise East Pacific Rise Mid-Atlantic Ridge Mid-Atlantic Ridge Mid-Atlantic Ridge Mid-Atlantic Ridge Mid-Atlantic Ridge Mid-Atlantic Ridge Mid-Atlantic Ridge

NEP NEP nEPR nEPR nEPR nEPR nEPR GAL sEPR sEPR ATL ATL ATL ATL ATL ATL ATL

Lau fields South Explorer field Middle Valley Endeavour Axial Seamount Cleft Segment G14 Escanaba Trough Guaymas Basin 21"N 13"N 11"N 9"N Galapagos fields 17"s 18"s Menez Gwen Lucky Strike Rainbow Broken Spur TAG Snake Pit Logatchev

"See Figure 2 for definitions.

2. HYDROTHERMAL VENTS

The definition of a hydrothermal vent is somewhat loose. Here we adopt a fairly restrictive one that is partly defined by the animals themselves. This approach is tautological, but it represents how most ecosystem definition is made. The vent faunas discussed here are restricted to spreading ridges, both mid-ocean and back-arc basins. Pertinent habitats not included are hot-

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spots such as Hawaii (Karl et al., 1989) and shallow geothermal systems such as those off Kamchatka (Zhirmunsky and Tarasov, 1990). Apart from one shrimp, these sites do not contain animals related to ridge vent inhabitants. That is not to say that more relationships will not be found in the future. Indeed, further study of such sites may be very important to the understanding of just what limits vent animals in terms of aspects such as depth or chemical source. 2.1. Ridge-crest Venting

On ridge crests, heated water emerges at restricted sites controlled by the underlying disposition of basalts and fractures. Undiluted water in concentrated jets can be as hot as 400°C and may form large chimneys as the dissolved load of minerals is precipitated. However, subsea-floor dilution can occur and much hydrothermal fluid emerges as warm (540°C) fluids diffusing through basalts and sulphide precipitates over areas up to several hundreds of metres square. The fluids are laden with many compounds including sulphides, metals, carbon dioxide and methane. The relative concentrations depend upon fluid source and subsurface interactions (Seyfried and Mottl, 1995). Nevertheless, vent waters throughout the world have a great similarity from the point of view of organisms; the major controls on composition are seawater and host rock (basalt) chemistry (Von Damm, 1995) neither of which shows gross differences around the globe. Biologically important factors such as heat and dissolved concentrations of sulphides, iron and manganese show far more change in the few years after an eruptive event on a single ridge segment than among sites around the world (Butterfield et al., 1997). Reviews of the biology of vents are available (Grassle, 1986; Tunnicliffe, 1991; Childress and Fisher, 1992; Lutz and Kennish, 1993; Childress, 1995; Jollivet, 1996) as are some recent compilations of all aspects of vents (Humphris et al., 1995; Karl, 1995; Parson et al., 1995; Cann et al., 1997). The distribution and extreme nature of the vent habitat are the features most pertinent to biogeography. 2.2. Habitat Distribution

Hydrothermal vents are found on active spreading ridges which are generally disposed in a linear fashion around the globe (Figure 2). On the global scale, the linear disposition has marked discontinuities. Connection through the southern oceans is interrupted by South America. There is much potential for regionalization of faunas through the northern oceans where most of

W 1

9 E

120

150

1

180

1

1

1

150

1

1

1

1

1

30

1

1

1

1

1

1

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Figure 2 Distribution of major vent sites around the world. Each dot may incorporate several vent fields. The initials refer to generalized regions discussed in the text: IND = southeast Indian Ridge, LFM = Lau, Fiji, Manus, Woodlark and Lihir sites; MBJ = Marianas, Bonin and Okinawa sites; NEP = Explorer, Juan de Fuca and Gorda Ridges; nEPR = 9" through 21"N East Pacific Rise (including Guaymas Basin); GAL = Galapagos Rift; sEPR = 17"s through 25"s;ATL = mid-Atlantic Ridge sites.

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the vent sites are known. The connectivity along the mid-ocean ridges is quite high compared to the back-arc basins of the western Pacific. There, ridges are short and barriers include subduction zones, islands and interposing plates. Active hydrothermalism occupies only a small portion of the spreading ridges; thus, the available habitat occurs at irregular intervals (Figure 3). The interval between vents depends on the nature of both volcanism and tectonism of that ridge. Vent habitats are “islands” in a deep-sea biome otherwise dependent upon photosynthetic productivity from the ocean surface. We look for vents almost exclusively within the axial valley of a ridge. Often, despite what seems like saturated coverage of a vent site, more vent openings are found with greater search efforts. The extent of our ignorance is still difficult to judge. For example, off-axis venting may be more prevalent that currently appreciated. Water does circulate on a large scale through the flanks of ridges (Davis et al., 1989). Off-axis venting with a few vent animals is known (Hekinian and Fouquet, 1985). Temporal stability and spatial patterning are both related to underlying magmatism and rock structure (Fornari and Embley, 1995). This dependency affects the predictability and stability of an isolated habitat. The habitat is neither permanent nor contiguous; dispersal and migration are the major links between neighbouring vents. 2.3. Habitat Character

Hydrothermal vents present some of the most unusual conditions for animals on this planet. The vent community is dependent upon microbial production that derives from autotrophic carbon fixation using reduced compounds in the vent fluids: chemosynthesis. Most producers, including the symbiotic microbes in animals, are sulphide oxidizers although methanotrophs are known (Distel et al., 1995). Because oxidation occurs very rapidly as vent water mixes with seawater, there is a narrow zone around the vents that provides suitable access to both sulphide and oxygen. Studies, including those by Johnson et al. (1986; 1988) and Martineu and Juniper (1997) document the spatial and temporal variability of these compounds over scales of seconds and centimetres. The high spatial variability in temperature and chemical composition (Johnson et al., 1988; Butterfield et al., 1994; Massoth et al., 1994) is an important character of a vent field. Little is known about animal responses to such variability but the great heterogeneity seen in animal abundances around vents likely relates to this phenomenon (Hessler et al., 1988; Grehan and Juniper, 1996). A specialized fauna has adapted to exploit the microbial production and to cope with extreme physico-chemical influences. The nature of these are

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U

2

1

3

6

4

7

0

Distance (km)

0 /

Distance (km)

0

0

\

\

\ \ \ \

5

10

16

Distance (km) Figure 3 Distribution of vents on a ridge to illustrate patchy nature of the habitat: (a) a cross-section of a ridge crest shows the axial valley usually bounded by faulted walls; vents are often located near a central fissure in the axial valley but water circulation can also be directed through faults near the walls; (b) a longitudinal section of the north East Pacific Rise includes several segments which tend to be higher in the central portion; vent fields (triangles) are disposed unevenly along segments; (c) one vent field may consist of numerous fluid openings in the axial valley but fluid flow is not spatially contiguous. (Reproduced from Van Dover, 1990, with permission from Elsevier Science.)

detailed in the reviews mentioned above. An important point for biogeographers, however, is that, because of the rigours of the habitat, the vent environment is populated by animals rarely found elsewhere. Limiting factors within the habitat include the steep gradients and extreme values for

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both temperature and concentrations of sulphide and metals. Behavioural, physiological, morphological and reproductive adaptations have been documented, including major reorganization of internal tissues and physiologies to house microbial symbionts (Fisher, 1990), biochemical adaptations to cope with sulphide poisoning (Arp and Childress, 1983; Powell and Somero, 1983; Juniper, 1994), behavioural and molecular responses to high temperature (Dixon et al., 1992; Chevaldonne and Jollivet, 1993), presence of metal-binding proteins (Cosson-Manevy et al., 1986; Nisbet and Fowler, 1996) and development of specialized sensory organs to locate hot chimneys (Van Dover et al., 1989), among others. The accumulation of polymetallic sulphides precipitated from hot water is a unique geological feature of hydrothermal vents. This complex of minerals can accrete rapidly and form large deposits within an axial valley; it is through sulphide chimneys that high temperature (over 200OC) venting occurs. Most ridge crests are bare of sediment so basalt and these sulphide deposits constitute the major substrata. In fluid-rich areas, space competition appears to be an important factor in community structuring (Fustec et al., 1987; Hessler et al., 1988). The variable flow through the sides of chimneys can support a complex mosaic of animal patches (Tunnicliffe and Juniper, 1990; Sarrazin et al., 1997), including a few species adapted to life at the top of chimneys (Chevaldonne and Jollivet, 1993; Juniper and Martineu, 1995). At a few sites on mid-ocean ridges, fluid emerges through an overlying blanket of sediment: at Middle Valley, Juan de Fuca Ridge, Escanaba Trough, Gorda Ridge and Guaymas Basin, northern East Pacific Rise (Table 1). The fluid chemistry at these sites is different, with lower maximum temperatures, metals mostly stripped and some methane present. The presence of sediment provides a substratum that can attract another suite of organisms. The sediment also probably redistributes the subsurface fluid flows (Grehan and Juniper, 1996). At several back-arc basin sites, sediments intermingle with basalt and sulphide deposits (e.g. Manus Basin, Lau Basin). The mix of substrata may be important for the ecological component of species distributions. 2.4, Composition of the Vent Fauna

2.4.1. Taxon Listings Collecting and identifying taxa are among the obvious tasks for the vent ecologist. The unusual nature of the vent animals has made this task interesting and relatively easy to attract the attention of systematists. Species lists have grown rapidly. Newman (1985) lists 58 species, Tunnicliffe (1991) lists

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236 species and here we present 443 species. Desbruykres and Segonzac (1997) present a similar listing with useful illustrations of many species. Tables 2 and 3 present the species currently known from hydrothermal vents. This data assembly represents a conservative approach on our part. We separate invertebrates (Table 2) from octopus plus fish listings (Table 3) because it is difficult to be sure which of the vagile organisms can be considered as “inhabitants” and which are fortuitous sightings in the vicinity of vents. We use two criteria for inclusion in the invertebrate listing. First, we include only the species published or that we could confirm to generic level. Many more taxa are presented in site listings to ordinal or family level but with genus yet unknown. Inclusion of these would increase the listing by nearly a hundred “species”, but would contribute almost nothing to biogeographic comparisons. But worse, unconfirmed citations can confuse future systematic work. Secondly, it can be difficult to determine if a species was really part of the vent community. If the author says it was, we accept it. If there was a reasonable probability of contamination, the species was left out. Contamination can occur as an exposed sample comes up through open ocean plankton assemblages; sometimes, rock samples from beyond the venting area will contribute a few organisms to the sample tray. Description of vent taxa is ongoing; the listing will be outdated by press time. In particular, a paper, currently in preparation by A. War& and P. Bouchet (pers. comm.) describes new and revises old gastropod taxa. 2.4.2. Phy let ic Composition While the vent habitat is extremely rich in autotrophic production, the inhabitants appear to have tied themselves to the habitat in evolutionary time; only a few are able to venture into the surrounding deep-sea or to colonize other sulphide-rich habitats. Despite the high productivity, the communities are low in diversity and trophic structure appears simpler than in other marine communities although relations are not easy to interpret (Tunnicliffe, 1991; Fisher et al., 1994; Southward et al., 1994; Van Dover and Fry, 1994). Van Dover (1990) and Tunnicliffe (1992) comment on the preponderance of three phyla found at vents: molluscs, arthropods and annelids (Figure 4). Most vent collections come from hard substratum habitats; a comparison with bare ridge crest faunas shows a marked difference in the resident invertebrate taxa at the phylum level (Figure 5). Comprehensive data from ridges are rare due to poor recovery but both Carey et al. (1990) and Copley et al. (1 996) document the predominance of cnidarians, sponges and echinoderms. Figure 5 also presents phyletic composition from Grassle and Maciolek’s (1 992) deep-sea sediment survey between 1400 and 2500 m

365

DEEP-SEA HYDROTHERMAL VENT FAUNA

Table 2 Names and locations of the known hydrothermal vent invertebrates. Taxa are listed only if generic identifications (actual or acknowledged as known or new) could be confirmed by us; thus many collected specimens remain unlisted; unk = unknown. The last column indicates the geographic region and the endemic status of the species: 1 =northeast Pacific, 2 = Guaymas Basin on East Pacific Rise, 3 = 21"N East Pacific Rise, 4 = 9-13"N East Pacific Rise, 5 = Galapagos Rift, 6 = 23-27"s East Pacific Rise, 7 = Fiji, Lau, Manus or Lihir Back-Arc Basins, 8 = Mariana, Bonin or Okinawa Back-Arc Basins, 9 = Mid-Atlantic Ridge; IND = southeast Indian Ridge; Y= endemic to hydrothermal vents, N = also known from the deep sea, S = also known at seeps, W = also known from whale carcasses, ? = status elsewhere unknown. Neither Protozoa nor Nematoda are included in this listing. Cnidaria Hydrozoa Hydroida Candelabridae

Candelabridae Siphonophora Rhodalidae Anthozoa Actiniaria Actinostolidae Actinostolidae Actinostolidae Actinostolidae Actinostolidae Cerinathidae Hormathiidae Scyphozoa Lucernariida Lucernariidae

Candelabrum phrygium Segonzac and Vervoort Candelabrum serpentarii Segonzac and Vervoort Thermopalia taraxaca Pugh Actinostola callosa Verrill Parasicyonis ingolfi Carlgren Cyanathea hydrothermala Doumenc and Van Praet Marianactis bythios Fautin and Hessler Gen n. sp. (D. Fautin, pers. comm.) Cerianthus sp. (in Desbruyeres et al., 1997) Chondrophellia coronata Verrill

4,N 9 3 4.Y

4.N

Lucernaria n. sp. (R. Larson, pers. comm.)

Nemertea Enopla Hoplonemertea cf. Emplectonematidae Thermanemertes valens Gibson, Rogers and Tunnicliffe Acanthocephala Palaeacanthocephala Echinorh ynchida Hypoechinorhynchidae Hypoechinorhynchus thermaceri de Buron Pogonophora Perviata Thecanephria Siphonobrachia lauensis Southward Lamellisabellidae 7,y

V. TUNNlCLlFFE ET AL.

Table 2 (continued) Obturata Axonobranchia Riftiidae Basibranchia Alaysiidae Arcovestiidae Lamellibrachiidae Lamellibrachiidae Ridgeiidae Tevniidae Tevniidae Escarpiidae unk Annelida Polychaeta Amphinomida Archinomidae Capitellida Maldanidae Maldanidae Maldanidae Maldanidae Eunicida Dorvilleidae Dorvilleidae Dorvilleidae Dorvilleidae Dorvilleidae Dorvilleidae Eunicidae Eu nicid ae Eunicidae Opheliida Scalibregmatidae Orbiniidae Orbiniidae Orbiniidae Orbiniidae Phyllodocida Glyceridae Hesionidae Hesionidae Hesionidae

Riftia pachyptila Jones Alaysia spiralis Southward Arcovestia ivanovi Southward and Galkin Lamellibrachia barhami Webb Lamellibrachia columna Southward Ridgeia piscesae Jones Oasisia alvinae Jones Tevnia jerichonana Jones Escarpia spicata Jones n. sp. (in Hashimoto et al., 1995)

Archinome (Euphrosine) rosacea (Blake ) Nicomache arwidssoni Blake Nicomache venticola Blake and Hilbig l,Y Capitella sp. (in Hashimoto et al., 1995) 8,? Capitella nr capitata (in Grassle et al., l,? 1985) Exallopus jumarsi Blake Ophryotrocha akessoni Blake Ophryotrocha globopalpata Blake and Hilbig Ophryotrocha platykephale Blake Ophryotrocha sp. (in Hashimoto et al., 1995) Parougia worfi Blake and Hilbig Eunice masudai Miura 1986 Eunice northioidea Moore 1903 Eunice pulvinopalpata Fauchald

2,y 2,4,5,Y 1 ,Y 2,Y 8,? 1 ,Y 8,N 8,N 3,Y

Axiothella millsi Pocklington and Fournier Leitoscoloplos pachybranchiatus Blake and Hilbig Orbiniella aciculata Blake Orbiniella hobsonae Blake and Hilbig Scoloplos ehlersi Blake

1, N

5,Y 1,Y 5,y

Glycera profundi Chamberlin Amphiduros axialensis Blake and Hilbig Hesiocaeca hessleri Blake Hesiodeira glabra Blake and Hilbig

2,N l,N 8,Y 1,Y

1,Y

367

DEEP-SEA HYDROTHERMAL VENT FAUNA

Hesionidae Hesionidae Hesionidae Hesionidae Lacydoniidae Nautiliniellidae Nautiliniellidae Nau tiliniellidae Nau tiliniellidae Nautiliniellidae Nereidae Nereidae Nereidae Nereidae Phyllodocidae Phyllodocidae Phyllodocidae Ph yllodocidae Polynoidae Polynoidae Polynoidae Pol ynoidae Pol ynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Pol ynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Pol ynoidae Polynoidae Polynoidae Polynoidae

Hesiolyra bergi Blake Hesiospina vestimentifera Blake Nereimyra alvinae Blake Orseis grasslei Blake Lacydonia n. sp. J. Blake, pers. comm. Iheyomytilidicola tridentatus Miura and Hashimoto Mytilidiphila enseiensis Miura and Hashimoto Mytilidiphila okinawaensis Miura and Hashimoto Shinkai longipedata Miura and Ohta Shinkai semilonga Miura and Hashimoto Ceratocephale pacijica Hartman Nereis piscesae Blake and Hilbig Nereis sandersi Blake Nereis sp. unpub. data (VT) Eulalia (Protomystides) papillosa Blake Galapagomystides aristata Blake Protomystides verenae Blake and Hilbig Protomystides sp. (in Gebruk et al., 1997) Bathycatalina jilamentosa Pettibone Bathykurila guaymasensis Pettibone Branchinotogluma burkensis Pettibone Branchinotogluma grasslei Pettibone Branchinotogluma hessleri Pettibone Branchinotogluma sandersi Pettibone Branchiplicatus cupreus Pettibone Branchipolynoepettiboneae Miura and Hashimoto Branchipolynoe seepensis Pettibone Branchipolynoe symmytilida Pettibone Harmothoe macnabi Pettibone Iphionella risensis Pettibone Iphionella n. sp. Galkin Lepidonotopodiumjimbriatum Pettibone Lepidonotopodium minutum Pettibone Lepidonotopodium piscesae Pettibone Lepidonotopodium riftense Pettibone Lepidonotopodium williamsae Pettibone Levensteiniella intermedia Pettibone Levensteiniella kincaidi Pettibone Levensteiniella raisae Pettibone Levensteiniella sp. (in Gebruk et al., 1997) Macellicephala galapagensis Pettibone Macellicephaloides alvini Pettibone

3AY 1,3,5,S 2,5,Y 1,2,y l,Y 8,Y 8,Y 8,Y 8,Y 8,Y 2 3 1,Y 2,3,4,5,6,Y 7,? 3,Y 5,Y l,Y 2,? 1 ,N 2,W 8,Y 1,2,3,4,5,Y 3,5,Y 1,2,3,5,Y 2,3,4,Y 7,8,Y

9,s 4,5,Y 1S , Y 3,y 7,Y 3,4,Y 8,Y 1 ,Y 2,3,5,Y 3,4,5,Y 1,Y 1,3,5,Y 8,Y 2,Y 5,Y 2,Y

V. TUNNICLIFFE ET AL.

Table 2 (continued) Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Polynoidae Sigalionidae Syllidae Sabellida Serpulidae Serpulidae Sp ionida Cirratulidae Cirratulidae Spionidae Spionidae Spionidae Spionidae Spionidae Spionidae Spionidae Terebellida Alvinellidae Alvinellidae Alvinellidae Alvinellidae Alvinellidae Alvinellidae Alvinellidae

Opisthotrochopodus alvinus Pettibone Opisthotrochopodus japonicus Miura and Hashimoto Opisthotrochopodus marianus Pettibone Opisthotrochopodus segonzaci Miura and Desbruyires Opisthotrochopodus trifurcus Miura and Desbruyeres Opisthotrochopodus tunnicliffeae Pettibone Thermiphione jijiensis Miura Thermiphione tufari HartmannSchroeder Thermopolynoe branchiata Miura Neolaenira racemosa Fauchald Sphaerosyllis ridgensis Blake and Hilbig Laminatubus alvini Ten Hove and Zibrowius Protis hydrothermica Ten Hove and Zibrowius Chaetozone n. sp. 1 (Blake, J., pers. comm.) Chaetozone n. sp.2 (Blake, J., pers. comm.) Laubieriellus grasslei Maciolek Lindaspio dibranchiata Blake Lindaspio southwardorum Blake and Maciolek Prionospio sandersi Maciolek Prionospio sp. Blake and Hilbig Xandaros acanthodes Maciolek N . gen. n. sp. (Blake, J., pers. comm.) Alvinella caudata Desbruyires and Laubier Alvinella pompejana Desbruyeres and Laubier Paralvinella bactericola Desbruyeres and Laubier Paralvinella dela Detinova Paralvinella fjiensis Desbruyeres and Laubier Paralvinella grasslei Desbruyeres and Laubier Paralvinella hessleri Desbruyeres and Laubier

2,3,5,Y 8,Y 8,Y 7,Y 7,Y 1,Y

7,y 6,Y 7,y 2,N 1,Y

369

DEEP-SEA HYDROTHERMAL VENT FAUNA

Alvinellidae Alvinellidae Alvinellidae Alvinellidae Ampharetidae Ampharetidae Ampharetidae Ampharetidae Hirudinea Rhynchobdella Piscicolidae Mollusca Polyplacop hora Chitonida Ischnochitonidae Ischnochitonidae Aplacophora Neomeniomorpha Chaetodermatidae Simrothiellidae Simrothiellidae Simrothiellidae Simrothiellidae Simrothiellidae Simrothiellidae Simrothiellidae Gastropoda Neomphalina Cyathermiidae Cyathermiidae Neomphalidae Neomphalidae Neomphalidae

Paralvinella palmiformis Desbruyeres and Laubier Paralvinella pandorae Desbruyeres and Laubier Paralvinella sulfincola Desbruyeres and Laubier Paralvinella unidentata Desbruyeres and Laubier Amathvs lutzi Desbruyeres and Laubier Amphhamytha galapagemis Zottoli

1 ,Y 1,4,Y

1 ,Y 1,Y 9,Y

Amphisamytha fauchaldi Solis-Weiss and Hernandez-Alacantara Grassleia hydrothermalis Solis-Weiss Bathybdella sawyeri Burreson

Leptochiton tenuidontus Saito and Okutani Thermochiton undocostatus Saito and Okutani Falcidens n. sp. (in Scheltema, 1990) Helicoradomenia juani Scheltema Helicoradomenia n. sp. 1 (in Scheltema, 1990) Helicoradomenia n. sp. 2 (in Scheltema, 1990) Helicoradomenia n. sp. 3 (in Scheltema, 1990) Helicoradomenia n. sp. 4. (in Scheltema, 1990) Helicoradomenia n. sp. 5 (in Scheltema, 1990) Helicoradomenia n. sp. 6 (in Scheltema, 1990) Cyathermia naticoides Waren and Bouchet Lacunoides exquisitus Warin and Bouchet Neomphalus fretterae McLean Symmetromphalus hageni Beck Symmetromphalus regularis McLean

315,Y 8,Y 3,Y 4,Y 5,Y 3,Y

3,4,y

5,y 3,4,5,y 7,Y 8,Y

370

V. TUNNICLIFFE ET AL.

Table 2 (continued)

Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Peltospiridae Pel tospiridae Pel tospiridae Peltospiridae Pel tospiridae Peltospiridae Peltospiridae Neritimorpha Phenacolepadidae Phenacolepadidae Phenacolepadidae Patellogastropoda Neolepetopsidae Neolepetopsidae Neolepetopsidae Neolepetopsidae Neolepetopsidae Neolepetopsidae Neolepetopsidae

Ctenopelta porifera Warkn and Bouchet Depressigyra globulus W a r h and Bouchet Echinopelta fistulosa McLean Hirtopelta hirta McLean Lirapex granularis Waren and Bouchet Lirapex humata W a r h and Bouchet Melanodrymia aurantiaca Hickman Melanodrymia brightae Waren and Bouchet 1993 Melanodrymia n. sp. (Shank et al., 1998) Nodopelta heminoda McLean Nodopelta subnoda McLean Pachydermia laevis Waren and Bouchet Pachydermia sculpta W a r h and Bouchet Peltospira delicata McLean Peltospira lamellifera Warkn and Bouchet Peltospira operculata McLean Peltospira sp. 1 Warkn and Bouchet Peltospira sp. 2 A. War&, pers. comm. Peltospira sp. 3 (in Van Dover et al., 1996) Planorbidella depressa W a r h and Bouchet Planorbidella planispira Waren and Bouchet Rhynchopelta concentrica McLean Rhynchopelta nux Okutani, Fujikura and Sasaki

3,Y 4,Y 3,Y * 3,Y 3,4,6,Y 1,Y 4,Y 3,4,Y 4,6,Y 3,4,6,Y 7,Y

4,6,Y 4,Y

'

3,4,6,Y 7,Y 5,? 9,?

