Oceanography and Marine Biology, An Annual Review, Volume 41: An Annual Review: Volume 41 (Oceanography and Marine Biology)

OCEANOGRAPHY and MARINE BIOLOGY AN ANNUAL REVIEW Volume 41 OCEANOGRAPHY and MARINE BIOLOGY AN ANNUAL REVIEW Volume ...

Author: R. N. Gibson | R. J. A. Atkinson

80 downloads 1265 Views 6MB Size Report

This content was uploaded by our users and we assume good faith they have the permission to share this book. If you own the copyright to this book and it is wrongfully on our website, we offer a simple DMCA procedure to remove your content from our site. Start by pressing the button below!

Report copyright / DMCA form

DOWNLOAD PDF

OCEANOGRAPHY and MARINE BIOLOGY AN ANNUAL REVIEW

Volume 41

OCEANOGRAPHY and MARINE BIOLOGY AN ANNUAL REVIEW

Volume 41 Editors R. N. Gibson The Dunstaffnage Marine Laboratory Oban, Argyll, Scotland [email protected]

R. J. A. Atkinson University Marine Biological Station Millport, Isle of Cumbrae, Scotland [email protected] Founded by Harold Barnes

First published 2003 by Taylor & Francis 11 New Fetter Lane, London EC4P 4EE Simultaneously published in the USA and Canada by Taylor & Francis Inc 29 West 35th Street, New York, NY 10001 Taylor & Francis is an imprint of the Taylor & Francis Group This edition published in the Taylor & Francis e-Library, 2004. © 2003 R. N. Gibson and R. J. A. Atkinson All rights reserved. No part of this book may be reprinted or reproduced or utilised in any form or by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying and recording, or in any information storage or retrieval system, without permission in writing from the publishers. Every effort has been made to ensure that the advice and information in this book is true and accurate at the time of going to press. However, neither the publisher nor the authors can accept any legal responsibility or liability for any errors or omissions that may be made. In the case of drug administration, any medical procedure or the use of technical equipment mentioned within this book, you are strongly advised to consult the manufacturer’s guidelines. British Library Cataloguing in Publication Data A catalogue record for this book is available from the British Library Library of Congress Cataloging in Publication Data A catalog record for this book has been requested ISBN 0-203-18057-7 Master e-book ISBN

ISBN 0-203-23081-7 (Adobe eReader Format) ISBN 0-415-25463-9 (Print Edition)

CONTENTS

Preface

vii

Oxygen minimum zone benthos: adaptation and community response to hypoxia

1

Lisa A. Levin

Antarctic marine benthic diversity

47

Andrew Clarke & Nadine M. Johnston

Inﬂuence of marine allochthonous input on sandy beach communities

115

I. Colombini & L. Chelazzi

The effects of sedimentation on rocky coast assemblages

161

Laura Airoldi

Exotic molluscs in the Mediterranean basin: current status and perspectives

237

Serge Gofas & Argyro Zenetos

Detritus in the epilithic algal matrix and its use by coral reef ﬁshes

279

Shaun K. Wilson, David R. Bellwood, J. Howard Choat & Miles J. Furnas

Ecology of whale falls at the deep-sea ﬂoor

311

Craig R. Smith & Amy R. Baco

The diet of harbour porpoise (Phocoena phocoena) in the northeast Atlantic: a review

355

M. B. Santos & G. J. Pierce Author index

391

Systematic index

419

Subject index

431

v

PREFACE

The forty-ﬁrst volume of this series contains eight reviews written by an international array of authors that, as usual, range widely in subject and taxonomic and geographic coverage. The majority of articles were solicited but the editors always welcome suggestions from potential authors for topics they consider could form the basis of appropriate contributions. Because an annual publication schedule necessarily places constraints on the timetable for submission, evaluation and acceptance of manuscripts, potential contributors are advised to make contact with the editors at an early stage of preparation so that the delay between submission and publication is minimised. In line with recent developments in electronic publishing, it is a pleasure to announce that the series will now be available in eBook format. Details can be found at the publisher’s eBookstore at http://www.ebookstore.tandf.co.uk The editors again gratefully acknowledge the willingness and speed with which authors complied with the editors’ suggestions, requests and questions and the efﬁciency of the copy editor and publishers in ensuring the regular annual appearance of each volume.

vii

Oceanography and Marine Biology: an Annual Review 2003, 41, 1–45 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