7,Y 3,4,6,Y 3,4,Y 8,Y

Shinkailepas kaikatensis Okutani Saito 7,8,Y and Hashimoto Shinkailepas tufari Beck 7,Y Olgasolaris tollmanni Beck 7,Y Eulepetopsis vitrea McLean Neolepetopsis densata McLean Neolepetopsis gordensis McLean Neolepetopsis occulata McLean Neolepetopsis verruca McLean Paralepetopsis rosemariae Beck Paralepetopsis n. sp. (J. McLean, pers.

comm.) Pseudococculinidae Cocculinidae

4,Y 1,Y

Amphiplica gordensis McLean Cocculina cf. craigsmithi ( J . McLean, pers. comm.)

DEEP-SEA HYDROTHERMAL VENT FAUNA

Pyropeltidae Pyropeltidae Pyropeltida Acmaeidae Acmaeidae Vetigastropoda Scissurellidae Scissurellidae Scissurellidae Scissurellidae Fissurellidae

Pyropelta bohlei Beck Pyropelta corymba McLean and Haszprunar Pyropelta musaica McLean and Haszprunar Bathyacmaea jonassoni Beck Bathyacmaea secunda Okutani, Fujikura and Sasaki

Skeneidae Lepetodrilidae Lepetodrilidae Lepetodrilidae

Sutilizona theca McLean Temnocinclis euripes McLean Temnozaga parilis McLean Anatoma sp. WarCn and Bouchet Puncturella parvinobilis Okutani, Fujikura and Sasaki Puncturella rimaizenaensis Okutani, Fujikura and Sasaki Puncturella solis Beck Puncturella n. sp. (J. McLean, pers. comm.) n. gen. n. sp. (J. McLean, pers. comm.) Clypeosectus curvus McLean Clypeosectus delectus McLean Pseudorimula marianae McLean Pseudorimula midatlantica McLean Bathymargarites symplector WarCn and Bouchet Fucaria n. sp. WarCn and Bouchet in ms Fucaria striata Waren and Bouchet Margarites shinkai Okutani, Fujikura and Sasaki Vetulonia phalcata WarCn and Bouchet Bruceiella globulus WarCn and Bouchet Leptogyra injlata Waren and Bouchet cf Leptogyra sp. (in Van Dover et al., 1996) Moelleriopsis n. sp. WarCn and Bouchet Protolira valvatoides Warkn and Bouchet Solutigyra reticulata WarCn and Bouchet Ventsia tricarinata WarCn and Bouchet Lepetodrilus corrugatus McLean Lepetodrilus cristatus McLean Lepetodrilus elevatus McLean

Lepetodrilidae Lepetodrilidae

Lepetodrilus fucensis McLean Lepetodrilus guaymasensis McLean

Fissurellidae Fissurellidae Fissurellidae Fissurellidae Clypeosectidae Clypeosectidae Clypeosectidae Clypeosectidae Trochidae Trochidae Trochidae Trochidae Trochidae Skeneidae Skeneidae Skeneidae Skeneidae Skeneidae Skeneidae

l,Y 1,y 3,4,5,6,Y 7,8,Y 9,Y 3,4,6,Y 1,Y l,Y 8,s 7,Y 1,Y 1,Y 9,? 4,Y 9,Y 34,y 1,Y 1,y 3,4,5,Y 3,4,5,6,7, 8,Y 1,Y 2,Y

V. TUNNICLIFFE ET AL.

Table 2 (continued) Lepetodrilidae Lepetodrilidae Lepetodrilidae Lepetodrilidae Lepetodrilidae Lepetodrilidae Gorgoleptidae Gorgoleptidae Gorgoleptidae Turbinidae unk Heterobranchia Hyalogyrinidae Hyalogyrinidae Hyalogyrinidae Xylodisculidae Caenogastropoda Rissoidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Provannidae Elachisnidae Buccinidae

Lepetodrilus japonicus Okutani, Fujikura and Sasaki Lepetodrilus ovalis McLean Lepetodrilus pustulosus McLean Lepetodrilus schrolli Beck Lepetodrilus tevnianus McLean Lepetodrilus sp. (in Van Dover et al., 1996) Gorgoleptis emarginatus McLean Gorgoleptis patulus McLean Gorgoleptis spiralis McLean Cantrainea jamsteci (Okutani and Fujikura) Helicrenion reticulatum WarCn and Bouchet Hyalogyra vitrinelloides W a r h and Bouchet Hyalogyrina grasslei War& and Bouchet Hyalogyrina n. sp. (A. Waren, pers. comm.) Xylodiscula major WarCn and Bouchet n. sp. (in Van Dover et al., 1996) Alviniconcha hessleri Okutani and Ohta Desbruyeresia cancellata War6n and Bouchet Desbruyeresia marianaensis Okutani Desbruyeresia melanioides W a r h and Bouchet Desbruyeresia spinosa W a r h and Bouchet Ifremeria nautilei Bouchet and Warkn Provanna buccinoides Waren and Bouchet Provanna glabra Okutani, Fujikura and Sasaki Provanna goniata WarCn and Bouchet Provanna ios Warkn and Bouchet Provanna laevis Waren and Ponder Provanna muricata Warin and Bouchet Provanna nassariaeformis Okutani Provanna segonzaci W a r h and Ponder Provanna variabilis W a r h and Bouchet Provanna sp. A. Warin pers. comm. Laeviphitus japonicus Okutani, Fujikura and Sasaki Buccinum viridum Dall

3.Y 5;Y 4,Y 8,Y 1,Y 1,Y 2,Y l,Y 1,Y

9,Y 1,8,Y 1,Y 8,Y 1,Y

7,Y

1,Y 7,Y 8,s 2,Y 3,4,5,6,Y 2,s 33,Y 8,Y 7,Y 1,s I,? 8,Y 1,N

373

DEEP-SEA HYDROTHERMAL VENT FAUNA

Buccinidae Buccinidae Buccinidae Buccinidae Turridae Turridae Turridae Turridae Turridae Turridae Turridae Cerithiellidae Bivalvia Eulamellibranchia Vesicomyidae Vesicom yidae Vesicom yidae Vesicomyidae Vesicomyidae Filibranchia Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae

Buccinum sp. Warkn and Bouchet Eosipho desbruyeresi Okutani and Ohta Eosipho n. sp. A. WarCn, pers. comm. Neptunea insulalis (Hashimoto et al., 1995) Oenopota ogasawarana Okutani, Fujikura and Sasaki Phymorhynchus sp. 2 WarCn and Bouchet Phymorhynchus sp. 3 Waren and Bouchet Phymorhynchus sp. 4 Waren and Bouchet Phymorhynchus warPni Sysoev and Kantor Phymorhynchus moskalevi Sysoev and Kantor Phymorhynchus starmeri Okutani and Ohta N. gen. n. sp. WarCn and Bouchet, in ms.

1,? 7,8,Y 6,Y 8,N 8,Y 9,? 1,?

4,5,? 7,Y 9,Y 7,Y l,Y

Calyptogena magnijica Boss and Turner Calyptogena solidissima Okutani, Fujikura and Sasaki C. pacijicalV. lepta complex (see Peek et al., 1997) V . gigaslc. kilmeri complex (see Peek et al., 1997) Ectenegena extenta (see Peek et al., 1997)

3,4,5,Y 8,Y

Bathymodiolus aduloides Hashimoto and Okutani Bathymodiolus brevior von Cosel, Metivier and Hashimoto Bathymodiolus elongatus von Cosel, Metivier and Hashimoto Bathymodiolus japonicus Hashimoto and Okutani Bathymodiolus platifrons Hashimoto and Okutani Bathymodiolus puteoserpentis von Cosel, Metivier and Hashimoto Bathymodiolus septemdierum Hashimoto and Okutani Bathymodiolus thermophilus Kenk and Wilson

8,s

1,2,? 1,2,W 1 ,s

7,Y 7,Y 8,s 8,s 9,Y 8,Y 4,5,Y

V. TUNNICLIFFE €1AL.

Table 2 (continued)

Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae Mytilidae Pectinidae Protobranchia Nuculaniidae Solemyidae Solemyidae Arthropoda Arachnida Acarina Halacaridae Halacaridae Halacaridae Halacaridae Halacaridae Halacaridae Pycnogonida Ammotheidae Ammotheidae Ammotheidae Ammotheidae Ammot heidae Ammotheidae Crustacea/Cirripedia Pedunculata Brachylepadidae Scalpellidae Scalpellidae Scalpellidae Scalpellidae Scalpellidae Scalpellidae Scalpellidae

Bathymodiolus n. sp. 1 Hessler and Lonsdale Bathymodiolus n. sp. 2 (R. Lutz, pers. comm.) Bathymodiolus n. sp. 3 (C. Metivier, pers. comm.) Bathymodiolid n. sp. (in Van Dover et al., 1996) Idas (Idasola) washingtonia Bernard Adipicola n. sp. (E. Southward, pers. comm.) Bathypecten vulcani Schein-Fatton

8,Y 9,Y 7,Y 9,Y 1,W,N l,Y 43,Y

Nuculana grasslei Allen 2,? Acharax alinae Metivier and van Cosel 7,Y Solemya johnsoni Dall 1 3

Bathyhalacarus sp. Bartsch Copidognathus alvinus Bartsch Copidognathuspapillatus Krantz Copidognathus nautilei Bartsch Halacarellus alvinus Bartsch Halacarellus auzendei Bartsch

9,? 9 3 1,4,5,7,Y 9,y 9,y 9,Y

Ammothea verenae Child Sericosura cochleifovea Child Sericosura cyrtoma Child and Segonzac Sericosura heteroscela Child and Segonzac Sericosura mitrata Gordon Sericosura venticola Child

1, y 8,Y 4,Y 9,Y

Neobrachylepas relica Newman and Yamaguchi Neolepas rapanuii Jones Neolepas zevinae Newman Neolepas n. sp. 1 (in Desbruyires et al., 1994) Neolepas n. sp. 2 (in Fujikura et al., 1995) Neolepas n. sp. 3 (in Southward et al., 1997) Scillaelepas n. sp. (in Newman, 1985) Gen. n. sp. (W. Newman, pers. comm.)

7,Y

9 3 1, y

6,Y 3,4,y 7,Y 8,Y IND,Y 3,Y 7,Y

DEEP-SEA HYDROTHERMAL VENT FAUNA

Sessi1ia Neoverrucidae Neoverrucidae Verrucidae Verrucidae Pachylasmatidae Pachylasmatidae Pachylasmatidae Crustacea/Copepoda Calanoida Spinocalanidae Poecilostomatoida Clausidiidae Clausidiidae Erebonasteridae Erebonasteridae Uncertain Oncaeidae Siphonostomatoida Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae

Imbracoverruca n. sp. (in Desbruyeres et al., 1994) Neoverruca brachylepadoformis Newman and Hessler Verruca n. sp. 1 (in Newman, 1985) Verruca n. sp. 2 (in Newman, 1985) Eochionelasmus ohtai Yamaguchi and Newman Eochionelasmus paquensis Yamaguchi and Newman Eochionelasmus n. sp. (in Desbruyeres et al., 1994) Isaacsicalanus paucisetus Fleminger Hyphalion captans Humes Hyphalion sp. (in Hashimoto et al., 1995) Amphicrossus altalis Humes and Huys Erebonaster protentipes Humes Laitmatobius crinitus Humes Oncaea praeclara Humes Aphotopontius acanthinus Humes and Lutz Aphotopontius arcuatus Humes Aphotopontius atlanteus Humes Aphotopontius baculigerus Humes Aphotopontius frexispina Humes Aphotopontius forcipatus Humes Aphotopontius hydronauticus Humes Aphotopontius limatulus Humes Aphotopontius mammillatus Humes Aphotopontius probolus Humes Aphotopontius temperatus Humes Benthoxynus spiculifer Humes Benthoxynus tumidiseta Humes Ceuthoecetes aliger Humes and Dojiri Ceuthoecetes acanthothrix Humes Ceuthoecetes cristatus Humes Ceuthoecetes introversus Humes Chasmatopontius thescalus Humes Cheramyzon abyssale Humes Exrima dolichopus Humes Exrima singula Humes Nilva torifera Humes Rhogobius contractus Humes

3,4,5,Y 3,4,5,y

V. TUNNICLIFFE ETAL.

Table 2 (continued)

Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Dirivultidae Ecbathyriontidae Uncertain Megapontiidae Harpacticoidea unk Crustacea/Ostracoda Mydocopida Philomedidae Mydocopida Philomedidae Podocopida Cytheruridae Cytheruridae ?

Crustacea/Pericarida Leptostraca Nebaliidae Amphipoda Amphilochidae Callopiidae Eusiridae

Rhogobius pressulus Humes Rimipontius mediospinifer Humes Scotoecetes introrsus Humes Stygiopontius appositus Humes Stygiopontius brevispina Humes Stygiopontius bulbisetiger Humes Stygiopontius cinctiger Humes Stygiopontius cladarus Humes Stygiopontius flexus Humes Stygiopontius hispidulus Humes Stygiopontius latulus Humes Stygiopontius lauensis Humes Stygiopontius lumiger Humes Stygiopontius mirus Humes Stygiopontius mucroniferus Humes Stygiopontius paxillifer Humes Stygiopontius pectinatus Humes Stygiopontius quadrispinosus Humes Stygiopontius regius Humes Stygiopontius rimivagus Humes Stygiopontius sentifer Humes Stygiopontius serratus Humes Stygiopontius stabilitus Humes Stygiopontius teres Humes Stygiopontius verruculatus Humes Ecbathyrion prolixicauda Humes Fissuricola caritus Humes Hyalopontius boxshalli Humes

n. sp. (R. Huys, pers. comm.) Euphilomedes climax Kornicker Prionotoleberis styx Kornicker

N. Gen. n. sp. (in Desbruyeres et al., 1994) Xylocythere n. sp. Van Harten Poseidonamicus sp. (in Gebruk et al., 1997) Dahlella caldariensis Hessler Gitanopsis alvina Hessler Oradarea longimana (Boeck) Bouvierella curtirama Bellan-Santini and Thurston

DEEP-SEA HYDROTHERMAL VENT FAUNA

Eusirid ae Lysianassidae Lysianassidae Lysianassidae Lysianassidae Lysianassidae Lysianassidae

Lysianassidae Lysianassidae Lysianassidae Pardaliscidae Pardaliscidae Sebidae Stegocephalidae Stegosessalidae Isopoda Criptoniscidae Tanaidacea Leptognathiidae Leptognathiidae Crustacea/Eucarida Decapoda Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Alvinocarididae Mirocarididae Mirocarididae Hippol ytidae

Luckia striki Bellan-Santini and Thurston Apotectonia heterostegos Barnard and Ingram Euonyx mytilus Barnard and Ingram Hirondellea glutonis Barnard and Ingram Orchomene abyssorum Barnard and Ingram Orchomene distinctus Barnard and Ingram Tectovalopsis diabolus Barnard and Ingram Tectovalopsis wegeneri Barnard and Ingram Transtectonia torrentis Barnard and Ingram VentGIla sulfuris Barnard and Ingram Halice hesmonectes Martin, France and Van Dover Pardalisca endeavouri Shaw Seba profundus Shaw Steleuthera ecoprophycea Bellan-Santini and Thurston Euandania aff. ingens Stebbing

2,3,4,5,Y 4,Y 1,Y 1,y 9,Y

9,y

Thermaloniscus cotylophorus Bourdon

4,Y

Leptognathia ventralis Hansen Typlotanais sp. Hansen

7 3 I,?

Alvinocaris longirostris Williams Alvinocaris lusca Williams and Chace Alvinocaris markensis Williams Alvinocaris sp. (in Desbruyeres et al., 1994) Chorocaris chacei Williams and Rona Chorocaris vandoverae Martin and Hessler Chorocaris sp. (in Desbruyeres et al., 1994) Zorania concordia Vereshchaka Rimicaris exoculata Williams and Rona Rimicaris aurantiaca Martin et al. Mirocaris (Chorocaris) fortunata Martin and Christiansen Mirocaris keldyshi Vereshchaka Lebbeus carinatus de Saint Laurent

8,Y 4,5,Y 9,y 1,Y 9,Y 7,8,Y 7,? 9,Y 9,Y 9,Y 9,y 9,Y 4,Y

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Table 2 (continued)

Hippolytidae Hippolytidae Nephropidae Oplophoridae Oplophoridae Caridea Chirostylidae Chirostylidae Chirostylidae Chirostylidae Galatheidae Galatheidae Galatheidae Galatheidae Galatheidae Galatheidae Gala theidae Galatheidae Galatheidae Bythograeidae Bythograeidae Bythograeidae Bythograeidae Bythograeidae Bythograeidae Bythograeidae Bythograeidae Geryonidae Lithodidae Lithodidae Majidae

Lebbeus washingtonianus (in Fujikura et al., 1995) Lebbeus sp. (in Desbruyeres et. al., 1994) Thimopides sp. (in Segonzac, 1992) Acanthephyra purpurea Milne-Edwards Systellapsis braueri Balss Paracrangon n. sp. (in Fujikura et al., 1995) Uroptychus bicavatus Baba and de Saint Laurent Uroptychus thermalis Baba and de Saint Laurent Uroptychus n. sp. (A. Williams, pers. comm.) N. Gen. n. sp. (A. Williams, pers. comm.) Munida magniantennulato Baba and Turkay Munidopsis alvisca Williams Munidopsis cf. crassa Segonzac Munidopsis lauensis Baba and de Saint Laurent Munidopsis lentigo Williams and Van Dover Munidopsis marianica Williams and Baba Munidopsis starmer Baba and de Saint Laurent Munidopsis subsquamosa Henderson Munidopsis sp. (in Fujikura et al., 1995) Austinograea alaysae Guinot Austinograea williamsi Hessler and Martin Bythograea intermedia de Saint Laurent Bythograea laubieri Guinot and Segonzac Bythograea microps de Saint Laurent Bythograea thermydron Williams Cyanagraea praedator de Saint Laurent Segonzacia mesatlantica Williams Chaceon af$nis Milne-Edwards and Bouvier Paralomis jamsteci Takeda and Hashimoto Paralomis sp. (in Desbruyeres et. al., 1994) Macroregonia macrochira Sakai

7,? 9,? 9 3 4 3 8.Y 7,y 7,Y 7,y 7,Y

9,? 7,Y 3,Y 8,Y 7,Y 3,4,5,N 8,? 7,Y 7,8,Y 5,Y 6,Y 3,4,Y 3,4,5,Y 3,4,Y 9,Y 9,N 8,Y 7,? 1 ,N

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DEEP-SEA HYDROTHERMAL VENT FAUNA

Echinodermata Echinoidea Echinoida Echinidae Holothuroidea Apodida Synaptidae Stelleroidea Asteroida Goniasteridae Ophiurida Ophiuridae Hemichordata Enteropneusta Saxipendiidae

Echinus alexandri Danielssen and Koren

9 3

Chirodota n. sp. (in Desbruyeres et al., 7,Y 1994) Cerarnaster misakiensis (in Hashimoto et al., 1995)

8,N

Ophioctenella acies Tyler et al.

9,Y

Saxipendiurn coronaturn Woodwick and 4,5,Y Sensenbaugh

off the eastern United States. Interestingly, the dominance pattern reflects that of vents with annelids constituting nearly 50% of the sediment collections. The similarity at the subclass/ordinal levels is less evident. For example, in sediments, the peracarid crustaceans dominated the Grassle and Maciolek samples whereas the decapod and copepod groups are most abundant at vents. The differences between bare ridge and sediments surely relate mostly to feeding adaptations. Thus one might make the same assumption to explain vent composition, although Tunnicliffe (1992) also implicates the protective nature of shells, carapaces and thick cuticles in the vent habitat in relation to the numerous chemical toxins.

3. OTHER RELATED FAUNAS Biologists have long known of associations between animals and high sulphide environments (Fenchel and Finlay, 1995). Sulphate reduction proceeds as free oxygen is rapidly consumed by microbes in organic-rich environments such as salt marshes, sewage outfalls, and other near-shore environments with high input of terrestrial matter. It is not unlikely that adaptation to high sulphide and even symbionts occurred early in metazoan evolution when oxygen concentrations were much lower (Reid and Brand, 1986; Fenchel and Finlay, 1995). Today, symbioses with sulphide-oxidizing

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Table 3 Vagrant taxa observed near hydrothermal vents. Most of these octopods and fish are most likely to be “visitors” at vents although several, such as Thermarces, are obviously abundant inhabitants of vents. Numbers and letters in the last column are the same annotations as Table 2.

Mollusca Cephalopoda Octopoda Cirroteuthidae Octopodidae Octopodidae Cbordata Chondrichthyes Chimaeriformes Chimaeridae Chimaeridae Scyliorhinidae Squaliformes Squalidae Squalidae Squalidae Squalidae Osteichthyes Anguilliformes Synaphobranchidae Synaphobranchidae Synaphobranchidae Synaphobrachidae Beryciformes Berycidae Gadiformes Gadidae Macrouridae Macrouridae Moridae Moridae Lophiiformes Chaunacidae Chaunacidae Notacanthiformes Notacanthidae Ophidiiformes Bythitidae

Cirroteuthis magna Hoyle Graneledone spp. Benthoctopus spp.

Hydrolagus mirabilis Collett Hydrolagus pallidus Hardy and Stehmann Apristurus maderensis Cadenat and Maul Centroscymnus coelolepis Bocage and Capello Etmopterus mirabilis (Collett) Etmopterus princeps Collett Etmopterus pusillus (Lowe) Zlyophis blachei Saldanha and Merrett Simenchelys parasitica Goode and Bean Synaphobranchus kaupi Johnson Thermobiotes mytilogeiton Geistodoerfer Beryx splendens Lowe Gaidropsarus n. sp. Saldanha and Biscoito Coelorhinchus cf. Iabiatus (Koehler) Nematonurus armatus Hector Lepidion ? schmidti Svetovidov Mora moro (Risso) Bathychaunax roseus (Barbour) Chaunax sp. Biscoito & Saldanha Polyacanthonotus cf. rissoanus (Filippi and Virany) Cataetyx laticeps Koefoed

9,N

38 1

DEEP-SEA HYDROTHERMAL VENT FAUNA

Perciformes Apogonidae Bythitidae Cycopteridae Zoarcidae Zoarcidae Zoarcidae Scorpaeniformes Scorpaenidae Zeiformes Oreosomatidae

Epigonus telescopus (Risso) Bythites hollisi Cohen, Rosenblatt and Haedrich Careproctus hyaleius Geistdoerfer Pachycara gynmium Anderson and Peden Pachycaru thermophilum Geistdoerfer Thermarces cerberus Rosenblatt and Cohen

9 3 3,4,5,Y

4,Y 1 ,N 9,Y 3,4,Y

Trachyscorpia cristulata (Koehler)

9,N

Neocyttus helgae (Holt and Byrne)

9,N

and/or methanotrophic microbes are recorded in many taxa such as bivalves (Felbeck, 1983; Southward, 1986), pogonophorans (Cavanaugh et al., 1981; Southward and Southward, 1988), nematodes (Giere et al., 1984) and sponges (Vacelet et al., 1995). Here, we address only those habitats of the deeper ocean that appear to have faunal interactions with the hydrothermal vent habitat.

Figure 4 Overall composition of the total invertebrate fauna (as listed in Table 2) recorded from vents.

V. TUNNICLIFFE ET AL.

Figure 5 Frequency (%) of major invertebrate phyla at vents compared to adjacent bare ridgecrest and deep-sea sediments. (Ridge data are compiled from Carey et al., 1990, and Copley et al., 1996, and sediment data are drawn from Grassle and Maciolek, 1992.)

3.1. Seeps

Extensive communities with obligate dependence on microbial oxidation processes thrive in the deep sea; here, the geochemical sources for energetic reactions can swamp the scanty input from overlying photosynthetic sources. At “cold seeps” remineralization of organic carbon at subduction zones or in hydrocarbon effluents produces extensive sulphide and methane compounds. Locations of known seeps below 500m depth are indicated in Figure 6. Most are located along subduction faults of the Pacific Ocean where they are known to 6000 m (the depth limit of research submersibles). A comprehensive review of 24 cold seep sites and the component organisms is presented by Sibuet and Olu (1998). Seeps can occur on both active subduction margins and on passive continental margins (both the west and east sides of North America, for example). The causes and nature of these seeping fluids are quite varied. On many active margins, the subducting plate scrapes much of its sediment burden against the overlying continental plate to form a wedge of sediments over the subduction zone known as an accretionary prism. These sediments, with their accumulated organic carbon, are compressed as the plate moves.

s

-~

E

~~

120

150

~~

180

150

120

~

9Ov

60

30

0

30

60

90%

Figure 6 Locations of margin cold seeps below 500 m (filled circles) and whale carcasses. Only the better-known seeps are shown: Jps = shallow Japan; Jpd = Japan Trench; Ale = Aleutian Trench; Ore = Oregon margin; Mon = Monterey Canyon; Per = Peru Margin; Lsp = Louisiana slope; Fla = Florida margin; Bar = Barbados Prism; Lfn = Laurentian fan. (More information about these and other sites can be found in Sibuet and O h , 1998.)

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Water within the sediment is forced out along weak bedding planes or faults. The carbon sources are organic: such as methane, petroleum, other hydrocarbon gases or from solid gas hydrates (Simoneit et al., 1990; Aharon, 1994; MacDonald et al., 1994). The extent of seepage communities in the deep ocean is unknown as margin environments are incompletely explored. Areal extent can be large (Olu et al., 1996). The seeps of the Gulf of Mexico have been active since the last glaciation period and the area has probably been seeping for at least 200000 years (Aharon et al., 1997). Sibuet and Olu (1998) invoke this long-term stability as a significant factor in the evolution of a diverse seep fauna.

3.2. Organic Remains Elsewhere in the deep sea, organic-rich areas appear to deplete local oxygen supplies and attract a resident fauna. One intriguing record of an organic sulphide association is that of pogonophorans in the cargo hold of a shipwreck off Portugal (Dando et al., 1992). More common are the animals collected on whale bones and sunken wood. The bones contain enough oils to maintain a sulphide microenvironment for many years (Smith et al., 1989). Deming et al. (1997) illustrate the chemosynthetic basis of the invertebrate community around whale remains. Numerous novel taxa have been collected from such finds, most of which are molluscan (Marshall, 1987; 1994; Fujioka et al., 1993; Bennett et al., 1994). These discoveries have prompted interesting speculation on the long-term evolution of “osteophiles” given the abundance of large marine reptiles and mammals in Mesozoic and Cenozoic seas (Squires et al., 1991; Hogler, 1994). Sunken wood in the deep sea is frequently a source of new taxa of molluscs (Turner, 1972; Marshall, 1988; McLean, 1991; WarCn and Bouchet, 1993). Shipworms (i.e. teredinid clams) dominate sunken wood in shallow waters but are replaced by the wood-boring bivalve Xylophaga in the deep sea (Turner, 1973). Both harbour bacterial symbionts related to wood digestion and share similar gill morphologies with symbiont-containing bivalves from other sulphide-rich environments, including vents: Lucinidae, Vesicomyidae, Modiolinae, and Solemyidae (Distel and Roberts, 1997). The mytilid genera Adipicola and Zdasola are known from both vents and sunken wood (Dell, 1987). The gastropods known from sunken wood are from diverse groups with genera shared with hydrothermal vents, hydrocarbon seeps and whale bone (Marshall, 1988; McLean, 1991; WarCn et al., 1993; WarCn and Bouchet, 1993). The ostracod genus Xylocythere is known from vents and sunken wood (Maddocks and

DEEP-SEA HYDROTHERMAL VENT FAUNA

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Steineck, 1987; van Harten, 1992). Many wood associated invertebrates thus have tolerance of a wide range of sulphide-rich habitats.

3.3. Other Sites

Shallow-water volcanic emissions are another source of reducing compounds. Microbial mats and sometimes dense suspension-feeding communities are reported at these sites (Karl et al., 1988; Fricke et al., 1989; Zhirmunsky and Tarasov, 1990). With little taxonomic systematic information available, it is still unclear how to treat these systems. A specifically adapted fauna does not appear to have developed, instead microbial production enhances the surrounding normal fauna. However, it should be noted that taxa with vent or seep affiliation may emerge at such sites (Williams and Dobbs, 1995).

4. THE BIOGEOGRAPHY OF FAUNAS

4.1 Distribution of Vents

Vents were first discovered on Galapagos Rift in 1977 (Corliss et al., 1979). With the discovery of vents at 21"N and 13"N on the East Pacific Rise (Spiess et al., 1980; Hekinian et al., 1983), the first speculations about faunal similarities and dispersal began. Desbruykres et al. (1982) note the similarities between 21"N and 13"N, including the absence of the clam Calyptogenu magniJica. They also drew a few comparisons with the Galapagos Rift site. It was soon clear that many elements of the fauna were similar - the tubeworms, clams and mussels, for instance - but the species richness was higher on EPR. Was this a sampling artefact, an historical/evolutionary phenomenon or a function of habitat? Although years of explorations of these sites have located many "missing" animals, first impressions of differences in relative abundances still remain valid. Galapagos does not have the high temperature smoker chimneys found on EPR - a habitat for a specific subcommunity such as Alvinellu, Eunice (polychaetes) and Cyanograea (crab) (Desbruykres et al., 1982). With the examination of vents in Guaymas Basin in 1982 (Lonsdale and Becker, 1985), similarities to other EPR sites were evident, but it was clear that the habitat factor influenced the assemblage present. Guaymas vents exit through thick sediment cover and many associated organisms are adapted to sediment rather than bare rock (Grassle, J.F., 1985).

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The 1980s saw many vent discoveries elsewhere in the world. Comparison with EPR and Galapagos was an initial goal for some biologists studying these sites, with questions of dispersal, colonization and divergence not far behind. The most striking differences arose when vents in the Atlantic Ocean were located (Rona et al., 1986). While similarities in some faunal groups became apparent, both the dominant invertebrates and the basic community structuring are substantially different from Pacific vents (Van Dover, 1995; Gebruk et al., 1997). Current work in the Indian Ocean represents a new frontier of exploration that begins to explore faunal pathways to the Atlantic (Southward et al., 1997). An examination of the global distribution of known vents (Figure 2) emphasizes many current questions concerning the effects of geography on vent characters. The highly linear disposition of vents along the midocean ridge is striking. The connectedness among southern oceans is a marked contrast to “terminations” in the northern oceans. Immediately, the potential for regionalization of faunas becomes evident. Modern-day barriers are evident: continental masses, extensive ocean plates and shallow seas. Pathways for dispersion are fairly obvious and the actual surface area suitable for colonization appears minute. This latter feature is perhaps important when considering the relatively low complexity and diversity of this ecosystem.

4.2. Regional Studies on the Relations of Vent Faunas

Once two hydrothermal vent sites were known, biogeographic comparisons began. The RISE Project Group (1980) noted great similarities between the fauna of 21”N and Galapagos but also commented: The cause of faunal differences is completely unknown, and biogeographic or ecological mechanisms are equally possible. Unfortunately, the Galapagos and East Pacific Rise are in the same ocean sector; therefore we must wait for more distant discoveries to demonstrate whether the community is widespread. (p. 1425).

The subsequent discovery of communities at 13”Nrevealed a great similarity to 21”N and Desbruybres et al. (1982) suggest the variable species compositions of these and the Guaymas site perhaps reflect chance species dispersal. However, discoveries in other “ocean sectors” soon appeared.

4.2.1. Northeast PaciJic The story of faunal relations became more complex with the discovery of vent fauna on the Juan de Fuca Ridge in 1983. While many animals were

DEEP-SEA HYDROTHERMAL VENT FAUNA

387

clearly related to those of the EPR and Galapagos, they were mostly different species and there were some different genera (Canadian American Seamount Expedition, 1985; Tunnicliffe et al., 1985). Juan de Fuca Ridge holds a special place in the history of modem geology. It was here that the timescale of ocean-floor magnetic reversals was conceived by Cox et al. (1964). A key paper published by Vine and Wilson (1965) both names Juan de Fuca Ridge and provides the fundamental model for modern plate tectonic theory. With work over the next decades by Atwater (1970), Rea and Dixon (1983), Riddihough (1984) and Engebretson et al. (1985), a history of the relations between Juan de Fuca and East Pacific Rise became clear. In the Mesozoic and early Cenozoic, one north-south ridge existed in the eastern Pacific (Figure 7). This ridge separated an eastern Farallon Plate from the Pacific Plate. However, the North American Plate has been drifting southwestward and overran the ancestral ridge about 30 million years ago. Today, Juan de Fuca Ridge remains as the northern fragment of this old ridge configuration. Presumably, the future will see the disappearance of the northeast ridges and the sundering of Baja California from North America. Tunnicliffe (1988) proposes that this vicariant event was responsible for the isolation of a northeastern Pacific fauna that still bears many similarities to that of the East Pacific Rise. From the Miocene onward, the dispersal barrier between EPR and the northern ridges grew larger and isolation increased. A “subsampling” of the original fauna may explain the absence of many elements of the EPR fauna that are missing in the northeast Pacific (e.g. endemic decapods). With a further 10 years of collections in both regions, the picture has not changed much. Eleven of the 80 identified species of the northeast Pacific (Explorer, Juan de Fuca and Gorda Ridges) are known from the East Pacific Rise or Galapagos sites. Long-range dispersal or slow speciation may explain these shared species. Another 27 species are in genera (mostly endemic) known from one of the EPR or Galapagos vent sites; this similarity suggests a longer evolutionary connection. 4.2.2. Western Pacific The first record of hydrothermalism in the western Pacific came in 1986 with bottom photographs of chimneys and large snails in Manus Basin (Both et al., 1986). An abundance of gastropods, subsequently determined to have microbial symbionts (Stein et al., 1988), was characteristic of subsequent observations of western Pacific vents including the Mariana and Fiji backarc basins (Waren and Bouchet, 1993). Hessler and Lonsdale (199 1) examine the fauna of the Mariana Trough in the context of biogeographic relations with other vent regions. Because over

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Present

56 Ma

65 Ma

110Ma

Pacific Figure 7 Mid-ocean ridge model for the northeast Pacific based on palaeomagnetic data. (From Tunnicliffe et al., 1996.)

half the species they found belonged to genera or families endemic to eastern Pacific ridges, they conclude that interchange with mid-ocean ridges has been an important part of the evolution of this fauna. Modern pathways connect through several basins to the north of New Zealand; a large barrier separates this back-arc complex from the southeast Indian Ridge (Figure 2). The western Pacific has a complex spreading history that, at the very least, bespeaks a region of continuous Cenozoic ridge activity despite changing

DEEP-SEA HYDROTHERMAL VENT FAUNA

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configurations. Hessler and Lonsdale (1991) and Tunnicliffe et al. (1996) review the geophysical models for the Cretaceous to early Tertiary. The salient points include a vigorous Cenozoic history of magmatic activity that would provide many suitable habitats, shifting configurations that would isolate or join adjacent faunas, termination of ridge connections to the Indian Ocean north of New Guinea in the early Tertiary, and a Mesozoic to mid-Cenozoic ridge connection into the western Pacific from the southern ridge between the Pacific and Antarctic Plates. Figure 8 illustrates the presence of an east-west Pacific ridge that both Hessler and Lonsdale (1991) and Tunnicliffe (1997) invoke as a possible interchange mechanism between the two sides of the ocean. The direction of faunal movement at such a time is not clear (Hessler and Lonsdale, 1991). Several hydrothermal sites are known around the northern periphery of the Philippine Sea Plate. Hashimoto et al. (1995) point out the regional similarities among these sites and with the cold seeps along the Japan Trench. These authors invoke deep-water circulation as an important mechanism for larval exchanges that bridge the large gaps among western Pacific hydrothermal sites. While there are some similarities at the generic level between the northwestern (MBJ) Pacific and northeastern Pacific (NEP), most of these genera also occur in the EPR region. Tunnicliffe

Figure 8 Possible Eocene northern Pacific plate boundaries illustrating a multiple plate system and an active mid-ocean ridge in the western Pacific. (From Tunnicliffe et al., 1996.)

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(1997) suggests that the MBJ region may become more distinct from the Australasian region (LFM) as more taxa are described. Only nine species are currently shared between these areas of the western Pacific (and the generic level similarities are higher than between the two sides of the northern Pacific). The communities of the western Pacific are notably heterogeneous in composition and assembly. The dominant biomass may be formed by a variety of taxa: mussels, clams, provannid snails, or barnacles. Vestimentiferans and anemones may be abundant also. This diversity may relate to habitat or differential niche exploitation (Desbruy6res et af., 1994). In their studies of the Manus, Fiji and Lau fields, Galkin (1992) and Desbruyires et af. (1994) note the strong similarities among these southern sites; they emphasize the need for further work - such as temporal studies in this region. We iterate this need: it is interesting to note the rather large number of species retrieved so far with relatively little collection effort.

4.2.3. Southern Pacijic

In the last few years, a great interest in "super-fast" spreading centres has arisen among many disciplines of vent work. Fornari and Embley (1995) summarize the early focus of geological workers on the southern East Pacific Rise between 14"s and 21"s. Several submersible dives by the French in 1984 identified areas of highly diffuse flow and numerous indications of recent lava eruptions (Renard et al., 1985). Many lines of evidence point to this region as a highly active centre with high magmatic activity that could sponsor extensive hydrothermalism. For example, Baker et af. (1995) describe "significant hydrothermal plumes" covering over 60% of the ridge crest from 13.8"s to 18.8"s (about 550 km distance). As yet, published biological descriptions of this region are limited. The samples from the late 1993 French dives are still under description. Juniper et af. (1990) examine the 1984 images noting species similar to northern EPR but in much lower concentrations. The 1993 observers examined 69 areas of hydrothermalism and described four distinct hydrothermal stages based mostly on the types of animal communities seen (Fouquet et af., 1994; Auzende et al., 1996). A very recent eruption had occurred in one area where only bacterial mats were seen. Geistdoerfer et al. (1995) expand the faunal descriptions and cite the collection of 30 species. They note the great similarity of the assemblage to that from north East Pacific Rise (although a few novel species are appearing in the literature: see Table 2). However, they make an important observation that the community assembly appeared different in several areas, such as the vast extent of fields of anemones and bivalves. Another dive series by a Japanese submersible examined the

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same site 10 months later to find the communities notably different (Embley et al., 1998); colonization had occurred where Auzende et al. (1996) reported only mats and shimmering water. Rapid community evolution in regions of unstable venting may influence biogeographic interpretations. The entire suite of regional species may not be sampled if all community stages are not visited. The question of habitat suitability rather than dispersal barriers may also be important: the unpredictable nature of a fast-spreading ridge axis may hinder some species from becoming established. 4.2.4. Atlantic Two publications address the Atlantic fauna and biogeographic relations in considerable detail: Van Dover (1995) and Gebruk et al. (1997). We summarize their works briefly here. The discovery of venting in the Atlantic expanded the models of hydrothermalism from medium and fast spreading ridges to include ridges with slow spreading rates. With those initial discoveries (Rona et al., 1986; Galkin and Moskalev, 1990), a new biogeographic realm emerged. The succeeding decade saw the location of five more vent sites that spanned 23" of latitude (Table 1). The depth range - from less than 820 m to over 3600 m - is intriguing for the potential in biological zonation. Another feature that may have significance for the evolution of these communities is the apparent long-term stability of the vent fields. Lalou er al. (1995) identify numerous cycles of venting using radiometric ageing of TAG sulphides; sporadic activity spans a period of nearly 150 000 years. Community organization is quite different from vents in the Pacific. Bresiliid shrimps and bathymodiolid mussels dominate the biomass while pogonophoran and alvinellid worms are absent. Gebruk et al. (1997) illustrate several sites where anemones and echinoderms are abundant in the peripheral fauna. The shrimp are the focus of many ongoing studies because of their abundance, their diversification and their unusual physiological features (Van Dover, 1995). These animals may be particularly amenable to studies that examine dispersal and communication among vents (Herring, 1997). Much of the Atlantic vent fauna is likely derived from the Pacific. There are many links to the higher diversity Pacific fauna and the Atlantic is also a younger ocean. The Atlantic opened in the Cretaceous and the full northsouth ridge from the southern ocean through to the Labrador Sea was established by the early Cenozoic. Tunnicliffe el al. (1996) examine several routes to the Atlantic including Caribbean, Cape Horn and Mediterranean. They conclude that the most likely route remains through the Indian Ocean but do not dismiss the possibility that a Tethyan equatorial dispersal route

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may have served some taxa as emphasized by Gebruk et al. (1997). Today, the St. Paul and Romanche transform faults may serve as major barriers between north and south Atlantic faunas if any of the latter exist (Van Dover, 1995). The similarity of several elements of this fauna to that of seeps suggests that there may be an intimate evolutionary relationship (Gebruk et al. 1997). With the cruises of the last few years, material is now available to explore systematic relations. 4.2.5. Indian Ocean

There have been no biological investigations to date on the ridges of the Indian Ocean. Evidence of hydrothermal venting is reported, however, from the triple junction region in the middle of the Ocean: vent plumes and images of inactive sulphide deposits (Pluger et al., 1990; Gamo et al., 1996). Dredged sulphides apparently contain structures reminiscent of entombed tubes (U. Munch, pers. comm.). Scheirer et al. (1998) report hydrothermal plumes on the Southeast Indian Ridge. Dredges returned lepadomorph barnacles from this site (Figure 2) closely allied to those known from the western Pacific in the Lau/Fiji region (Southward et al., 1997). Van Dover (1995) and Tunnicliffe and Fowler (1996) suggest that the Indian Ocean is a key pathway to the southern Atlantic. Plumes have been located recently on the Southwest Indian Ridge (C. German, pers. comm.); thus an active pathway of distribution may exist through the Indian basin. The northern extension of the Central Indian Ridge enters the Gulf of Aden. Aggregations of shrimp have been reported from this Aden extension (Juniper et al., 1990), but at an order of magnitude less than along the Mid-Atlantic Ridge (Gebruk et al. 1997).

4.3. SeepIWhalelVent Relations

Is the hydrothermal vent fauna similar to those of other sulphide-rich environments? Yes and no: it rather depends upon the point you are trying to make. The first discoveries of seeps and whale skeletons emphasized the similarities to vents because of the unusual mode of nutrition (Southward, 1989) and the potential that many taxa would be found to bridge the habitats (Hecker, 1985; Kulm et al., 1986; Laubier et al., 1986; Smith et al., 1989). There are many physiological similarities among the animals (Fiala-Medioni and Le Pennec, 1987; Fisher, 1990; Scott and Fisher, 1995), but in the context of biogeographic work at the species level, there is relatively little overlap among the faunas (Tunnicliffe et al., 1996; Sibuet and Olu, 1998).