OXYGEN MINIMUM ZONE BENTHOS: ADAPTATION AND COMMUNITY RESPONSE TO HYPOXIA LISA A. LEVIN Integrative Oceanography Division, Scripps Institution of Oceanography, La Jolla, CA 92093-0218 USA e-mail: [email protected] Abstract Mid-water oxygen minima (0.5 ml l1 dissolved O2) intercept the continental margins along much of the eastern Paciﬁc Ocean, off west Africa and in the Arabian Sea and Bay of Bengal, creating extensive stretches of sea ﬂoor exposed to permanent, severe oxygen depletion. These seaﬂoor oxygen minimum zones (OMZs) typically occur at bathyal depths between 200 m and 1000 m, and are major sites of carbon burial along the continental margins. Despite extreme oxygen depletion, protozoan and metazoan assemblages thrive in these environments. Metazoan adaptations include small, thin bodies, enhanced respiratory surface area, blood pigments such as haemoglobin, biogenic structure formation for stability in soupy sediments, an increased number of pyruvate oxidoreductases, and the presence of sulphide-oxidising symbionts. The organic-rich sediments of these regions often support mats of large sulphideoxidising bacteria (Thioploca, Beggiatoa, Thiomargarita), and high-density, low-diversity metazoan assemblages. Densities of protistan and metazoan meiofauna are typically elevated in OMZs, probably due to high tolerance of hypoxia, an abundant food supply, and release from predation. Macrofauna and megafauna often exhibit dense aggregations at OMZ edges, but depressed densities and low diversity in the OMZ core, where oxygen concentration is lowest. Taxa most tolerant of severe oxygen depletion (0.2 ml l1) in seaﬂoor OMZs include calcareous foraminiferans, nematodes, and annelids. Agglutinated protozoans, harpacticoid copepods, and calciﬁed invertebrates are typically less tolerant. High dominance and relatively low species richness are exhibited by foraminiferans, metazoan meiofauna, and macrofauna within OMZs. At dissolved oxygen concentrations below 0.15 ml l1, bioturbation is reduced, the mixed layer is shallow, and chemosynthesis-based nutrition (via heterotrophy and symbiosis) becomes important. OMZs represent a major oceanographic boundary for many species. As they expand and contract over geological time, OMZs may inﬂuence genetic diversity and play a key role in the evolution of species at bathyal depths. These ecosystems may preview the types of adaptations, species, and processes that will prevail with increasing hypoxia over ecological and evolutionary time. However, many questions remain unanswered concerning controls on faunal standing stocks in OMZs, and the physiological, enzymatic, metabolic, reproductive and molecular adaptations that permit benthic animals to live in OMZs. As global warming and eutrophication reduce oxygenation of the world ocean, there is a pressing need to understand the functional consequences of oxygen depletion in marine ecosystems.

1

LISA A. LEVIN

Introduction Oxygen depletion is widespread in the world oceans (Kamykowski & Zentara 1990), occurring as permanent, seasonal and episodic features. Persistent low oxygen is evident in midwater oxygen minimum zones (OMZs), deﬁned as regions where oxygen concentrations are 0.5 ml l1 (or about 7.5% saturation; 22 M). These features are sometimes called oxygen minimum layers or oxygen-deﬁcient zones, and are present at different water depths ranging from shelf to upper bathyal zones (10–1300 m). Oxygen minima are created through biochemical oxygen consumption, with circulation affecting their distribution and position within the water column (Wyrtki 1962). Where oxygen minima intercept the continental margin or seamounts, they have large effects on benthic assemblages. The present review represents the ﬁrst detailed synthesis of benthic responses to permanent oxygen depletion in OMZs, covering shelf and bathyal depths. The goal is to describe the general features of the seaﬂoor environments that are intercepted by OMZs and to review what is known about the effects of OMZs on benthic organisms and biotic processes. While the structure and composition of selected OMZ communities have been described in several regions of the world, this review represents the ﬁrst attempt to synthesise OMZ patterns and trends across geographic locations, different taxonomic groups (protozoans, metazoan meiofauna, macrofauna, and megafauna) and different levels of biotic organisation (organisms, communities, and ecosystems). Although a review of the literature reveals signiﬁcant gaps in our knowledge of OMZ physiology, population dynamics and ecosystem function, recent observations suggest that OMZs are an important frontier for discovery of new adaptations and processes at many levels. A range of terms has been used to describe different dissolved oxygen concentrations. This paper will adopt the following nomenclature. Anoxia refers to the complete absence of dissolved oxygen. The term microxic will be used to describe oxygen when it is measurable up to concentrations of 0.1 ml l1, following Bernhard & Sen Gupta (1999). Dysoxic or dysaerobic refers to oxygen concentrations of 0.1 ml l1 to 1.0 ml l1, and oxic (aerobic) waters contain 1 ml l1 O2 (Rhoads & Morse 1971). The term hypoxic generally refers to low oxygen conditions that are physiologically stressful. This will vary among taxa, though Kamykowski & Zentara (1990) deﬁne hypoxia as 0.2 ml l1. OMZ waters are those with 0.5 ml l1 O2. The discussion below provides an overview of knowledge about benthos in regions of the sea ﬂoor that are intercepted by permanent, open-ocean OMZs. It does not address benthic responses to oxygen depletion in very shallow water, enclosed seas, fjords, silled basins that occur outside OMZs, or in seasonally hypoxic waters. Comprehensive recent reviews of benthic response to shallow-water hypoxia associated with eutrophication can be found in Diaz & Rosenberg (1995, 2001), Gray et al. (2002), and Karlson et al. (2002). The many facets of seasonal hypoxia in the Gulf of Mexico are presented in Rabalais & Turner (2001). A number of reviews have considered themes related to the OMZ topics addressed here. Thiel (1978) reviewed benthos in upwelling regions, but a considerable amount of work has taken place in the past 25 yr. Rogers (2000) and Levin et al. (2001) reviewed aspects of OMZs related to diversity pattern and generation. Bernhard & Sen Gupta (1999) reviewed adaptations, morphology and assemblage characteristics of Foraminifera in oxygen-depleted waters.