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Figure 9 illustrates that few of over 440 species recorded from hydrothermal vents are yet known at seeps or on skeletons. Ten seep animals overlap with vents and five whale bone animals overlap. In addition, recent reports indicate that at least two species are found in all three habitats: the vestimentiferan Escarpia spicata (Feldman et al., 1998) and the gastropod Pyropelta corymba (A. WarCn, pers. comm.). There is little doubt that the overlap numbers will grow with increased exploration; current work suggests the number of shared species with whale bones may be higher (Smith et al., 1998). Overall, however, these non-vent habitats are not important as dispersal stepping stones for the current vent fauna overall nor should such sites confuse a biogeographic analysis at the species level. Despite the dominance of symbiont-containing hosts (pogonophorans and molluscs) at both

Figure 9 Overlap of species among sulphide-rich habitats and the deep sea. Vent and seep habitats are represented by triangles overlapping with each other and with the deep sea. Species displayed within the triangle represent the number of endemics, while the number of overlap species are also indicated (? means species not yet identified). The whale represents the species found on decaying bones on the ocean floor. Of the total whale bone species shown, two species are known from whale bones, seeps and vents; one is found at bone, vent and deep sea; four are known from vents and bone, while one other is known only from bone and seeps; the remainder are endemic to bones.

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vents and seeps, there are only a few species that transcend the habitat boundaries (Sibuet and O h , 1998). There are, however, several interesting links at the individual taxon level that should make biogeographic analyses of individual clades interesting. Note, for example, the occurrence of the polynoid Branchipolynoe seepensis at both Florida seeps and Atlantic vents where it occurs as a commensal of mussels. Chevaldonnt et al. (1997a) examine phylogenetic relations of this genus and suggest that genetic differentiation is sufficient to represent different species. About half the species in Table 2 that cross into seep or carcass habitats are known from sedimented vent sites; sites such as Middle Valley (Juan de Fuca Ridge), Escanaba Trough (Gorda Ridge) and Guaymas Basin may emulate seep conditions better than the bare rock ridge crest sites. Southward et al. (1996) note the habitat separation between the vestimentiferans Lamellibrachia barhami (known from seeps) and the vent endemic Ridgeia piscesae. There are enough similarities at higher taxonomic levels to suggest strong evolutionary links. McArthur and Tunnicliffe (1998) discuss the mosaic origins of the vent fauna including relations to the deep sea and other sulphide-rich habitats. About 50% of vent genera are found in sulphiderich habitats that require physiological adaptations to the reducing conditions. Chevaldonnt and Olu (1996) examine the anomuran crabs at vents and seeps; noting that the general biological traits of these animals adapt them well to at least the periphery of sulphide-rich habitats, if not the harsher “interior”. The rate and mechanism of sulphide delivery is different between vent and seep/whale (Scott and Fisher, 1995); perhaps this difference is partly responsible for the difference between vent and seep/whale organisms. Seeps and carcasses may provide an entry point into the vent habitat. Molecular studies detail the phylogenetic relations between vent and seep (plus whale carcass) taxa: vestimentiferans (Williams et al., 1993; Black et al., 1997; Feldman et al., 1998), vesicomyids (Vrijenhoek et al., 1994; Peek et al., 1997) and mytilids (Craddock et al., 1995a).

5. LOCAL TO REGIONAL-SCALE PROCESSES

Many biogeographers recognize distinct ecological and historical approaches, each dealing with different time and spatial scales. While there is much overlap, it is convenient to examine processes on the intra- versus interregional scales. An ecological approach includes consideration of mechanisms of communication and dispersal among communities. An historical approach to regional faunas encompasses the events that cause barrier formation and breakdown. There is a need for both approaches in vent

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biogeography as we deal with organisms with potentially dispersive stages that live in an ecosystem associated with the primary driving force of shifting barriers on our planet: sea-floor spreading. At this point in the study of biogeography, it is a challenge to tease out the role of these two approaches. 5.1. Reproductive Strategies

Relatively few accounts of reproduction and development of vent fauna have been published; time-series samples from vent populations are difficult to obtain. Culture of seep species in the lab has only recently been achieved (Young et al., 1996). Nonetheless, single samples of animals from different areas have yielded a great deal of information (e.g., Lutz et al., 1980; Le Pennec et al., 1984; Berg, 1985; Turner et al., 1985; Van Dover et al., 1985; Berg and Van Dover, 1987; Cary et al., 1989; McHugh, 1989, 1995; Gustafson et al., 1991; Gardiner et al., 1992; McHugh and Tunnicliffe, 1994; Zal et al., 1994, 1995). These studies show that vent species exhibit a wide range of reproductive and developmental modes reflecting phylogenetic origins of the species rather than the opportunistic life history strategies initially predicted for such an ephemeral and patchy environment (DesbruyZres and Laubier, 1984; Grassle, J.P., 1985). While planktotrophic larval development has been inferred in some cases, lifecycles of many vent species are thought to exhibit development modes expected to restrict dispersal, i.e. lecithotrophy or brooding of larvae (see Tunnicliffe, 1991; Jollivet, 1996). Readers are referred to a forthcoming review of reproduction and development of vent fauna (P. Tyler, pers. comm.) for a full account of life history biology of vent organisms. 5.2. Dispersal

There is no doubt that the factors governing dispersal of vent animals are important to vent animal distributions. Short-range dispersal certainly must affect within-region distributions and gene flow. Long-range and “sweepstakes” dissemination may explain some anomalously large species ranges. We will not review vent animal dispersal here. A recent summary (Mullineaux and France, 1995) encapsulates most of the dispersal phenomena. They find no universal dispersal mechanism. Analysis of dispersion capability of these animals requires understanding the life span of the larvae and the regional water transport. While we know little about either yet, the potential for extensive local movement exists. Rapid colonization of new vents from distant populations confirms the dispersal potential of some vent species (Tunnicliffe et al., 1997). It is not unreasonable to expect

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a variety of different strategies among vent animals: some will be those adapted to rapid population shifts and exploitative colonization of new vents and others that conserve populations in stable areas of venting (Vrijenhoek, 1997). The proportion of such species in a region may depend upon regional ridge dynamics (Juniper and Tunnicliffe, 1997). Larvae cannot swim the great distances between vent sites so must rely on transport by ocean currents. Currents near the ridge show daily tidal reversals but a progressive drift parallels the ridge crest; mean velocities range between 1 and 5 cm s-' (Cannon and Pashinski, 1990; Thomson et al., 1990). Peak velocities of tidal flows could move larvae over a kilometre in a few hours (Mullineaux and France, 1995). Instead of this bottom-flow transport, however, larvae may be carried in the higher hydrothermal plume (Kim et al., 1994). Vent larvae may have a higher concentration in these plumes relative to normal deep-sea larvae (Mullineaux et al., 1995). While these plumes rise about 300 m above the sea floor, the massive expulsion of hydrothermal fluids in megaplumes that rise over 1OOOm (Baker, 1994) may provide a rare but effective mechanism of long-distance transport. 5.3. Gene Flow

The ephemeral existence of individual vents and patchy distribution of venting and vent fields on a variety of scales generate many questions about population interactions. Examples include: local survival of individual species, maintenance of genetic continuity among vent fields on a ridge and the likelihood of genetic discontinuities and allopatric speciation among the major ridge systems. High species provinciality of the vent fauna argues for genetic discontinuity and cladogenesis among the major ridge systems; the predominance of monospecific families and the broad distribution of species within individual ridges suggests genetic panmixis within individual ridges. Oceanographic barriers and large vent-barren distances could possibly be the only limit to successful dispersal of vent larvae. Alternatively, morphologically-based species definitions may poorly reflect actual species boundaries and ranges (e.g. Craddock et al., 1995a; Vrijenhoek et al., 1994). Initial expectations for dispersal and genetic exchange between hydrothermal vent populations were based on investigations of the varied larval stages of hydrothermal vent animals. The most consistent contrast was between those animals with direct development or short-lived, non-feeding lecithotrophic larvae and those with feeding planktotrophic larvae that may spend weeks or months in the water column. Lutz et al. (1984) predict that the former larvae would disperse in a vent-to-vent stepwise process, while the latter would exhibit panmixis over large distances. Inferences about larval ecology of vent molluscs were based on expansions of Thorson's

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(1950) shell apex theory, in which the morphology of larval shells preserved in the shells of adults records the mode of larval development - lecithotrophy or planktotrophy (e.g. Lutz et al., 1980). Several factors make molecular genetics the ideal tool for examining population structure and species boundaries for vent species: concerns about the applicability of Thorson’s rule to deep-sea molluscs, the large gaps in knowledge about reproductive and larval biology of most other vent species, and the unknown influences of metapopulation dynamics and extinction/colonization cycles upon gene flow and population subdivision. Not surprisingly, much of the published work on population genetics was conducted on widespread species; one may not expect, and obviously not test for, extensive gene flow in species that do not have wide distributions. Vrijenhoek (1997) synthesizes the population genetic studies of vent animals - we summarize his results here. Molecular investigations generally illustrate extensive communication among vents within a ridge system. It is now clear that larval shell morphology of vent molluscs is a poor indicator of dispersal patterns - presumed lecithotrophs such as Calyptogena magnifica, Eulepetopsis vitrea and Lepetodrilus pustulosus exhibit extensive gene flow and no isolation of sampled populations by distance (Karl et al., 1996; Craddock et al., 1997). Overall, genetic investigations of bivalves (Vrijenhoek et al., 1994; Craddock et al., 1995b; Karl et al., 1996) and archaeogastropods (Craddock et al., 1997) at the East Pacific Rise, Galapagos Rift, and Guaymas Basin support an island model of larval dispersal within and among ridge systems. In such a scenario, larvae from different locations form a single well-mixed migrant pool that feeds larvae to all three ridge systems (Vrijenhoek, 1997). Similarly, populations of Bathymodiolus at both Lau and North-Fiji vents have high gene flow between them (Moraga et al., 1994). In contrast, populations of the gastropod Alvinoconcha spp. do not exhibit genetic continuity over the same range in the western Pacific, although genetic distances may be representative of species boundaries instead of isolation of populations by distance (Denis et al., 1993). Gaps between segments within ridge systems appear to restrict alvinellid polychaete populations to stepping-stone dispersal among segments and genetic divergence among ridge systems is high, suggestive of isolation by distance (Jollivet et al., 1995a; Vrijenhoek, 1997). Molecular investigations of vestimentiferans (Black et al., 1994; Southward et al., 1996) reveal some isolation by distance and thus vestimentiferans are also inferred to utilize stepping-stone dispersal (Vrijenhoek, 1997). However, estimated rates of gene flow were sufficient to maintain genetic continuity among populations of Rijtia pachyptila on the East Pacific Rise, Galapagos Rift, and Guaymas Basin (Black et al., 1994) and

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along a 450km stretch of the Juan de Fuca Ridge for Ridgeia piscesae (Southward et al., 1996). Among the Crustacea, Creasey et al. (1996) find high gene flow and no genetic subdivision among populations of the shrimp Rimicaris exoculata over a 370 km distance on the Mid-Atlantic Ridge. The direct developing amphipod Ventiella sulfuris exhibited extensive gene flow along the East Pacific Rise, and isolation by distance was observed only across the Rivera Fracture Zone (France et al., 1992). However, genetic divergence exceeded that predicted through isolation by distance when France et al. (1992) compared populations of the East Pacific Rise and Galapagos Rift - topographic or oceanographic features may present a strong barrier to dispersal for vent species without a free-living larval stage (Vrijenhoek, 1997). Other genetic signatures from vent species are revealing. Vrijenhoek (1997) finds rare alleles (i.e. alleles with low frequencies in the population) uncommon in vent species and suggested that frequent genetic bottlenecks and founder events resulting from vent extinction/colonization cycles result in their removal. Comparison of levels of polymorphism and occupancy, defined as “the proportion of sampled sites at which a particular species was found” (Vrijenhoek, 1997, p. 290), reveal that ridge spreading rate and faunal succession following initial colonization of nascent sites may control distribution of species both within and among ridge systems. Maintenance of genetic diversity or even viable populations may be dependent upon a species’ relationship to successional processes - late successional stages required by some species may never exist at rapidly spreading ridges where individual vents have short life spans (Vrijenhoek, 1997).

6. REGIONAL TO GLOBAL-SCALE PROCESSES

Examination of the inter-regional patterns of any biota requires a shift in approach. As gene flow diminishes over large distances, the role of propagule dispersal becomes less pronounced. Just as any region holds several kinds of communities, we also see differences in character among regions influencing the type and diversity of species present. Historical events that have shaped the manner of species accumulation are usually evoked on this scale. 6.1. Ridge Characteristics

Variations in the sea-floor spreading process, within and between regions, affect the nature of the hydrothermal habitat. Location and behaviour of

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the underlying magma chamber are primary controls on heat source while faulting properties will determine the extent and depth of hydrothermal circulation. Fornari and Embley (1995) present a comprehensive review of these controls and their variability around the world. The major manifestation of regional differences is variation in sea-floor spreading rates. Rates range from the slow-spreading Mid-Atlantic Ridge at 2 4 cm yr-I, to intermediate (Galapagos, Juan de Fuca), to fast (parts of northern EPR) to superfast on southern EPR where rates exceed 15cm yr-I. Spreading rate appears to influence the abundance of hydrothermal sites and the overall hydrothermal flux through the crust (Baker et al., 1995). The southern EPR, for instance, sponsors extensive hydrothermal plumes over hundreds of kilometres of ridge crest (Urabe et al., 1995). Segmentation of the ridge crest is also a fundamental feature that relates to the behaviours of individual magma chambers (Macdonald et al., 1991; Haymon, 1996). There is extensive variability in the extent of segmentation and the ramifications for hydrothermalism. Venting tends to occur on the shoalest points of segments but there may be many intervening segments with no venting (Fornari and Embley, 1995). Fragmentation is greatest in the western Pacific where ridges are very short. One may expect greater provincialism here. The Mid-Atlantic Ridge shows much greater depth variability both along segments and between segments compared to the relatively smooth southern EPR. Ridge-crest behaviour and placement also influence the heterogeneity of vent habitats. There is much segment-to-segment variability in depth, eruption frequency, cracking behaviour and substratum. In the northeast Pacific, for instance, hydrothermalism is known in a wide range of geologic settings: sedimented valleys, deeply faulted areas and neovolcanic zones (Fornari and Embley, 1995). Axial Volcano sits astride the Cobb hotspot that provides another source of magmatic influence (Hammond, 1990). The extent of venting among segments is highly variable in time and space as indicated by the hydrothermal plumes over the ridge (Baker and Hammond, 1992). 6.2. Vicariance

Mantle convection drives the movements of the lithospheric plates. Over the millennia, continental positions change, oceans are reshaped or even obliterated, and new ridges form. Vent faunal evolution must be intimately connected with this process; an examination of ridge and plate history in the context of biogeographic relations reflects an appreciation of the role of vicariance events in structuring modern distribution patterns. However, a major challenge for geophysicists is the reconstruction of past plate

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histories. That reconstruction is based mainly on interpretation of paleomagnetic signatures in the oceanic plates. As the magnetic poles reverse, the polarization of cooling lavas at the spreading ridges switches. That new lava, on both sides of the ridge, is slowly pushed laterally as newer lavas emerge and neighbouring tectonic plates pull apart. Today we see matching stripes of polarization on each side of the ridge crest. Older ridges can thus be reconstructed by “removing” the younger crust and putting matching stripes back together. The oldest ocean crust is of Jurassic age in the western Pacific - subduction has destroyed older material. Reconstructions become more contentious when other evidence must be used. Weijermars (1989), Tunnicliffe (1991), Van Dover (1995), Tunnicliffe et al. (1996) provide summaries of the Mesozoic and Cenozoic global histories of ridges and plates. Several major shifts in relative ridge positions influence interpretation of biogeographic relations: severing of the East Pacific Rise connection with the Juan de Fuca Ridge region, isolation of the Atlantic from the eastern Pacific (in Caribbean and Cape Horn regions) and continuity of connection from the eastern Pacific to the western Pacific and Indian oceans. The western Pacific has a complex history but fragmented ridges are recorded in the area throughout the Cenozoic. Speculations on how each regional fauna is affected are presented in Section 5. A global analysis of similarities among the faunas listed in Table 2 can illustrate the relations among regions at the species, genus and family levels (Jollivet, 1996; Tunnicliffe and Fowler, 1996; Tunnicliffe et al., 1996). Such a similarity analysis leads to several observations: seep and whale faunas fall outside the vent groups; similarity levels are generally low except between 13”N and 21”N; and the overall pattern reflects modem proximity. Distance between regions can be viewed two ways: either direct distance travelled by modern larvae or distance along ridge crests that represent a distribution pathway both past and present. While the difference may appear somewhat subtle, the processes involved are different. If larval communication among modern faunas dictates the similarities we see, then direct distance should be a better predictor of relationships. If past pathways and ridge connectedness is more important then perhaps a ridge distance is a better predictor. One is faced with a problem in fractal geometry in trying to assess ridge distances that may affect the outcome. Tunnicliffe et al. (1996) present such distances and perform a Mantel matrix comparison between each distance set and the similarity coefficients. Only ridge pathway was significant. Despite its position equidistant, by direct measure, from Galapagos and Juan de Fuca Ridge, 21”N shows a much stronger affinity to Galapagos. The longer ridge pathway and the barrier of North America explain its lower affiliation with the Juan de Fuca fauna.

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Long distance dispersal is still a possibility for many organisms. Examination of past routes may reveal patterns that are less specific to the vent fauna (Gebruk et al., 1997). A example is the Tethyan dispersal pathway that existed from late Cretaceous to early Tertiary (Figure 10). Many modern marine faunas contain elements that reflect this ancient pathway: corals, seagrasses and mangroves, among others (Ricklefs and Latham, 19943). A pathway through the equatorial regions may have provided a Mediterranean access to the Atlantic. The shallow ocean currents were probably from east to west (Specht, 1981). Such currents would not have carried larvae from the eastern Pacific to the Atlantic.

Figure 10 Configuration of the continents during the Paleocene-early Eocene. Arrows indicate hypothesized dominant current flow through the Tethys Sea. Shaded areas are continental masses; lines are continental shelves. (Adapted from Ricklefs and Latham, 1993.)

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7. THE DISTRIBUTION PAlTERNS OF TAXA

A closer examination of the distribution patterns among and within taxa should provide the answers to several questions such as: do species and genera show different patterns? Are non-endemic species more widespread? Do all taxa show the same distribution patterns? Are whole faunal patterns driven by just a few dominant taxa? We are still on the low end of sufficient data to answer such questions, but a preliminary analysis is useful.

7.1. Generic/Specific Distributions

There are not many species that have widespread distributions (Figure 11). Over 75% of vent species occur at only one site. Those that occur at four sites are mostly at Galapagos, nEPR sites and Guaymas Basin. An interesting four-site occurrence is the mite Copidognathus papillatus at Galapagos, 13"N, Juan de Fuca, and Lau/Fiji; mites have no pelagic dispersal stages. The limpet Lepetodrilus elevatus is currently recorded at five sites (Table 2); there is, however, some disagreement on identity of the western Pacific specimens based on molecular information (R. Vrijenhoek, pers. comm.). The species with the widest range is the ampharetid polychaete, Amphisamytha galapagensis (Figure 1 l), that is currently recorded from every vent site except the Atlantic (although it is recorded from the Florida seeps: Petrecca and Grassle, 1990). As expected, genera show wider ranges; about 40% are known from more than one site (Figure 11). Restricted ranges reflect the large number of monospecific genera described from the vent fauna. Figure 12 illustrates the numbers of genera that have different numbers of species. Most are in the one or two species per genus categories but a few have a large number of species described in the genus. Stygiopontius, the dirivultid copepod, holds the record at 22 species. The speciose genera (Figure 12) have the widest distributions. They have undergone an extensive radiation in the vent habitat. Do non-endemic taxa tend to have wider distributions? It seems not. No species found at four or more sites is currently known from the deep sea. The maldanid polychaete, Nicomache arwidssoni (four sites), and the ampharetid Amphisamytha galapagensis are also recorded at Florida seeps (Blake, 1985; Petrecca and Grassle, 1990). At the generic level, a quarter of the 28 genera found at four or more sites are also deep-sea genera. Figure 13 presents the distributions of gastropod families known from vents. While the sample size is not large, the most widely distributed families are vent

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Number of vent sites where one taxon is found Figure 11 Representation of the range of invertebrate vent species and genera. The number of sites at which each species and each genus occurs was tabulated. “Sites” are those listed in Table 2. The dominant pattern is that the great majority of species are found at only one site. Genera tend to have wider ranges.

endemics. The Provannidae include several species known from seeps and whale skeletons. The observations can be summarized as: 1. Most species records are single-site. 2. A few species have unusual distributions around the world. 3. Monospecific genera predominate. 4. Broadest distribution range occurs among the endemics, not the invading deep-sea species.

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Figure 12 Numbers of invertebrate genera with 1-22 species recorded in the genus. 135 of 232 genera have only one species described at vents. The most speciose genera are named. Of these, only the galatheid crab Munidopsis is a nonendemic deep-sea genus.

7.2. Pogonophorans

7.2.1. Distribution Obturate pogonophorans (vestimentiferans) are known only from habitats in which both oxygen and dissolved sulphide are available as hot or cold emissions. Ten species, representing eight described genera, are known from hydrothermal vents; eight species are restricted to vents, and two are also known from cold seeps (Table 2). One vent endemic perviate pogonophoran species, Siphonobrachia lauensis, is known from the western Pacific (Southward, 1991). There is a very high degree of endemicity for the vestimentiferans, with eight of ten species, six of eight described genera, and five of seven families known only from a single vent region. The northeast Pacific sites are inhabited by a single vent endemic species, Ridgeia piscesae, and one species also known at seeps. The East Pacific Rise province is distinguished by the presence of three vent endemic species, one of which (Rijtia pachyptila) is also found at Guaymas Basin. Ridgeia piscesae and Riftia pachyptila are the most widely distributed vent endemic pogonophoran species, each occurring at several sites within their provincial limits. Otherwise, only one other species,

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Scissurellidae Buccinidae Tumdae Trochidae Neolepetopsidae Neomphalidae Peltospiridae Provannidae Clypeosectidae Lepetodrilidae 0

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Figure 13 Gastropod families with the widest vent distributions. Families known also from the deep sea are not among the most widely distributed families at vents.

Oasisia alvinae on EPR, is known from more than a single vent site. Five vent endemic species characterize the western Pacific province. There are no pogonophorans known from the Mid-Atlantic Ridge hydrothermal sites. Lamellibrachia is the most widespread vestimentiferan genus (Figure 14); it has been reported from cold seeps on the east and west continental margins of both the Pacific and Atlantic Oceans, and also from two hydrothermal sites, L. columna at the Lau Back-arc Basin in the western Pacific, and L. barhami at Middle Valley on the Juan de Fuca Ridge (Van der Land and Nmrevang, 1975; Suess et al., 1985; Maiie-Garzon and Montero, 1986; MacDonald et al., 1989; Hashimoto et al., 1989; Jollivet et al., 1990; Southward, 1991; Dando et al., 1992; Williams et al., 1993; Miura et al., 1997). Cold seeps are also inhabited by species of Escarpia (Webb, 1969; Jones, 1985), one of which, E. spicata, is known from the Guaymas Basin (Jones, 1985). Considering the wide geographic distributions of L. barhami and E. spicata, their occurrence at single vent sites indicates that habitat features such as sedimentation, not geographic isolation, control their colonization of these sites.

Figure 14 Distribution of vestimentiferan tubeworm genera in vent, seep and other habitats. Each genus is represented by one symbol. Note the highly localized distribution of vent genera compared to the global spread of Lamellibruchiu. (Adapted from Southward et ul., 1996.)

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7.2.2. Development and Dispersal Vestimentiferans have separate sexes, and fertilization in some species is thought to occur just internal or external to the gonopores of the female (Southward and Coates, 1989). Maximum egg size for Riftia pachyptila is 105 pm, and spawning of gametes by Rijtia pachyptila has been observed in situ (Van Dover, 1994). Eggs of Lamellibrachia and Escarpia measure 105p m and 115 pm, respectively, and they develop into non-feeding, ciliated larva that remain in the water column during development (Young et al., 1996). A lecithotrophic larval stage is thought to provide the dispersive phase in the lifecycle of all vestimentiferans (Southward, 1988; Jones and Gardiner, 1989; Gardiner and Jones, 1994; Southward et al., 1996). It is thought that lamellibrachiid larvae may remain in the water column for at least three weeks (Young et al., 1996; Young, pers. comm. 1997); development times for other vestimentiferans are not known. However, during development the entrainment of larvae in hydrothermal plumes combined with tidal currents could potentially carry them as far as 50 km from their origin within a few weeks. No vestimentiferan larvae have been identified from plume plankton samples, and long-distance interbasin transport at this timescale seems unlikely (Southward and Galkin, 1997).

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7.2.3. Gene flow Studies of genetic differentiation among populations have focused on two vestimentiferan species, Riftia pachyptila and Ridgeia piscesae. In Rijtia pachyptila, the early allozyme studies of Bucklin (1988) indicate low genetic variability within populations at Galapagos Rift and 21"N. However, using a different suite of enzymes and polymorphism scoring criteria, Black et al. (1994) report relatively high levels of variation within Rijtia pachyptila populations, and gene flow estimates that dispersal among populations of this species occur at levels high enough to prevent significant population differentiation. The general absence of rare alleles at most polymorphic loci was suggested to be consistent with frequent restrictions in R. pachyptila population sizes, and the relatively high heterozygosity levels suggested a high intrinsic rate of population increase. Decreasing estimates of gene flow with increasing geographic distances between populations indicates that dispersal in this species follows the stepping-stone model of dispersal (Black et al., 1994). Ridgeia piscesae was originally considered to constitute two separate species. However, examination of morphological and allozyme variation among populations from the Juan de Fuca Ridge, Explorer Ridge and

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Gorda Ridge showed that it is a single species with great plasticity in tube morphology (Southward et al., 1995, 1996). Allozyme and RFLP studies of Ridgeia piscesae populations show that, as with Riftia pachyptila at EPR, along-ridge transport of R. piscesae in the NE Pacific occurs (Southward et al., 1996). However, unlike R. pachyptila, R. piscesae populations do not show an "isolation-by-distance" pattern of gene frequencies (Southward et al., 1996). The Juan de Fuca Ridge is a smaller system than EPR, which may account for this difference. The transform fault barriers between ridges in the NE Pacific are also small, and there is no genetic differentiation between R. piscesae populations from the Juan de Fuca Ridge and Explorer Ridge, which are separated by an offset of 160km (Southward et al., 1996). However, RFLP data show genetic differentiation between the Juan de Fuca Ridge and the Gorda Ridge, indicating that the offset of 360km separating these two ridges presents a real barrier to dispersal (Southward et al., 1996). 7.2.4. Phylogenetic Relationships

In one of the first phylogenetic analyses of vestimentiferan relationships, Williams et al. (1993) examined 322 bp of 28s rDNA for five vestimentiferan genera. While there are limitations to these data, the phylogenetic analysis showed strong support for a sister relationship between Ridgeia and Tevnia, but did not support monophyly of the Lamellibrachiidae (Lamellibrachia and Escarpia). A molecular phylogeny of the seep vestimentiferans collected around Japan (Lamellibrachia and Escarpia) based on the mitochondria1 gene, cytochrome oxidase I, supports monophyly of the Lamellibrachiidae (Kojima et al., 1997); however, only one unidentified perviate pogonophoran and one polychaete were included so the relationships of the two genera to other pogonophorans could not be assessed. Another recent study using cytochrome oxidase I to examine evolutionary relationships among vestimentiferans included 10 vestimentiferan species representing six genera (Black et al., 1997). Phylogenetic analysis of these data supported placement of the vestimentiferans as a clade within the Pogonophora, with Lamellibrachia as the basal taxon within the vestimentiferan clade, and a well-supported sister relationship between Ridgeia and Oasisia (Figure 15). Based on the levels of sequence divergence among vestimentiferans, the authors suggest that extant vestimentiferans constitute a recent radiation, perhaps less than 100 million years old, that diversified as a paraphyletic assemblage of seep-associated taxa and then gave rise to a clade of vent-endemic taxa (Black et al., 1997). The family Tevniidae, represented by two species, is restricted to the relatively young ridge systems of 13"N and 21"N, indicating recent evolution of this taxon.

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Paraivinellapalmiformis ( ~ c h - ) Figure 15 Phylogenetic hypothesis for vestimentiferans based on the mitochondrial gene, cytochrome c oxidase I DNA sequences. This tree suggests that vestimentiferans diversified as a paraphyletic assemblage of seep-associated taxa then gave rise to a clade of vent-endemic taxa. SJR = south Juan de Fuca Ridge; SER = south Explorer Ridge, 9"N and 18"s are one East Pacific Rise; WF = west Florida; GB = Guaymas Basin; LB = Lau Basin; MV = Middle Valley; OM = Oregon Margin. (After Black et al., 1997.)

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7.2.5. Conclusions The very high degree of endemicity for this group indicates barriers to dispersal among venting regions, and the distribution of vent endemic pogonophorans into provinces that lack shared species or genera is striking evidence that long-term tectonic processes have subdivided vent endemic pogonophoran communities. The lack of overlapping distributions between vestimentiferans of the northeast Pacific and East Pacific Rise reflects the isolation of these provinces following the subduction of the Farallon plate below the North American Plate 28 million years ago, while isolation of the western and eastern Pacific vestimentiferan fauna may date to the midTertiary. A planktonic larval stage, described for Lamellibrachia and Escarpia, and inferred for other vestimentiferans, provides a means of dispersal among sites within a vent region, and gene flow estimates support the expected pattern of isolation by distance for some populations of vestimentiferans. However, dispersal across ridge axes and over large geographic distances is not evident, which is not surprising given the important role vicariance has apparently played in the evolution of this group. The hypothesis that vent vestimentiferans are recently derived from a cold seep ancestor is supported by phylogenetic analysis of mitochondria1 data; further analyses of relationships between vent and seep taxa using different loci would be fruitful. Additional molecular data may also provide a robust phylogeny of vestimentiferan relationships needed to infer the branching patterns within the group, and to estimate divergence times. 7.3. Polychaete Family Alvinellidae

7.3.1. Distribution All known species in the family Alvinellidae are endemic to hydrothermal vents of the Pacific ocean. As with vestimentiferans, the degree of endemicity for alvinellids is very high. The genus Alvinella is known from three vent regions on the East Pacific Rise, while the second genus, Paralvinella, is pan-Pacific (Table 2). Of the nine described Paralvinella species, four are known from a single vent region in the eastern Pacific, and three are restricted to vents in the western Pacific. The distribution of alvinellids into provinces of vent regions is not as distinct as that of vestimentiferans. Four species are reported from the East Pacific Rise; however, one of them, Parallvinella grasslei, also occurs at Guaymas Basin, and a different subspecies of another, P. pandorae, is known from the northeast Pacific. P. pandorae is the only alvinellid

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known to cross provincial boundaries. The northeast Pacific sites are inhabited by four unique species, and the western Pacific is distinguished by the presence of two species. No alvinellids are known from the Mid-Atlantic Ridge. 7.3.2. Development and Dispersal Alvinellids studied to date are gonochoric and those species investigated are thought to have lecithotrophic larvae that undergo planktonic or direct development. Larvae or embryos of alvinellids are unknown, so their dispersal abilities are inferred indirectly from gamete morphology. Paralvinella palmiformis is thought to undergo demersal lecithotrophic development (McHugh, 1989). Direct development is suggested for P . pandorae, but no embryos or egg masses have been found in maternal tubes (McHugh, 1989, 1995). Internal fertilization followed by direct benthic development is inferred from sperm ultrastructure and large egg size (275pm) in Paralvinella grasslei (Zal et al., 1994, 1995). Fewer life history studies of Alvinella spp. have been undertaken. Chevaldonne and Jollivet (1993) suggest that a behaviour in which Alvinella spp. sometimes enter nearby tube openings may be a mating behaviour, in which case fertilization in these species may be internal or within the maternal tube. Maximum egg sizes in Alvinella pompejana and A . caudata are approximately 200 p m (Chevaldonne et al., 1997b), and probably lead to lecithotrophic larvae. 7.3.3. Gene Flow Several studies have used allozyme surveys to examine genetic divergence among alvinellid species, and genetic differentiation among alvinellid populations (Autem et al., 1985; Jollivet, 1993; Tunnicliffe et al., 1993; Jollivet et al. 1995a,b). The level of genetic diversity for alvinellids is high and Paralvinella spp. have higher percentages of polymorphic loci and observed heterozygosity than the two Alvinella spp. (Autem et al., 1985; Jollivet, 1993; Tunnicliffe et al., 1993; Jollivet et al., 1995b). The unstable nature of vent habitats may be responsible for maintenance of this high genetic diversity. Relatively high levels of genetic differentiation among populations of Alvinella pompejana, A . caudata, and Paralvinella grasslei from vents within venting regions along the East Pacific Rise seem to confirm the low dispersal ability of alvinellids (Jollivet et al., 1995a). However, genetic variation among disjunct populations separated by at least 1000 km is of the same magnitude as that found within the 13"N site; Jollivet et al. (1995a) propose

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that this may reflect uniform balancing selection on the enzyme loci, driven by the rapid aerobic-anaerobic alternation of vent conditions. ChevaldonnC et al. (1997b) examine whether the evidence of high genetic exchange among populations could be reconciled with the inferred limitations of larval dispersal in alvinellids; according to their model, dispersal between sites could not be achieved given what is known about intersite distances, bottom currents and alvinellid life histories. The authors propose two alternative explanations for this discrepancy; either larval dispersal can occur, or the flickering of hydrothermal activity at sites along ridge axes over geological time allows short-range dispersal processes to maintain gene flow (ChevaldonnC et al., 1997b). To decide between the two alternative explanations, additional information on the reproductive biology and development of alvinellids is needed. 7.3.4. Phylogenetic Relationships Hypotheses of alvinellid relationships based on distance analysis of allozyme data (Jollivet, 1993; Jollivet et al., 1995a), distance and parsimony analyses of 28R rRNA (Ftral et al., 1994), and evolutionary taxonomy (DesbruyZres and Laubier, 1991, 1993) have been presented. All support monophyly of the Alvinellidae, although the sister group for the family has not been firmly established. Monophyly of the two alvinellid genera is also supported, and the various hypotheses present no conflicts regarding the relationships among alvinellid species. Figure 16 represents a summary tree of evolutionary relationships of alvinellid species. Because six of the 11 species are site-endemic, little insight can be gained from an area cladogram for the group. The two Alvinella species overlap entirely in their geographic distribution. However, the Paralvinella species within each region or province do not form monophyletic groups. Instead, there is a pattern of pan-Pacific sister relationships for Paralvinella species, indicating that parallel speciation events have occurred in four separate lineages of the genus. The cladogenetic events leading to these pan-Pacific sister groups may reflect vicariant events stemming from the isolation of the western and eastern Pacific ridge systems in the Tertiary (Tunnicliffe et al., 1996). Alternatively, the speciation events may have occurred since that time, in which case pan-Pacific dispersal for colonization and subsequent isolation would be invoked to explain the pattern. The separation between subspecies of Paralvinella pandorae and between the sister species, P . palmiformis and P . grasslei, probably occurred approximately 25 million years ago, when the Farallon plate was subducted below the American plate. Assuming a separation time of 35 million years ago for these events, and knowing the genetic divergences between these taxa,

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P. unidentata

Western Pacific

P.hessleri

Figure 16 A summary tree of evolutionary relationships of alvinellid species based on the hypotheses of Desbruyeres and Laubier (1991, 1993), Jollivet (1993), FCral et al. (1994), and Jollivet et al. (1995a). A. = Alvinella; P . = Paralvinella. The diagram indicates the major groupings of the alvinellids as interpreted from morphological and genetic information. Note the pattern of pan-Pacific sister relationships for Paralvinella species, indicating that parallel speciation events have occurred in three separate lineages of the genus.

Jollivet et al. (1995a) estimate that the three clades of Paralvinella species diverged approximately 150 million years ago, during the Mesozoic. This implies that the Alvinella-Paralvinella split is even older; however, the distribution of Alvinella species is restricted to the relatively young ridge systems at 13"Nand 21"N. 7.3.5. Conclusions The contrasting patterns of distribution between the vestimentiferans and the alvinellids apparently reflects the more ancient divergences within the alvinellid clade. As has been done for vestimentiferans (Black et al., 1997), molecular clock analyses could test the timing of alvinellid divergences and

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provide evidence to discriminate between the vicariance hypothesis and the dispersal hypothesis for the pan-Pacific sister group relationships. Future investigations of the reproductive biology of additional species, and estimates of gene flow and genetic divergence based on DNA sequences, will strengthen the inferences we make regarding the biogeography of this group.

7.4. Polychaete Family Ampharetidae

7.4.1, Distribution Five species of ampharetid polychaetes have been reported from deep-sea hydrothermal vents. Amphisamytha galapagensis has the widest known geographic range of all species reported from hydrothermal vents spanning the Pacific and one seep location. A congeneric species, A . fauchaldi, is known from the Guaymas Basin on the EPR (Solis-Weiss and HernandezAlcantara, 1994). Three other ampharetids species are vent endemics known from only one vent region (see Table 2); two of these species represent monotypic, vent endemic genera.

7.4.2. Development and Dispersal Amphisamytha galapagensis is gonochoristic and all stages of egg and sperm development are present in populations sampled at different times of the year. Eggs reach 240 pm in diameter and fecundity is relatively high; sperm ultrastructure is of the ectaquasperm type. From comparison with other polychaetes, A . galapagensis appears to undergo continuous reproduction with external fertilization resulting in demersal, lecithotrophic larvae (McHugh and Tunnicliffe, 1994). Although adult distribution is wide, there is no indication that long range dispersal is accomplished by the presumably short-lived demersal larvae. No studies of population genetics in Amphisamytha galapagensis have been undertaken. Having such a wide distribution pattern, A . galapagensis could provide a great deal of insight into the biogeography of hydrothermal vent fauna. No phylogenetic analysis of vent associated ampharetids and their relationships to non-vent species has been published either. Further research on the ampharetids is needed.

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7.5. Gastropoda

7.5.1. Distribution Overall, endemicity of hydrothermal vent gastropods is high with 86% of species and 61YOof genera restricted to vents. Hydrothermal vent gastropod communities can be divided into five provinces that lack any shared species: the northeast Pacific (Juan de Fuca, Explorer, Gorda Ridges), Guaymas Basin, the western Pacific (MBJ + LFM), the mid-Atlantic Ridge and the East Pacific Rise/Galapagos Rift system. Gastropod faunas within the five provinces are generally homogeneous throughout each province (also discussed in Tunnicliffe, 1988; Waren and Bouchet, 1993). Molecular estimates of gene flow indicate communication between molluscan populations along the East Pacific Rise and between the East Pacific Rise and the Galapagos Rift (Vrijenhoek et al., 1994; Craddock et al., 1995b, 1997; Karl et al., 1996). Species distributions reflect this trend less than 30% of species found at any one location in the East Pacific Rise/ Galapagos Rift system are localized endemics. Waren and Bouchet (1993) found the East Pacific Rise and Galapagos Rift gastropod faunas indicative of a single province and we concur. Species of the vent endemic families Neomphalidae, Peltospiridae, Lepetodrilidae, Neolepetopsidae (also known from hydrocarbon seeps), Clypeosectidae, Provannidae (also known from hydrocarbon seeps), and the vent endemic genera Bathymargarites (Trochidae) and Solutigyra (Skeneidae) are known at multiple locations on the East Pacific Rise/Galapagos Rift system. Dispersal ability appears sufficient for a large part of the gastropod fauna of this region to overcome the distances between vent fields. While genetic evidence supports communication of vestimentiferans of this region with those of Guaymas Basin (Black et al., 1994), the gastropod fauna of Guaymas Basin is distinct, with only five species of gastropods, three locally endemic and two also known from hydrocarbon seeps or sunken whale bone. Three of the four genera at Guaymas Basin, are known from hydrocarbon seeps, whale falls, or sunken plants (Hyalogyrina, Pyropelta, Provanna) and the fourth has a global vent distribution (Lepetodrilus). Accepting that species distributions and genetic evidence support extensive dispersal of larvae within the East Pacific Rise/Galapagos Rift system, it is likely that sedimentation, not geographic isolation, controls gastropod faunal composition at Guaymas Basin. While tectonic movements have isolated the northeast Pacific vent gastropod communities from those of the East Pacific Rise/Galapagos Rift such that they now do not share any gastropod species, we confirm Waren and Bouchet’s (1993) finding that geographic distance has not completely isolated the different gastropod communities of the Western Pacific.

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Ignoring the anomalous Lepetodrilus elevatus, four vent endemic gastropod species range from MBJ to LFM. These are the exception as 85% of LFM and 74% of MBJ species are localized endemics, conforming to Beck’s (1992) conclusion that tectonic vicariance over the last five million years has geographically isolated populations of the northwestern and southwestern Pacific. Beck (1992) suggests that vicariance has led to allopatric speciation within the neomphalid genus Symmetromphalus. Molecular genetic investigations of the broadly distributed western Pacific species would be very useful. The gastropod fauna of the western Pacific differs from that of the EPR in that a higher proportion of families known outside vents have invaded western Pacific vent communities and a higher proportion of species have congenerics from the adjacent deep sea, hydrocarbon seeps or whale bones. The same is true for gastropod communities in the northeast Pacific. The faster rate of spreading at the East Pacific Rise may present more severe physiological extremes or more frequent extinction/recolonization events than at other ridge and back-arc systems, thus presenting a more restrictive environment for invasion. The Mid-Atlantic Ridge gastropod fauna appears small despite considerable exploration - only eight species are known although more are under description (A. WarCn, pers. comm.). 7.5.2. Patterns of Endemism Globally, gastropod species found at vents can be divided into three groups:

1. those with deep-sea congenerics or confamilials and representing genera with provincial distribution at vents; 2. representatives of groups found in a variety of sulphide-rich habitats; 3. members of extensive in situ radiations. With only one exception, endemic genera from families found outside sulphide-rich environments are restricted to a single ridge or back-arc basin system. The exception is the trochid genus Fucariu, found at LFM and the Juan de Fuca Ridge. The phenacolepadid genus Shinkailepas is found at both LFM and MBJ and the skeneid Solutigyra and trochid Bathymargarites are found at multiple sites along the East Pacific Rise. Some regionalism in invasion is apparent: the Phenacolepadidae, Turbinidae, Elachisnidae and Acmaeidae have only invaded western Pacific vents. The families Neolepetopsidae, Provannidae, Hyalogyrinidae, Xylodisculidae, and the cocculiniform families Pseudococculinidae, Cocculinidae, and Pyropeltidae are found in a variety of sulphide-rich environments such as hydrocarbon seeps and biogenic substrata such as whale

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bone and wood (McLean, 1990a, 1991, 1992a; Waren and Bouchet, 1993; Waren et al., 1993). The Hyalogyrinidae and Xylodisculidae are also known from shallow water sunken wood (Waren and Bouchet, 1993; Waren et al., 1993). The distribution of genera from families that are found in multiple habitats cannot be expected to be informative about historical vent connections - dispersal pathways other than ridge and back-arc basin systems have been available to them. Noting the disjunct distribution of the neolepetopsid genus Paralepetopsis (Florida seeps and Edison Seamount vents), Beck (1996) hypothesizes a common ancestor of the two species existed when the Isthmus of Panama was open. The implication is that species of the genus could have historically dispersed between vent sites by use of hydrocarbon seeps at continental margins. It will be interesting to see if species of Parafepetopsis are discovered at south American continental margins or at Indian Ocean spreading centres. Molecular evidence (A.G.M., in progress) shows that the Pyropeltidae is actually a derived group of the Pseudococculinidae, a family known from multiple biogenic substrata. The distribution of the Pyropeltidae reflects its phylogenetic origins as it is known from vents, seeps, and sunken whale bone - it appears to be tolerant of a wide range of sulphide-rich habitats. Of the strictly vent-endemic genera, the greatest number are shared along the East Pacific Rise and between the EPR and Galapagos Rift, following the pattern of common species. Unlike any other region, the EPR/Galapagos has two gastropod families that are strictly endemic, the Gorgoleptidae and Cyathermiidae (a subset of Neomphalidae delineated by McLean, 1990b). Like Wartn and Bouchet (1993), we find that the most species-rich and widespread families are only known from chemosynthetic environments. The EPR/Galapagos region shares four genera with the northeast Pacific and three with southwestern Pacific (Lau, Fiji and Manus back-arc basins). Other than the anomalous distribution of Fucaria, the northeast Pacific or MBJ vents do not share species with communities on the other side of the Pacific. The distribution of endemic gastropod genera suggests historical connection via the southern Pacific. The species of Rhynchopelta in the western Pacific is being redescribed as Lepetodrilus (L.A. Beck, pers. comm.), a genus known from all vent communities. The relationships of the Mid-Atlantic Ridge gastropod fauna are hard to judge because of the few species; ecology may be more important than history. The MAR shares the clypeosectid genus Pseudorimula with southwestern Pacific vents (as noted by McLean, 1990b, 1992b) but shares the peltospirid genus Peltospira with both sides of the southern Pacific.

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7.5.3. Development, Dispersal and Gene Flow Vent gastropods were very important for the initial examinations of reproduction, dispersal, and maintenance of vent populations. Studies of larval shell morphology predicted that the majority of vent gastropod species had short-lived lecithotrophic larvae (e.g. WarCn and Bouchet, 1989, 1993) and thus dispersal among vents was likely to follow a stepping-stone model. This implied that gastropod demes were geographically small and that the short lives of vents at spreading centres such as the East Pacific Rise and Galapagos Rift would result in frequent extinction of local gastropod populations, with the subsequent loss of genetic diversity. Susceptibility to local extinction argues against long-term survival of vent species. Species would arise by allopatry, be isolated by continued vicariance, and become extinct owing to cessation of local venting or other local crises (e.g. megaplumes). Contrary to these expectations, molecular evidence supports extensive within and among-ridge gene flow for vent molluscs (Vrijenhoek et al., 1994; Craddock et al., 1995b, 1997; Karl et al., 1996). Additional evidence of long-term in situ survival of vent gastropod lineages (McArthur and Tunnicliffe, 1998), suggests that high dispersal and gene flow greatly reduce the influence of local geological processes upon lineage survival. It may be that only long-term tectonic processes - plate and continental migration may subdivide vent gastropod populations and that provinciality may be the primary contributor to species diversity. Preliminary molecular investigations suggested possible stepping-stone dispersal for Lepetodrilus elevatus but sampling was limited (Craddock et al., 1997). A related species, L. pustulosus, exhibits no isolation of populations by distance (Craddock et al., 1997). L. elevatus is reported from the western and eastern Pacific (Table 2). Given the distances involved and the lack of cosmopolitan distribution of any other vent gastropod species, L. elevatus may actually consist of multiple species (Beck, 1993; War& and Bouchet, 1993). 7.6. Bivalves

The bivalve fauna found at vents is represented by symbiont-associated families also known from other sulphide-rich habitats (McArthur and Tunnicliffe, 1998). Nearly all species are restricted to one region but most genera are widely distributed. Molecular data support extensive gene flow for both Bathymodiolus thermophilus and Calyptogena magni$ca within and among the East Pacific Rise, Galapagos Rift and Guaymas Basin vent communities (Craddock et al., 1995b; Karl et al., 1996). But the unreliability of the current taxonomy of vesicomyid species is also evident (Vrijenhoek et

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al., 1994). Moraga et al. (1994) illustrate gene flow between western Pacific Bathymodiolus at Lau and North-Fiji back-arc basins with no genetic continuity of these with Bathymodiolus thermophilus populations on the East Pacific Rise. Craddock et al. (1995a) demonstrate that vent Bathymodiolus probably originated from seep ancestors but there is no genetic communication between modern vent and seep populations. Their results are ambiguous about the affinity of the Mid-Atlantic Ridge Bathymodiolus to Pacific species.

7.7. Copepods

Several groups of copepods are recorded from vents but not all have been studied thoroughly. Their retrieval requires retention of washings on a fine screen and assiduous picking. The copepods in the Family Dirivultidae represent a full 10% of the vent species: Arthur Humes has described 47 species in 17 publications. This family has but one representative outside vents - a separate genus described from a seep vestimentiferan off California. Figure 17 illustrates the distribution of the described species. There is a plethora of species at the 9"N to 21"N sites (Guaymas adds only two species). Stygiopontius comprises 22 species world-wide; it is the most speciose genus. Aphotopontius is found at all sites except the western Pacific. It is quite likely that more copepods will be found in the western Pacific basins, but currently copepods are reported from only two sites: Marianas and Lau (Humes, 1990, 1991). One genus and species is known to occur at both locations. Stygiopontius is represented by several western Pacific species. Rather astonishingly, one of those species from the Marianas, S. pectinatus, is known also from the Atlantic at both TAG and Snake Pit where it is the most abundant species. One other curious distribution is the occurrence of Aphotopontius forcipatus at both Snake Pit and Juan de Fuca Ridge. Humes (1991) discusses the eastern Pacific pattern of copepod distribution, including several other families of siphonostomes and poecilostomes. He notes the peculiar absence of Stygiopontius from Galapagos and invokes habitat as an important factor reducing copepod diversity at Guaymas. Dispersal by means of a nauplius or copepodid larval stage is inferred. While initial conjecture may have placed dirivultids as parasites of vestimentiferans, the diversity of associated fauna and habitats observed over the years suggests a variety of strategies that may include grazing of bacteria (Dinet et al., 1988). We have found these animals in a wide range of habitats and at several stages of vent succession. There appears to be a large centre of diversification for dirivultid copepods on the northern East Pacific Rise. As the same phenomenon can be

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Figure 17 Distribution of copepods in the family Dirivultidae. Each symbol represents a different genus (all of which are vent endemics). Under each symbol is the number of species in each genus at that site: at the northern East Pacific Rise sites, there are nine genera with a total of 28 species. The two most speciose genera are illustrated. (Adapted from Humes, 1987.)

noted in several other groups, there may be a common cause rooted in historical or habitat factors. Indeed, copepod diversification may have followed that of other groups. Investigation into the phylogenetic versus biogeographic relationships of the siphonostome copepods could be a fruitful future endeavour.

8. PATTERNS IN DIVERSITY

An interesting challenge over the next years will be to examine the causes of faunal differences among vents. Current concerns over global “biodiversity” have focused attention of ecologists on the underlying causes of patterns in diversity. We address a few of the interesting issues in this section but the reader can find another approach in Jollivet (1996).

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8.1. Gradients

The general diversity pattern for vent fauna from eastern Pacific to western Pacific is from higher to lower diversity. This gradient is contrary to that shown by shallower faunas. The Indo-West Pacific region is much richer than the eastern Pacific (and Caribbean) in mangroves (Ricklefs and Latham, 1993), corals and seagrasses (McCoy and Heck, 1976) and molluscs (Vermeij, 1978). Many explanations exist and now recent workers focus on higher origination rates in the west and extinction events in the east associated with Panama uplift and glaciations (Ricklefs and Latham, 1993; Jackson et al., 1993). Such events would not be expected to influence the deeper faunas of vents. Rate of origination could be an explanation for the higher eastern Pacific diversity. The underlying phenomena need closer examination. Latitudinal gradients exert strong controls on the marine fauna of shallow waters (Huston, 1994). While there is some argument about the strength of such controls in the deep sea, latitudinal gradients are demonstrated for gastropods in the Atlantic (Rex et al., 1993). If such gradients exist only in response to differences in surface production, then latitude may not be an important component of biogeographic patterns at vents. Nonetheless, as the mid-ocean ridge system spans arctic to equatorial regions, the opportunity to test is available. Figure 18 presents patterns in diversity in the eastern Pacific. The curves for the three largest groups of vent taxa are not consistent and do not peak at the equator. But there are many confounding factors such as areas sampled and complexity of habitat at the sites graphed. The matter remains open to a clearer test. Water depth is variable and of arguable importance (Van Dover et al., 1993). Vent depths range from 680m at Minami-Ensei off Japan and less than 850m at Menez Gwen site to 3600m depth at TAG and Snake Pit on the Mid-Atlantic Ridge. Because pressure increases by one atmosphere every 10m depth, it seems plausible to consider it as a factor in adaptation to life in the deep sea. Early work recorded marked declines in metabolic activity rates with depth in fish and crustaceans (see Childress et al., 1990); hypotheses of food limitations (Smith and Hessler, 1974), enzymatic adaptations (Somero and Siebenaller, 1979). These and other adaptations were considered to be induced by pressure. With the discovery of the hydrothermal vents, the opportunity arose to look at the effects of depth on metabolism when food availability and low temperature were not covariables. In studies on vent crabs, fish, tubeworms and bivalves, it has been found that metabolic activities are comparable to those of the shallow-water counterparts (Smith, 1984; Childress and Mickel, 1985; Felbeck et al., 1985; Childress et al., 1990). Thus depth-related metabolic effects are not apparently expressed at vents although they appear important in active pelagic

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Figure 18 Diversity of vent faunas of the eastern Pacific plotted against latitude. Solid dark line is total fauna from southern EPR, Galapagos, 13"N,21"N,Guaymas Basin and northeast Pacific. Circles are numbers of gastropod taxa, triangles are polychaetes, and squares are crustaceans.

non-vent groups for reasons associated with their mobile life-style (Childress, 1995). The depth zonation patterns reported in the deep sea (cf. reviews in Advances in Marine Biology, volume 32) are evidently not related to physical factors. Patterns among groups and geographic locations are not consistent, so that causal factors are difficult to define (Carney et al., 1983; Gage and Tyler, 1991).

8.2. Global Differences

There are clear differences in species diversity among vent regions. Collection effort is highly variable but the patterns found have remained consistent over the last decade. The northern East Pacific Rise holds the greatest number of species. From 21"N to the Guaymas Basin spans 6" latitude, bare rock and sediment, in two major sites; there are 121 recorded species (Table 2). In comparison, from Juan de Fuca to Explorer Ridge

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spans 5.5” latitude, bare rock and sediment, in five major sites (Table 1): yet there are but 80 recorded species. As systematic effort remains to be matched in other areas of the world it is unwise to speculate extensively. The northern EPR may have sponsored more extensive diversification than other regions. “Diversity anomalies may arise historically from region-specific differences in the origin of clades, rates of diversification within clades, propensity for dispersal between regions, extinction, or some combination of these” (Ricklefs and Latham, 1993; p. 221).

8.3. Causes 8.3.1. Speciation and Extinction

Processes that promote speciation sponsor differences in diversity. Ridge segmentation probably enhances separation of populations and increases divergence. Jollivet (1996) and Vrijenhoek (1997) summarize genetic studies to examine different models of gene flow and many of these have been addressed in preceding sections. A surprising number of studies find little genetic differentiation between populations spread along a ridge crest. Metapopulation dynamics and high gene flow may be one explanation (Vrijenhoek, 1997) but Jollivet et a f . (1995a) suggest that the fluctuating vent conditions impose a stabilizing selection that maintains a degree of similarity among populations. Genetic studies between ridges, on the other hand, confirm the role of major disjunctions in curtailing gene flow (Jollivet, 1996). Figure 19 summarizes the interaction between the hydrothermal vent conditions and the various factors that inhibit and promote speciation. Extinction must play an important role in determining diversity but there is little available information. Because the stability of the vent habitat is linked to an ephemeral fluid source, local extinctions are common. Extinction of a segment fauna will occur when the underlying magma chamber cools. An entire region may be affected if spreading dynamics change. Dating of Mid-Atlantic sulphides reveals episodicity in venting (Lalou et a/., 1995) so local extinction at these widespread sites must have occurred often. The modern Mid-Atlantic fauna may be a product of widespread extinctions during quiescent periods (Juniper and Tunnicliffe, 1997). Indeed, the Atlantic fauna as a whole may be relatively young with secondary invasions.

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Figure 19 Summary diagram to illustrate the factors presumed to induce speciation in the hot vent environment. (Adapted from Jollivet, 1996.)

8.3.2. Ecological Controls Mantle and crustal processes control the distribution and longevity of vent communities. Juniper and Tunnicliffe (1997) examine the implications of these processes for the development of vent communities. They use the model of Baker et al. (1995) to relate spreading rate to vent habitat frequency. More abundant habitat will sponsor a larger number of species (Ricklefs, 1987). But as greater habitat frequency will also facilitate gene flow one cannot always predict which factor will promote a greater regional species pool. One may expect greater between-site differences on the MidAtlantic Ridge, but perhaps a smaller overall species pool. Differences in overall diversity may have many explanations, all of which demand closer examination. The island nature of hydrothermal vents makes them an ideal system in which to study the local to regional scale of diversity. Figure 20 summarizes some of the ways in which diversity may vary with different habitat conditions. Eruption frequency and habitat stability vary among biogeographic regions so much that disturbance may be a major factor influencing diversity (Tunnicliffe et al., 1997).

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LOW >

-

%

Disturbance

Heterogeneity

Habitat Area

Regional Age High

>-

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Figure 20 Summary representation of the influence of major physical controls on diversity of ecological communities. (From Juniper and Tunnicliffe, 1997.)

Habitat heterogeneity is increased if there is a range of substrata, of venting dilutions, and of depths within a ridge. Some regions have multiple extended vent fields while others are quite limited. Regional age can vary from long-established ridges to relatively recent propagating rifts. Diversity controls are not only habitat phenomena. Ricklefs (1987) presents a compelling argument concerning the inadequacy of “local processes”, by which he means competition and predation for the most part, to explain large differences among regions. He argues strongly for the examination of the origin and role of regional species pools and how they regulate local species pools. In concert with current concepts in diversity studies, there are many larger aspects to consider: differential extinction, differential clade origins, varying rates of diversification and the history of the region (Ricklefs and Schluter, 1993). There remain many challenges in vent biogeography.

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9. SUMMARY

The hydrothermal vent fauna presents an unusual opportunity to test current concepts in biogeography and diversity regulation. The island nature of the habitat, the dependence on a geophysically controlled energy source, the separation of the fauna from the surrounding deep sea and the regionalization around the globe constitute a very interesting system. Here, we can compare ecological and historical roles in species accumulation. The relative importance of local versus regional controls of diversity regulation should be possible to assess. It is easy to make a list of the shortcomings of current vent studies topping the list is the unevenness of collection efforts and systematic studies. Nevertheless, the discipline has come a very long way from early exploration and description efforts. New techniques have made an important contribution such as construction of phylogenies from molecular data and experimental approaches to measurement of larval dispersal. Expansion of regional studies by the entry of various national efforts to this deep-sea science has introduced new theoretical and practical approaches. Training of young scientists remains a priority and incorporation of models from non-vent disciplines will contribute vital advances in hydrothermal vent biogeography.

ACKNOWLEDGEMENTS We thank Alan Southward for his encouragement in writing this review and the helpful comments of two reviewers. As always, the help of many systematists involved in vent work must be acknowledged. Thoughtful insight from S.K. Juniper was welcome. The support of L. Franklin in production of the manuscript was vital. A.G.M. was supported by a Smithsonian Institution Postdoctoral Fellowship.

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442

V. TUNNICLIFFE €1 AL.

Vrijenhoek, R.C. (1997). Gene flow and genetic diversity in naturally fragmented metapopulations of deep-sea hydrothermal vent animals. Journal of Heredity 88, 285-293. Vrijenhoek, R.C., Schutz, S.J., Gustafson, R.G. and Lutz, R.A. (1994). Cryptic species of deep-sea clams (Mollusca, Bivalvia, Vesicomyidae) in hydrothermal vent and cold-water seep environments. Deep-sea Research 41, 1171-1 189. Wartn, A. and Bouchet, P. (1989). New gastropods from east Pacific hydrothermal vents. Zoologica Scripta 18, 67-102. Wartn, A. and Bouchet, P. (1993). New records, species, genera, and a new family of gastropods from hydrothermal vents and hydrocarbon seeps. Zoologica Scripta 22, 1-90. Wartn, A., Gofas, S. and Schander, C. (1993). Systematicposition of three European heterobranch gastropods. The Veliger 36, 1-1 5 . Webb, M. (1969). Lomellibrachia barhami, gen. nov., sp. nov. (Pogonophora), from the northeast Pacific. Bulletin of Marine Science 19, 19-47. Wegener, A. (1924). The Origin of Continents and Oceans. Methuen Press, London, UK. Weijermars, R. (1989). Global tectonics since the breakup of Pangea 180 million years ago: evolution maps and lithospheric budget. Earth-Science Reviews 26, 11 3 162. Williams, A.G. and Dobbs, F.C. (1995). A new genus and species of caridean shrimp (Crustacea: Decapoda: Bresiliidae) from hydrothermal vents on Loibi Seamount, Hawaii. Proceedings of the Biological Society of Washington 108, 228-237. Williams, N.A., Dixon, D.R., Southward, E.C. and Holland, P.W. (1993). Molecular evolution and diversification of the vestimentiferan tube worms. Journal of the Marine Biological Association of the United Kingdom 73, 437452. Young, C.M., Vazquez, E., Metaxas, A. and Tyler, P.A. (1996). Embryology of vestimentiferan tube worms from deep-sea methane/sulfide seeps. Nature, London 381, 514-516. Zal, F., Desbruykres, D. and Jouin-Toulmond, C. (1994). Sexual dimorphism in Paralvinella grasslei, a polychaete annelid from deep-sea hydrothermal vents. Comptes Rendm de I'Academie des Sciences de Paris, Sciences de la vie 311,4248. Zal, F., Jollivet, D., Chevaldonnk, P. and Desbruykres, D. (1995). Reproductive biology and population structure of the deep-sea hydrothermal vent worm Paralvinella grasslei (Polychaeta: Alvinellidae) at 13"N on the East Pacific Rise. Marine Biology 122, 637-648. Zhirmunsky, A.V. and Tarasov, V.G. (1990). Unusual marine ecosystem in the flooded crater of Ushishir Volcano. Marine Ecology Progress Series 65, 95-102.

Taxonomic Index Note: Page references in italics refer to Figures; those in bold refer to Tables

Acanthaster planci, 230 Acanthephyra purpurea, 378 Acanthina angelica, 49 Acharax alinae, 374 Achatina fulica, 33, 34 Acmaea (Macclintockia) scabra, 32 persona, 32 (Tectura) digitalis, 32 (Tectura) tessulata, 12 Actinostola callosa, 365 Adipicola. 384 n. sp., 374 Aequipecten opercularis, 287 AIaysia spiralis, 366 Alvinella, 385, 410, 41 1, 412 caudata, 368,41 1, 413 pompejana, 368,411, 413 Alviniconcha hessleri, 372 Alvinocaris, 377 longirostris, 377 lusca. 377 markensis, 377 Alvinoconcha, 397 Amathys lutzi. 369 Ammodytes. 257 hexapterus, 270 Ammothea verenae. 374 Ampelisca spinipes, 287 Amphicrossus altalis. 375 Amphiduros axialensis, 366 Amphiplica gordensis, 370 Amphisamytha fauchaldi, 369, 414 galapagensis, 369, 402, 414 Amphiura, 287 filiformis, 235 Anatoma, 371 Ancylusjuviatilis, 18, 22, 37, 41

Anisodoris nobilis, 33 Aphotopontius acanthinus, 375 arcuatus, 375 atlanteus, 375 baculigerus, 375 jlexispina, 375 forcipatus. 375, 419 hydronauticus, 375 limatulus, 375 mammillatus. 375 probolus, 375 temperatus, 375 Apotectonia heterostegos, 377 Apristurus maderensis, 380 Archachatina marginata, 33 Archidoris montereyensis, 33, 49 pseudoargus, 33 Archinome (Euphrosine) rosacea, 366 Arcovestia ivanovi, 366 Arctica Blandica, 213, 214 Arctocephalus australis australis, 274 Arenicola marina, 2 1 1 Ariolimax columbianus, 9, 13, 22, 31, 33 Arion ater, 7 empiricorium, 33 Astarte, 221 Asterias rubens, 288 Astropecten irregularis, 235 Atapozoa, 33 Austinograea alaysae, 378 williamsi, 318 Axiothella millsi. 366 Balaena glacialis, 275 Balaenoptera

444 Balaenoptera (contd.) acutorostrata, 210 borealis, 264 musculus, 259 physalus, 215 Balanus glandula, 48, 49 Balistapus undulatus, 230 Barleeia, 24 Bathyacmaea jonassoni, 311 secuna'a, 311 Bathybdella sawyeri. 369 Bathycatalinajilamentosa, 361 Bathychaunax roseus, 380 Bathyhalacarus, 314 Bathykurila guaymasensis. 361 Bathymargarites, 4 1 5, 4 1 6 symplector, 311 Bathymodiolid n. sp., 314 Bathymodiolus, 391, 419 aduloides, 313 brevior, 373 elongatus, 313 japonicus, 313 n. sp.1, 314 n. sp.2, 314 n. sp.3, 314 platifrons, 313 puteoserpentis, 313 septemdierum, 313 thermophilus, 373, 4 1 8, 4 19 Bathypecten vulcani, 314 Benthoctopus, 380 Benthoxynus spiculifer. 315 tumidiseta, 315 Beryx splendens, 380 Bilhynia tentaculata, 3 1 Bouvierella curtirama, 376 Brachyramphus marmoratus, 210 Branchinotogluma burkensis, 367 grasslei, 361 hessleri. 367 sandersi. 367 Branchiplicatus cupreus, 367 Branchipolynot? pettiboneae, 367 seepensis, 367, 394 symmytilida, 361 Brevoortia tyrannus. 8 1

TAXONOMIC INDEX

Bruceiella globulus, 311 Buccinum, 313 undatum. 6, 18, 22, 33, 31, 41, 288 viridum, 312 Bucephala clangula, 210 Busycon canaliculatum. 4, 5 , 6,I unhtum, 5 Bythites hollisi. 381 Bythograea intermedia, 318 laubieri, 318 microps, 318 thermydron, 318 Calliostoma, 33 canaliculatum, 33 zizyphinum, 3 5 Calyptogena kilmeri, 313 magnijica, 313, 385, 391, 418 pacijica, 373 solidissima, 313 Cancer magister, 224 Candelabrum phrygium, 365 serpentarii, 365 Cantrainea jamsteci, 312 Capitella, 366 nr capitata, 366 Carcinus maenas, 289 mediterraneus, 8 1 Careproctus hyaleius, 381 Cassis tuberosa, 24 Cataetyx laticeps, 380 Cellana, 33 grata, 21, 28, 40 toreuma, 6, 22 Centroscymnus coelolepis, 380 Cepaea nemoralis, 22, 31, 41 Cepphus columba, 210 Ceramaster misakiensis, 319 Cerastoderma edule, 13, 24, 21 1 Ceratocephale pacijica, 367, 373 Cerianthus, 365 Ceuthoecetes acanthothrix. 315 aliger, 315 cristatus, 315 introversus. 315

TAXONOMIC INDEX

Chaceon aflinis, 378 Chaetozone n. sp.1, 368 n. sp.2, 368 Charonia lampas, 7 Chasmatopontius thescalus. 375 Chaunax, 380 Chelinus trilobatus, 230 Cheramyzon abyssale, 375 Chirondota n. sp., 379 Chiton pelliserpentis, 19,37, 38, 40, 41 Chondrophellia coronata, 365 Chorocaris, 377 chacei, 377 vandoverae, 377 Chthamalus anisopoma. 49 Cirroteuthis magna, 380 Clupea harengus, 264 pallasi, 265 Clypeosectus curvus, 371 delectus, 371 Cocculina cf. craigsmithi, 370 Coelorhinchus cf. labiatus, 380 Collisella digitalis, 6 (Macclintockia) scabra, 28, 46 scabra, 6 (Tectura) digitalis, 28 Concholepas concholepas, 22, 31, 41, 41 Copidognathus alvinus, 374 nautilei, 374 papillatus. 374,402 Corbiculafluminae, 24 Coris julis, 255 Corolla spectabilis, 3 1 Corophium, 44 Corymorpha pendula, 221 Crangon. 281 Crangon crangon, 21 1 Crassostrea gigas, 11 rhizophorae, 183 virginica, 15, 28, 183 Crepidulafornicata, 31 Ctenopelta porifera, 370 Cyanagraea praedator, 378 Cyanathea hydrothermala, 365 Cyanograea, 385

445 Cyathermia naticoides. 369 Cyprinus carpio, 84 Dahlella caldariensis, 376 Daphnia magna. 253 Dendroporna maximum, 3 1 Depressigyra globulus, 370 Desbruyeresia cancellata. 372 marianaensis, 372 melanioides, 372 spinosa, 372 Diadema antillarum, 231 Dicentrarchus labrax, 310 Dichelopandalus leptoceros, 22 1 Doriopsilla albopunctata, 33 Dreissena polymorpha, 1 1 Drupella. 230, 232 Dysidea fragilis, 33 Ecbathyrion prolixicauda, 376 Echinocardium cordatum, 213 Echinopelta fistulosa, 370 Echinus alexandri, 379 Ectenegena extenta, 373 Engraulis. 251 mordax, 276 ringens. 205 Enhydra h i s , 270 Ensis siliqua, 2 1I Eochionelasmus n. sp., 375 ohtai, 375 paquensis, 375 Eosipho desbruyeresi, 373 n. sp., 373 Epigonus telescopus, 381 Epinephelus guttatus, 254 striatus, 254 Erebonaster protentipes, 375 Escarpia, 405, 401, 408, 410 laminata, 409 spicata, 366,393, 405, 409 Etmopterus mirabilis, 380 princeps. 380 pusillus, 380 Euandania a f f . ingens, 377 Eulalia (Protomystides)papillosa, 367

446 Eulepetopsis vitrea, 310, 391 Eumetiopiasjubatus, 214 Eunice masudai. 366 northioidea, 366 pulvinopalpata. 366 Eunicella verrucosa, 223 Euonyx mytilus, 311 Euphausia superba, 264 Euphilomedes climax, 316 Exallopus jumarsi, 366 Exrima dolichopus, 315 singula, 315 Falcidens n. sp., 369 Fissuricola caritus, 316 Fossaria modicella, 20 Fratercula arctica, 268 Fucaria, 416, 411 n. sp., 311 striata, 311 Fucus distichus, 49 Fulmaris glacialis, 296 Fulmarus glacialis, 285 Funiculina quadrangularis, 223 Gadinalia nivea, 3 1 Gadus macrocephalus, 265 morrhua, 205 Gaidropsarus n. sp., 380 Galapagomystides aristata, 361 Galathealinum brachiosum, 409 Gitanopsis alvina. 316 Gleba cordata, 31 Glycera profundi, 366 Gorgoleptis emarginatus, 312 patulus, 312 spiralis, 312 Graneledone, 380 Grassleia hydrothermalis, 369 Haematopus ostralegus, 27 1 Halacarellus alvinus, 374 auzendei, 314 Halice hesmonectes, 311 Haliotis discus hannai. 48

TAXONOMIC INDEX

tuberculata, 23, 31, 38, 39, 40,41 Harmothoe macnabi, 361 Helcion pellucidum, 35 Helicoradomenia juani, 369 n. sp.1, 369 n. sp.2, 369 n. sp.3, 369 n. sp.4, 369 n. sp.5, 369 n. sp.6, 369 Helicrenion reticulatum, 372 Helix aspersa, 10, 32, 33 Hesiocaeca hessleri, 366 Hesiodeira glabra, 366 Hesiolyra bergi, 361 Hesiospina vestimentifera, 367 Hippoglossoides platessoides, 289 Hippoglossus hippoglossus, 240 Hirondellea glutonis, 311 Hirtopelta hirta, 310 Hirudo medicinalis, 409 Histrionicus histrionicus, 210 Homarus americanus, 224 Hoplostethus atlanticus, 210 Hyalogyra vitrinelloides, 372 Hyalogyrina, 415 grasslei, 312 n. sp., 312 Hyalopontius boxshalli, 316 Hydrobia ulvae, 24, 3 1 ventrosa, 31, 41 Hydrolagus mirabilis, 380 pallidus, 380 Hyphalion, 315 captans, 375 Hypoechinorhynchus thermaceri. 365 Hypselodoris webbi, 33 Idas (Idasola) washingtonia, 314 Idasola, 384 Ijiemerita nautilei, 372 Iheyomytilidicola tridentatus. 367 Ilionia prisca, 35 IIyanassa obsoleta, 10, 18,22, 26, 31, 38, 39, 40,41 Ilyophis blachei, 388 Imbracoverruca n. sp., 375 Iorania concordia. 311

TAXONOMIC INDEX

Iphionella n. sp.,367 risensis, 367 Isaacsicalanuspaucisetus, 375 Janthina, 35 Jaxea nocturna, 2 19 Lacuna, 25 variegata, 25 vincta, 25 Lacunoides exquisitus, 369 Lacydonia n. sp., 367 Laeviphitus japonicus, 372 Laitmatobius crinitus, 375 Lamellibrachia, 405, 406, 401, 408, 410 barhami, 366,394, 405, 409 columna, 366,405, 409 Laminatubus alvini, 368 Larus, 285 audouinii, 286 Laubieriellus grasslei, 368 Lebbeus, 378 carinatus. 377 washingtonianus, 378 Leitoscoloplos pachybranchiatus, 366 Lepetodrilus, 372, 41 5, 411 corrugatus, 371 cristatus, 371 elevatus, 371,402, 416, 418 fucensis, 371 guaymasensis, 371 japonicus, 372 ovalis, 372 pustulosus, 372,391, 418 schrolli, 372 tevnianus. 372 Lepidion ? schmidti, 380 Lepidonotopodium fimbriatum, 367 minutum. 367 piscesae. 367 rifttense. 367 williamsae, 367 Leptochiton tenuidontus, 369 Leptognathia ventralis, 377 Leptogyra, 371 injata, 371 Lethrinus, 219 Levensteiniella, 367 intermedia, 367

447 kincaidi, 367 raisae, 367 Limanda limanda, 287 Limax maximus, 8, 31 pseudopavus. 26 Lindaspio dibranchiata, 368 southwardorum, 368 Liocarcinus, 281 Lirapex granularis, 370 humata, 370 Lithothamnion, 2 1I Littorina aspera, 9 irrorata, 26, 45 littorea, 10, 13, 14, 19, 20, 29, 45, 50 sitkana, 24 Loligo vulgaris, 33 Lopholatilus chamaeleonticeps, 255 Lottia, 48 gigantea, 6,28, 46 Lucernaria n. sp., 365 Luckia striki, 377 Lunda cirrhata, 210 Lutjanus, 219 Lymnaea peregra, 31, 41 stagnalis, 19, 31 truncatula, 4, 6, 19 Macellicephala galapagensis, 367 Macellicephaloiaks alvini, 367 Macoma balthica, 28 Mallotus villosus, 251 Margarites shinkai, 371 Marianactis bythios, 365 Megaptera novaeangliae, 215 Melanodrymia aurantiaca, 370 brightae, 370 n. sp., 370 Melanogrammus aegle3nus, 262 Mercenaria mercenaria, 21 1 Merlangius merlangus, 262 Metapenaeus macleayi, 120, I46 Microstomus kitt, 3 10 Milax sowerbii, 23, 24, 33

448 Mirocaris (Chorocaris) fortunata, 311 keldyshi, 311 Moelleriopsis n. sp., 311 Monodonta turbinata, 22 Mora moro, 380 Munida magniantennulato, 378 Munidopsis. 318, 404 alvisca, 318 cf. crassa, 318 lauensis, 318 lentigo. 318 marianica, 318 starmer, 318 subsquamosa, 318 Mya arenaria, 2 1 7 Mycteroperca microlepis, 254 venenosa, 254 Mytilidiphila enseiensis, 361 okinawaensis. 367 Mytilus edulis, .13-15, 16, 17, 17, 18, 24, 27, 35 Myxicola infundibulum, 219 Nacella concinna, 5, 6, 21 Natica gualteriana, 31 Navanax inermis, 41 Nematonurus armatus, 380 Nembrotha, 33 Neobrachylepas relica, 314 Neocyttus helgae, 381 Neolaenira racemosa, 368 Neolepas n. sp.1, 314 n. sp.2, 314 n. sp.3, 314 rapanuii, 314 zevinae, 314 Neolepetopsis densata. 310 gordensis, 310 occulata, 370 verruca, 310 Neomphalus fretterae, 369 Neoverruca brachylepadoformis,315 Neptunea insulalis, 313 Nereimyra alvinae, 361 Nereis. 361 piscesae, 361

TAXONOMIC INDEX

sandersi, 361 Nicomache arwihsoni, 366,402 venticola, 366 Nilva torifera, 315 Nodopelta heminoda, 310 subnoda, 310 Nucella, 49 emarginata, 4, 6, 24, 28 lamellosa, 48 lapillus, 25 Nuculana grasslei, 314 Oasisia, 408 alvinae, 366,405, 409 Oenopota ogasawarana, 313 Olgasolaris tollmanni, 310 Olivelia columeliaris, 3 1 Oncaea praeclara, 315 Ophioctenella acies, 319 Ophryotrocha, 366 akessoni, 366 globopalpata, 366 platykephale, 366 Opisthotrochopodus alvinus, 368 japonicus. 368 marianus, 368 segonzaci, 368 trifurcus, 368 tunnicliffeae, 368 Oradarea longimana, 316 Orbiniella aciculata, 366 hobsonae, 366 Orchomene abyssorum. 311 distinctus, 377 Orcinus orca, 286 Oreochromis aureus, 85 mossambicus, 83, 84, 253 Orseis grasslei, 361 Otala lactea. 32 Owenia, 44 Pachycara gynmium, 381 thermophilum, 381 Pachydermia

TAXONOMIC INDEX

laevis, 370 sculpta, 370 Pagurus bernardus, 288 Palinurus marginatus, 224 Paracentrotus lividus, 23 1 Paracrangon n. sp., 378 Paralepetopsis, 4 17 n. sp., 370 rosemariae, 370 Paralomis. 378 jamsteci, 378 Paralvinella, 410, 41 1, 412 bactericola, 368,413 &la, 368,413 jjiensis, 368,413 grasslei, 368,41C12, 413 hessleri, 368,413 p , irlandei, 413 p. pandorae, 413 palmiformis, 369,409, 411, 412, 413 pandorae, 369, 410, 411, 412 sulJTncola. 369, 413 unidentata, 369, 413 Parasicyonis ingolfi. 365 Pardalisca endeavouri, 377 Parougia worfi, 366 Patella caerulea. 22 depressa, 13 longicosta, 48 ulyssopensis (P. aspera), 13 vulgata, 4, 6, 8, 9, 10, 11, 12, 12, 13, 19, 21, 28, 29. 30, 33, 34, 37, 38, 40,41, 44, 47 Pecten maximus, 21 1 Pectinaria, 44 Peltospira, 417 delicata, 370 lamellifera, 370 operculata, 370 sp.1, 370 sp.2, 370 sp.3, 370 Penaeus setiferus, 146 Pennatula phosphorea, 223 Pentapora foliacea, 223 Phalacrocorax aristotelis, 268 Phoca vitulina, 270 vitulina vitulina L., 275 Phocoena ahocoena. 275-, -

449 Phyllidia varricosa, 33 Phymorhynchus moskalevi, 373 sp.2, 373 sp.3, 373 sp.4, 373 starmeri, 373 warkni. 373 Physa acuta, 37, 41 Placopecten magellanicus, 14 Planorbidella depressa, 370 planispira, 370 Planorbis contortus, 18, 22, 37, 41 Pleuronectes platessa, 2 11 Poecilia reticulata, 253 Polyacanthonotus cf. rissoanus, 380 Poseidonamicus, 376 Posidonia oceanica, 222, 229 Prionospio, 368 sandersi, 368 Prionocoleberis styx, 376 Protis hydrothermica, 368 Protolira valvatoides. 371 Protomystides, 367 verenae, 367 Provanna, 372, 415 buccinoides, 372 glabra, 372 goniata, 372 ios, 372 laevis, 372 muricata, 372 nassariaeformis, 372 segonzaci. 372 variabilis, 372 Pseudorimula, 417 marianae, 371 midatlantica, 371 Puncturella n. sp., 371 parvinobilis, 371 rimaiazenaensis, 371 solis, 371 Pyropelta, 4 15 bohlei, 371 corymba, 371, 393 musaica, 371 Raia. 248 batis, 237

450 Raja (contd.) clavata, 310 Ralfia, 48 Rhogobius contractus, 375 pressulus, 376 Rhynchopelta, 417 concentrica, 370 nux, 370 Ridgeia, 408, 409 piscesae, 366,394, 398, 404, 4079 409 Riftia pachyptila, 366,398,404, 4079 408’ 409 Rimicaris aurantiaca, 377 exoculara, 377,398 Rimipontius mediospinver, 376 Rissa tridactyla, 268 Sabellaria spinulosa, 2 19 Sardina, 257 Sardinops sagax, 276 Sarotherodon mossambicus, 83, Saxipendium coronatum, 379 84-85 Scillaelepas n. sp., 374 Scoloplos ehlersi, 366 Scomber scombrus, 264 Scotoecetes introrsus, 376 Scyllarisdes squammosus, 224 Seba profundus, 377 Sebastes alutus, 256 Segonzacia, mesatlantica, 378 Semibalanus balanoides. 50 cariosus, 48, 49 Sericosura cochlevovea, 374 cyrtoma, 374 heteroscela. 374 mitrata, 374 venticola, 374 Serpulorbis squamigerus, 31 Shinkai longipedata, 367 semilonga, 367 Shinkailepas, 41 6 kaikatensis, 370 tufari, 370 Simenchelys parasitica, 380 Siphonaria, 45

TAXONOMIC: INDEX

atra, 21, 22, 40 japonica, 2 1, 22, 40 Sirius, 21, 22 Siphonobrachia lauensis, 365,404 Solea solea, 21 1 Solemya johnsoni, 374 Solutigyra. 415,416 reticulata, 371 Somateria mollissima, 271 Sphaerosyllis ridgensis, 368 Sphincterochila boisseri. 32 Squalus acanthias, 280 Steleuthera ecoprophycea, 577 Stercorarius skua, 268 parasiticus, 268 Stereolepis gigas, 237 Sterna paradisaea, 268 Strombus gigas, 23, 33 Strongylocentrotus droebachiensis, 233 Stygiopontius, 402, 419 appositus, 376 brevispina, 376 bulbisetiger, 376 cinctiger, 376 cladarus, 376 flexus, 376 hispidulus, 376 latulus, 376 lauensis, 376 lumiger, 376 mirus, 376 mucronferus, 376 paxillifr, 376 pectinatus, 376,41 9 quadrispinosus. 376 regius, 376 rimivagus, 376 sentfler, 376 serratus, 376 stabilitus, 376 teres, 376 verruculatus, 376 Sula banana, 268 Sutilizona theca, 371 Symmetromphalus, 416 hageni, 369 regularis, 369 Synaphobranchus kaupi, 380 Systellapsis braueri, 378

TAXONOMIC INDEX

Tapes philippinarum, 2 1 1 Tectovalopsis diabolus, 377 wegeneri. 377 Tectura (Collisella) scutum, 49 Temnocinclis euripes, 371 Temnozaga parilis, 371 Tetraselmis suecica, 28 Tevnia, 408 jerichonana, 366,409 Thalassoma bifascia turn, 255 Thermaloniscus cotylophorus, 377 Thermanermertes valens, 365 Thermarces cerberus, 381 Thermiphione jijiensis. 368 tufari, 368 Thermobiotes mytilogeiton, 380 Thermochiton undocostatus, 369 Thermopalia taraxaca, 365 Thermopolynoe branchiata, 368 Thimopides. 378 Thunnus thymus, 259 Tilapia mossambica, 83, 84, 85 rendalli, 84 Trachurus. 265 Trachyscorpia cristulata, 381 Transtectonia torrentis, 377 Tricolia pulloides, 24

45 1 Trimusculus reticulatus. 3 1, 33 Trisopterus esmarki, 251 Tubifex tubifex, 409 Typlotanais sp., 377 Umbonium, 31 vestiarium, 3 1, 32 Uria. 268 Uroptychus bicavatus, 378 n. sp., 378 thermalis. 378 (Vesicomya) gigas, 373 (Vesicomya) lepta, 373 Venerupispullastra, 13 Ventiella sulfuris, 377,398 Ventsia tricarinata, 371 Ventuloniaphalcata, 371 Verruca n. sp.1, 375 n. sp.2, 375 Virgularia mirabilis. 223 Xandaros acanthodes, 368 Xylocythere, 384 n. sp., 376 Xylodiscula major, 372 Xylophaga, 384

Subject Index Note: Page references in italics refer to Figures; those in bold refer to Tables Alvinellidae development and dispersal, 41 1 distribution, 410-1 1 gene flow, 41 1-12 phylogenetic relationships, 412-1 3, 413 Ampharetidae development and dispersal, 414 distribution, 414 ATP, 10 Barents Sea ecosystem, 267 Benthic fauna and habitat, 208-36 degradation of reef habitats, 232-3 direct effects of fishing gears, 209-27 drive netting, 225-7 effects of trawls and dredges on epifauna, 219-22, 220 effects of trawls and dredges on infauna, 2 13- 19 explosives, 225-7 indirect effects on habitat, 227-33 mobile epibenthic invertebrate fauna, 288 natural vs. fishing disturbance, 233-5, 234 poisons, 225-7 potential effects of sediment resuspension, 228, 228 static fishing gears, 223-5 Biogeography, 3554,356 approaches, 3 5 6 7 hydrothermal vent fauna, 385-94 Birds, 258 changes in kittiwake and fulmar populations, 285, 285 prey removal, 268-74,269 Bivalves, 418-19 Cenozoic ridge, 388-9

Coastal lagoon fisheries, 73-199 anthropogenic constraints, 134-5 1 anthropogenic variables, 137 biological characteristics, 135 biological variables, 115 change in catch per unit effort according to number of fishermen per unit area, 146 change in fishery yield according to mean depth, 127 according to water area, I26 collected lagoon data, 184-99 correlation between area of immersed vegetation (SC)and Res, 143 correlation between fishery yield and maximum depth, 127 data description and collection, 86-1 16,87-109, 111-12 definitions of lagoon, 77-9 descriptive statistics, 117-19, 134 diversity of fishery data and fishery techniques, 110 effect of brushparks, 83 effect of devices and practices on catch, 82 effect of fish barrages, 83 effect of freshwater run-off on yield, 139 effect of introduction of exotic fish species, 83-5 effect of tide height on yield, 140 environmental constraints, 134-5 1 environmental data, 114-15 fishing effort and catch per unit effort, 144-5 fishing practices, 116 fishing pressure, 150 flushing index (FZ),115 geographical constraints, 117-34 geographical data, 110, 119-21, 121

SUBJECT INDEX

geographical variables, 124 maximum sustainable yield ( M S Y ) , 130, 132, 148 mean and maximum depths, 114 morphoedaphic index (ME& 133, 148 morphometrical constraints, 1 17-34 morphometrical data, 113-14, 121-8 morphometrical variables, 124 multivariate analysis, 128-9 overview, 77-86 physical characteristics and variabilities, 79 physico-chemical characteristics, 135 physico-chemical data, 140-2 physico-chemical variables, 1 15, 136 productivity, 149 significant relationships, 124, 136 statistical analysis, 110, 116-17 statistical approaches, 86 studentized residuals (Res), 117 vs. wetland area, 144 tidal prism (TP), 115, 138 type I-V lagoons, 78-9, 118 under-estimation ( U E ) , 110, 113, 119 water exchange characteristics, 135 data, 135-9 variables, 136 yield, 80-2, 80, 110, 122, 131-3, 138 and anthropogenic characteristics, 135

and latitude, 119 data, 179-83 distribution, 123 per unit area according to tidal prism (TP), 138 per unit area as function of fishing pressure, 145 records, 117, 118 vs. latitude, 125 vs. mean nitrite-nitrogen concentration, 142 vs. minimum salinity, 141 vs. minimum temperature, 141 vs. radiation balance, 125 vs. under-estimation (UE), 123 vs. water volume, 128 year-to-year variations, 120 Copepods, 4 19-20,420 Coral reef ecosystem, 230-1,231 Cousin Reserve, 250

453 Discards, 283-90, 284, 289, 313 Dispersal, 395-6, 401, 401 Dredges, 209-1 3 effects on epifauna, 219-22, 220 effects on infauna, 213-19 Drive netting, 225-7 East Pacific Rise (EPR), 354, 362, 365, 386, 387, 398, 399, 404, 405, 410, 414-16, 420, 422 ECOPATH models, 298 Ecosystem processes, 306-8 modelling, 296-9 Explosives, 225-7 Feeding system, 27-32 mucus in, 27-32, 29, 46-7 scavengers, 288-9 Fish community structure, 236-57 annual size spectra, 247 biomass changes and biomass composition, 243 changes in multispecies communities, 248-50 composition of non-cryptic diurnally active reef-associated fish community in Seychelles’ marine reserves, 250 diversity, 237-46 diversity loss and ecosystem stability, 240-5 indirect diversity losses, 240 intraspecific changes in life histories, 251-4 intraspecific diversity, 245-6 life history traits, 248-56 local extinctions and redundancy, 237-40 mean biomass of piscivorous length and mean species richness, 242 relationships between structure of target fish communities in Fijian fishing grounds, 251 reproduction, 254-6 sex ratios of fish populations, 254-6 size selective fishing, 254-6 size structure of, 2 4 6 7 species extinction, 240-1 temporal changes in estimated abundance of nine species, 244

454 Fish population biology, 258-9 Fisheries. See Coastal lagoon fisheries; Fish community structure Fishing effects on marine ecosystems, 201-351 active fishing techniques, 209-22 degradation of reef habitats, 232-3 drive netting, 225-7 ecosystem management, 305-6 ecosystem processes, 306-8 effect of fishing effort, 308 explosives, 225-7 fish community structure, 236-57 global discards and landings, 208 global estimates of primary production, 204 indirect effects on habitat, 227-33 investigating marine food webs, 295-6 management, 303-1 1 marine reserves, 309-10 modelling ecosystem processes, 296-9 poisons, 225-7 population management, 305-6 potential effects of sediment resuspension, 228, 228 protected areas, 308-10 research, 293 research sites, 299-300 reversibility, 290-1 role of fisheries in marine ecosystems, 303-4 small-scale fishing effects, 301 spatial and temporal scales of study, 300-2 species replacement, 276-83, 277, 278 static fishing gears, 223-5 statistical basis for correlative studies, 284-5 study methods, 292-302 trophic interactions, 306-8 vs. natural disturbance, 233-5, 234 see also Benthic fauna; Trophic interactions Fishing effort in shell-fisheries and mortality or emigration of shorebirds, 274 Fishing intensity and mean biomass of fishes recorded, 241 vs. mean number of grouper species, 238

SUBJECT INDEX

Forage fishes, 257 Galapagos Rift, 354, 365, 385-7, 398, 399,415 Gastropoda development and dispersal, 418 distribution, 415-16 gene flow, 418 patterns of endemism, 416-17 Gene flow, 396-8 Alvinellidae, 41 1-12 Gastropoda, 418 vestimentiferans, 407 Georges Bank, 219, 220,246, 279, 280, 281, 306 Glycoconjugates, 3 Glycoproteins, 3, 5 Glycosaminoglycans(GAGS), 3,5,7, 13, 15 Gulf of Mexico, 384 Gulf of St. Lawrence, 279 Hydrothermal vent fauna Atlantic, 391-2 biogeography, 385-94 causes of diversity, 423-5, 425 composition, 363-79 dispersal, 395-6, 401, 401 distribution, 385-6 distribution patterns of taxa, 402-20 ecological controls, 424-5 extinction, 423 gene flow, 396-8 generic/specific distributions, 402-3, 403,404,405 global species differences, 422-3 gradients, 421-2 high sulphide environments, 379-81 Indian Ocean, 392 local to regional-scale processes, 394-8 Northeast Pacific, 386-7, 388, 389 organic remains, 384-5 overall composition, 381 patterns in diversity, 420-5, 422 phyletic composition, 364-79 pogonophorans, 40410 problems, 357 regional studies, 386-92 regional to global-scale processes, 398-401

SUBJECT INDEX

representation of major invertebrate phyla, 382 reproductive strategies, 395 ridge characteristics, 399 seep/whale/vent relations, 3 9 2 4 seeps, 3824,383 Southern Pacific, 390-1 speciation, 423, 424 species currently known, 364, 365-81 sulphide-rich habitats, 3 9 3 4 taxon listings, 363-4 vagrant taxa, 380-1 vestimentiferans distribution, 406 vicariance, 399-401 Western Pacific, 387-90, 389 Hydrothermal vents, 353-442 definition, 358-9 distribution of major sites, 360 distribution on a ridge, 362 habitat character, 361-3 habitat distribution, 359-61 primary sites of biological collections, 358 ridge-crest venting, 359 Indian Ocean, 392 Juan de Fuca Ridge, 386-7 Lagoons. See Coastal lagoon fisheries Life history traits, 248-56 intraspecific changes, 251-4 Mammals, 258 prey removal, 274-6 Manus Basin, 387 Mariana Trough, 387 Marine food webs, 295-6 Marine reserves, 309-10 Mean beam trawling intensity, 212 Metabolic faecal loss (MFL), 43 Mid-Atlantic Ridge, 399, 405, 416, 421 Middle Atlantic Bight, 228, 228, 255 Mucopolysaccharides, 5 Mucus adhesive nature, 30, 47 biochemical content, 6 composition, 4-8 ecology, 43-50 energy-rich, 26

455 fate of, 43-5 from marine molluscs, 1-71 functions, 2-3, 9-10, 23-35 high viscosity, 10 in biological interactions, 45-50 in distribution of species, 48 in energy budgets, 35-43, 38,41, 46-7 in feeding, 27-32, 29, 4 6 7 in gardening behaviour, 47-8 in locomotion, 23-5, 40 in non-molluscan groups, 3 in protection, 3 2 4 lubricatory, 18 nomenclature, 3 overview, 50-1 production, 9-23, 20, 37, 41, 50 production rates, 19 properties, 8-9 structure, 7 trails, 22-3, 2 S 7 , 25, 45-7 viscosity, 7 Mucus-secreting cell, 13 Multispecies virtual population analysis (MSVPA), 305 Norwegian Sea-Barents Sea ecosystem, 266 Peruvian upwelling ecosystem, 278, 278 Philippine Sea Plate, 389 Phylogenetic diversity, 261 Planktivores, 265 Pogonophorans, 404-10 Poisons, 2 2 5 7 Predation, 258, 280 Predator-prey relationships, 260, 262, 295 Predator removal, 259-65, 260 Prey populations, 264 Prey release, 262 Prey removal, 265-76 birds, 268-74, 269 fishes, 265-8 mammals, 274-6 Prey species, 262 Prince William Sound, 270 Protein-polysaccharide complexes, 5 Proteoglycans, 5 Pull-seining, 225-7

SUBJECT INDEX

Reproduction, 254-6 strategies, 395 Reversibility of fishing effects, 290-1 Ridge characteristics, 399 RISE Project Group, 386 Rivera Fracture Zone, 398

area trawled by English North Sea fleet, 216 effects on epifauna, 219-22, 220 effects on infauna, 213-19 sediment suspension, 228,228 Trophic interactions, 257-92, 306-8

Ste. Anne Reserve, 250 Sandeel populations, 263, 263 Scavengers, 28390 feeding preferences, 288-9 Scope for growth (SFG), 41-2 Seabirds. See Birds Seeps, 382-4, 383, 392-4 Sex ratios of fish populations, 254-6 Shetland Islands, 271 Shorebirds and shell-fisheries interactions, 272 mortality or emigration, 274 see also Birds Species replacement, 276-83, 277, 278

Valliculture sensu lato, 82 Vestimentiferans development and dispersal, 407 distribution, 404-10, 406 gene flow, 407 phylogenetic relationships, 408-10,

Trawls, 209-13, 287

409

Vicariance, 399-401 Virtual population analysis (VPA), 305 Whales, 264, 392-4 Yield. See Coastal lagoon fisheries

Note: Titla of papers have been converted into subjects and a specific article may therefore appear more than once Abyssal and hadal zones, zoogeography, 32, 325 Abyssal macrobenthos, trophic structure, 32, 427 Acetabularia, marine alga, recent advances in research, 14, 123 Algal-invertebrate interactions, 3, 1 Antarctic benthos, 10, 1 Antarctic fishes, comparative physiology, 24, 32 1 Ascidians biology, 9, 1 physiology, 12, 2 Atlantic, Northeast, meiobenthos, 30, 1 Baltic Sea, autrophic and heterotrophic picoplankton, 29, 73 Barnacles, growth, 22, 199 Bathyal zone, biogeography, 32, 389 Benthic marine infaunal studies, development and application of analytical methods, 26, 169 Benthos abyssal macrobenthos, trophic structure, 32, 427 Antarctic, 10, 1 Northeast Atlantic meiobenthos, 30, 1 sampling methods, 2, 171 Biogeography, hydrothermal vent fauna, 34, 353 Blood groups, marine animal, 2, 85 Blue whiting, North Atlantic, population biology, 19, 257 Brachiopods, living, biology, 28, 175 Bryozoans, marine, physiology and ecology, 14, 285 Bullia digitalis, 25, 179 Calanoid copepods, biology of, 33 Cephalopods flotation mechanisms in modern and fossil, 11, 197 recent studies on spawning, embryonic development, and hatching, 25, 85 Chaetognaths, biology, 6, 271 Cladocerans, marine, reproductive biology, 31, 80 Climatic changes, biological response in the sea, 14, 1 Clupeid fish behaviour and physiology, 1, 262 biology, 20, 1 parasites, 24, 263

CUMULATIVE INDEX OF TITLES

Copepods association with marine invertebrates, 16, 1 calanoid, biology of, 33 respiration and feeding, 11, 57 Coral reefs adaptations to physical environmental stress, 31, 222 assessing effects of stress, 22, 1 biology, 1, 209 communities, modification relative to past and present prospective Central American seaways, 19,91 ecology and taxonomy of Halimeda: primary producer of coral reefs, 17, 1 Crustaceans, spermatophores and sperm transfer, 29, 129 Ctenophores, nutritional ecology 15, 249 Die1 vertical migrations of marine fishes: an obligate of facultative process?, 26, 115 Donax serra, 25, 179 Fxhinoids, photosensitivity, 13, 1 Eels, North Atlantic freshwater, breeding, 1, 137 Effluents, effects on marine and estuarine organisms, 3, 63 Environmental simulation experiments upon marine and estuarine animals, 19, 133 Euphausiids, biology, 7, 1, 18, 373 Fish alimentary canal and digestion in teleosts, 13, 109 Antarctic fish,comparative physiology, 24, 321 artificial propagation of marine fish, 2, 1 clupeid behaviour and physiology, 1, 262 clupeid biology, 10, 1 clupeid parasites, 24, 263 die1 vertical migrations, obligate or facultative process?, 26, 115 diseases, 4, 1 egg quality, 26, 71 gustatory system, 13, 53 migration, physiological mechanisms, 13, 241 North Atlantic freshwater eels, 1, 137 nutrition, 10, 383 parasites in deep-sea environment, 11, 121 photoreception and vision, 1, 171 predation on eggs and larvae of marine fishes and the recruitment problem, 25, 1 production and upwelling, 9, 255 year class strength, and plankton production, update of matchmismatch hypothesis, 26, 249 Fish farming, estuarine, 8, 119 Fish larvae appraisal of condition measures for marine fish larvae, 31, 217 field investigations of the early life stages of marine fish, 28, 1 turbulence and feeding ecology, role of microscale, 31, 170 Fish migration, physiological mechanisms in the migration of marine and amphihaline fish, 13, 248 Fisheries, lagoons, constraints on, 34, 73

CUMULATIVE INDEX OF TITLES

459

Fisheries, management of resources, 6, 1 Fisheries and seabird communities, competition, 20, 225 Fishing effects, on marine ecosystems, 34,201 Frontal systems, aspects of biology, 23, 163 Gastropods comparison between Donax serra and Bullia digitalis, bivalve molluscs, 25, 179 intertidal, ecology, 16, 111 marine, burrowing habit, 28, 389 Gonatid squids, subarctic North Pacific, ecology, biogeography, niche diversity and role in ecosystem, 32, 243 Gonionemus, erratic distribution: the occurrence in relation to oyster distribution, 14, 25 1 Habitat selection by aquatic invertebrates, 10, 271 Halibut Hippoglossus hippoglossus, biology, 26, 1 Halimeda, 17, 1 Herring CIupea harengus L. and other clupeids behaviour and physiology, 1, 262 biology, 20, 1 relationships with its parasites, 24, 263 Human affairs, marine biology, 15, 233 Hybridization in the sea, 31, 2 Hydrothermal vent communities, deep sea, ecology, 23, 301 Hydrothermal vent fauna, mid-Atlantic ridge, ecology and biogeography, 32, 93 Hydrothermal vents, biogeography of, 34,353 Indo-West Pacific region, mangrove swamps and forests, fauna and flora, 6, 74 Isopoda, oniscid, biology of the genus Tyros, 30,89 Japan, scallop industry, 20, 309 Japanese oyster culture industry, recent developments, 21, 1 Lagoon fisheries, constraints on, 34,73 Learning by marine invertebrates, 3, 1 Mangrove swamps and forests of Indo-West Pacific region, general account of fauna and flora, 6,74 Marine animals blood groups, 2, 85 neoplasia, 12, 151 Marine ecosystems, effects of fishing on, 34,201 Marine toxins venomous and poisonous animals, 3, 256 venomous and poisonous plants and animals, 21, 59 Meiobenthos of the Deep Northeast Atlantic, 30, 1 Mesoplankton distribution patterns, 32, 9 and macroplankton, some problems of vertical distribution in ocean, 32, 1

460

CUMULATIVE INDEX OF TITLES

Metabolic energy balance in marine invertebrates, influence of temperature on maintenance, 17, 329 Microbiology, marine, present status of some aspects, 2, 133 Molluscs hosts for symbioses, 5, 1 wood-boring teredinids, biology, 9, 336 Molluscs, bivalve effects of environmental stress, 22, 101 and gastropods: comparison between Donax serra and Bullia digitalis, 25, 179 rearing, 1, 1 scatological studies, 8, 307 Molluscs, marine, mucus in, 34, 1 Mucus, in marine molluscs, 34, 1 "N, natural variations in the marine environment, 24,389 Nazca submarine ridge, composition and distribution of fauna, 32, 145 Nitrogen cycle, and phosphorus cycle, plankton, 9, 102 Oil pollution of the seas, problems, 8, 215 Oniscid isopods, biology of the genus Tylos, 30, 89 Oysters, living, speciation, 13, 357 Particulate and organic matter in sea water, 8, 1 Pelagic invertebrates, social aggregation, 30, 155 Penaeida, biology, 27, 1 Petroleum hydrocarbons and related compounds, 15, 289 Phoronida, biology, 19, 1 Phosphorus cycle, and nitrogen cycle, plankton, 9, 102 Pigments of marine invertebrates, 16, 309 Plankton distribution, vertical, in the ocean, 32, 4 laboratory culture of marine holozooplankton and its contribution to studies of planktonic food webs, 16, 21 1 and nitrogen and phosphorus cycles of sea, 9, 102 oceanic phytoplankton, outline of geographical distribution, 32, 527 parasitology of marine zooplankton, 25, 112 phytoplankton, circadian periodicities in natural populations, 12, 326 phytoplankton blooms, harmful or exceptional, 31, 302 picoplankton, Baltic Sea, 29, 73 production, and year class strength in fish populations: update of the matchmismatch hypothesis, 26, 249 Pollution studies with marine plankton Part 1: Petroleum hydrocarbons and related compounds, 15, 289 Part 2: Heavy metals, 15, 381 Pseudocalanus, biology, 15, 1 Pycnogonida, biology, 24,1 Sala y G6mez submarine ridge,composition and distribution of fauna, 32, 145 Salmon, acclimatization experiments, Southern Hemisphere, 17, 398 Scallop industry in Japan, 20, 309

CUMULATIVE INDEX OF TITLES

46 1

Sea anemones, nutrition, 22, 65 Seabird communities, and fisheries, competition, 20, 225 Seamounts, biology, 30,305 Seaweeds of economic importance, aspects of biology, 3, 105 population and community ecology, 23, 1 Shrimps, pelagic, biology, 12, 233 Siphonophore biology, 24, 97 Sole (Solea solea), Bristol Channel, 29, 215 Spermatophores and sperm transfer in marine crustaceans, 29, 129 Squids gonatid squids, subarctic North Pacific, ecology, biogeography, niche diversity and role in ecosystem, 32, 243 oceanic, review of systematics and ecology, 4, 93 Taurine in marine invertebrates, 9, 205 Teredinid molluscs, wood-boring, biology, 9, 336 Tropical marine environment, aspects of stress, 10, 217 Turbulence feeding ecology of larval fish, role of microscale, 31, 170 phytoplankton cell size, and structure of pelagic food webs, 29, 1 5 1 0 s biology, 30,89

Cumulative Index of Authors

Ahmed, J., 13, 357 Akberali, H. B., 22, 102 Allen, J. A., 9, 205 Ansell, A. D., 28, 175 Arakawa., K. Y.,8, 307 Arnaud, F., 24,1 Bailey, K. M., 25, 1 Bailey, R. S., 19, 257 Balakrishnan Nair, M., 9, 336 Bamber, R. N., 24,1 Bett, B. J., 30, 1 Blaxter, J. H. S., 1, 262, 20, 1 Boletzky, S. V., 25, 85 Boney, A. D., 3, 105 Bonotto, S., 14, 123 Bourget, E., 22, 200 Branch, G. M., 17, 329 Brinkhurst, R. O., 26, 169 Brown, A. C., 25, 179,28, 389, 30, 89 Brown, B. E., 22, 1, 31, 222 Bruun, A. F., 1, 137 Burd, B. J., 26, 169 Campbell, J. I., 10, 271 Carroz, J. E., 6, 1 Chapman, A. R. O., 23, 1 Cheng, T. C. 5, 1 Clarke, M. R., 4, 93 Collins, M. J., 28, 175 Corkett, C. J., 15, 1 Comer, E. D. S., 9, 102, 15, 289 Cowey, C. B., 10, 383 Crisp, D. J., 22, 200 Curry,G. B., 28, 175 Cushing, D. H., 9, 255, 14, 1, 26, 249 Cushing, J. E., 2, 85 Dall. W., 27, 1 Davenport, J., 19, 133 Davies, H. C., 1, 1 Davies, M. S., 34, 1 Davis, A. G., 9, 102, 15, 381

Dell, R. K., 10, 1 Denton, E. J., 11, 197 Dickson, R. R., 14, 1 Dinet, A., 30, 1 Dower, J. F., 31, 170 Edwards, C., 14, 251 Egloff, D. A., 31, 80 Emig, C. C., 19, 1 Evans, H. E., 13, 53 Ferrero, T., 30, 1 Ferron, A., 30,217 Fisher, L. R., 7, 1 Fofonoff, P.W., 31, 80 Fontaine, M., 13, 241 Furness, R. W., 20, 225 Galkin, S. V., 32, 93 Gardner, J. P. A., 31, 2 Garrett, M. R., 9, 205 Gebruk, A. V., 32, 93 Ghirardelli, E., 6, 271 Gilpin-Brown, J. B., 11, 197 Glynn, P. W., 19, 91 Gooday, A. J., 30, 1 Goodbody, I., 12, 2 Gotto, R. V., 16, 1 Grassle, J. F., 23, 301 Gulland, J. A., 6, 1 Harris, R. P.,16, 211 Haug, T., 26, 1 Hawkins, S. J., 34, 1 Heath, M. R., 28, 1 Hickling, C. F., 8, 119 Hill, B. J., 27, 1 Hillis-Colinvaux, L.,17, 1 Holliday, F. G. T., 1, 262 Holme, N. A., 2, 171 Holmefjord, I., 26, 71 Horwood, J., 29, 215 Houde, E. D., 25, 1 Howard, L. S., 22, 1

CUMULATIVE INDEX OF AUTHORS

Hunter, J. R., 20, 1 James, M. A. 28, 175 Jennings, S., 34,201 Joyeux, J.-C., 34,73 Kaiser, M. J., 34,201 Kapoor, B. G., 13, 53, 13, 109 Kennedy, G. Y., 16, 309 Kierboe, T., 29, 1 Kjersvik, E., 26, 71 Kuosa, H., 29, 73 Kuparinen, J., 29, 73 Lambshead, P. J. D., 30, 1 Le Fevre, J., 23, 163 Leggett, W. C., 30, 217, 31, 170 Loosanoff, V. L., 1, 1 Lurquin, P., 14, 123 McArthur, A. G., 34,353 Macdonald, J. A., 24, 321 McHugh, D., 34,353 Mackenzie, K., 24, 263 Mackie, G. O., 24, 97 McLaren, I. A,, 15, 1 Macnae. W., 6, 74 Mangor-Jensen, A., 26, 71 Marshall, S. M., 11, 57 Mauchline, J., 7, 1, 18, 1, 33, 1-660 Mawdesley-Thomas, L. E., 12, 151 Mazza, A., 14, 123 Meadows, P. S., 10, 271 Millar, R. H., 9, 1 Miller, T. J., 31, 170 Millot, N., 13, 1 Mironov, A. N., 32, 144 Montgomery, J. C., 24, 321 Moore, H. B., 10, 217 Moskalev, L. I., 32, 93 Naylor, E., 3, 63 Neilson, J. D., 26, 115 Nelson-Smith, A., 8, 215 Nemec, A., 26, 169 Nesis, K. N., 32, 144, 32, 243 Newell, R. C., 17, 329 Nicol, J. A. C., 1, 171 Noble, E. R., 11, 121 Odendaal, F. J., 30,89 Omori, M., 12,233 Onbi, T., 31, 80 Owens, N. J. P., 24, 389 Paffenhofer, G. A., 16,211 Parin, N. V., 32, 144 Peck, L. S., 28, 175

463 Perry, R. I., 26, 115 Pevzner, R. A., 13, 53 Pfannkuche, O., 30, 1 Pugh. P. R. 24,97 Purcell, J. E., 24, 97 Reeve, M. R., 15, 249 Rhodes, M. C., 28, 175 Richardson, K., 31, 302 Riley, G. A., 8, 1 Ritz, D. A., 30, 155 Rogers, A. D., 30,305 Rothlisberg, P. C., 27, 1 Russell, F. E., 3, 256, 21, 60 Russell, F. S., 15, 233 Ryland, J. S., 14, 285 Saraswathy, M.,9, 336 Sargent, J. R., 10, 383 Scholes, R. B., 2, 133 Semina, H. J., 32, 527 Shelbourne, J. E., 2, 1 Shewan, J. M., 2, 133 Sindermann, C. J., 4, 1 Smit, H., 13, 109 Sokolova, M. N., 32, 427 Soltwedel, T., 30, 1 Sournia, A., 12, 326 Southward, A. J., 32, 93 Staples, D. J., 27, 1 Stenton-Dozey, J. M. E., 25, 179 Stewart, L., 17, 397 Subramoniam, T., 29, 129 Taylor, D. L., 11, 1 Thiodorides, J., 25, 117 Trueman, E. R., 22, 102,25, 179,28,389 Tunnicliffe, V., 34,353 Underwood, A. J., 16, 111 Van-Praet, M., 22, 66 Vanreusel, A., 30, 1 Ventilla, R. F., 20, 309, 21, 1 Vereshchaka, A. L., 32, 93 Verighina, I. A., 13, 109 Vincx, M., 30, 1 Vinogradov, M. E., 32, 1 Vinogradova, N. G., 32, 325 Walters, M. A., 15, 249 Ward, A. B., 34,73 Wells, M. J., 3, 1 Wells, R. M. G., 24,321 Yonge, C. M., 1, 209 Zezina, 0. N., 32, 389

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