2

OXYGEN MINIMUM ZONE BENTHOS

The physical and geological nature of oxygen minimum zones Distribution, formation, and temporal stability of OMZs OMZs generally form where strong upwelling leads to high surface productivity that sinks and degrades, depleting oxygen within the water column. However, OMZ formation also requires stagnant circulation, long residence times (the absence of oxygen exchange), and the presence of oxygen-depleted source waters (Sarmiento et al. 1988). In the absence of exchange by circulation, oxygen is supplied to the OMZ by vertical and horizontal diffusion and by water ascending from below (Wyrtki 1962). Often OMZs support bacterial denitriﬁcation in which nitrate ions are used for oxidation of organic matter; in the process they are reduced to molecular nitrogen with nitrite as an intermediate (Codispoti & Christiansen 1989). Nitriﬁcation, the oxidation of nitrite and ammonium also occurs in these waters (Ward et al. 1989). The largest OMZs reside at bathyal depths in the eastern Paciﬁc Ocean, in the Arabian Sea, in the Bay of Bengal and off southwest Africa (Fig. 1) (Kamykowski & Zentara 1990). The Baltic Sea, Black Sea, Gulf of Aden, Philippine region, northwest Paciﬁc margin and Norwegian fjords exhibit hypoxia (0.2 ml l1) irrespective of sampling depth. Deep-water hypoxia is found also in some basins, for example in Baja California, in the southern California borderland, in Saanich Inlet and in some fjords (Dean et al. 1994, Diaz & Rosenberg 1995). The very extensive OMZ development in the eastern Paciﬁc Ocean (Fig. 1) can be attributed to the fact that intermediate depth waters of this region are older and have overall lower oxygen concentration than other water masses (Wyrtki 1966).

Figure 1 Distribution of the world oxygen minimum zones. Open water oxygen minima are shown in black, hypoxic enclosed seas and fjords are stippled. (Adapted from Diaz & Rosenberg 1995.)

3

LISA A. LEVIN

All OMZs exhibit a similar general oxygen proﬁle but the oxygen levels, OMZ thickness and depth of occurrence vary regionally (Fig. 2). The upper boundary of the OMZ (0.5 ml l1) may come to within 10 m or 50 m of the sea surface off Central America and Peru (Wyrtki 1973), but may occur as deep as 500 m or 600 m off California or Oregon. Typically, a vertical proﬁle of dissolved oxygen concentration through an OMZ exhibits a steep drop in oxygen from the surface to the upper boundary. Below this there is a zone of continuous low oxygen. The lower OMZ boundary exhibits a more gradual increase in oxygen with water depth (Fig. 2). The shape of the oxygen proﬁle is due to an exponential decrease in oxygen consumption with depth (Wyrtki 1962). The thickness of the OMZ is strongly inﬂuenced by circulation and by the oxygen content of the ocean region. Off Mexico and in the Arabian Sea, the OMZ is over 1000 m thick (Wyrtki 1973, Wishner et al. 1990), but off Chile, the OMZ is 400 m thick (Wyrtki 1966) (Fig. 2). OMZ thickness increases in the north Paciﬁc because the water masses are older and have lower oxygen content than in the south Paciﬁc (Wyrtki 1966). Along continental margins, minimum oxygen concentrations typically occur between 200 m and 700 m. Oxygen concentrations may approach zero, accompanied by denitriﬁcation, but sulphate reduction and the production of hydrogen sulphide rarely occur in the water column (Morrison et al. 1999). A second, deeper oxygen minimum occurs in all the southern hemisphere oceans and in the north

Figure 2 Water column dissolved oxygen proﬁles through OMZs in different regions of the world ocean.

4

OXYGEN MINIMUM ZONE BENTHOS

Paciﬁc (Anderson 1972) but concentrations are typically not low enough to be considered part of the OMZ. The surface area of the sea ﬂoor intercepted by oxygen minimum zones is substantial on a global basis. Helly & Levin (unpubl. obs.) have used seaﬂoor topographic data and US National Oceanographic Data Center oxygen data to estimate that the ocean contains over 106 km2 of sea ﬂoor overlain with bottom water 0.5 ml l1 O2. OMZs differ from many shallow-water dysoxic regions in exhibiting stable, persistent low oxygen over ecological and geological timescales, such that sessile species will live out many generations in continuous low oxygen. However, at the upper OMZ boundary, oxygen concentrations can vary with internal tides (Levin et al. 1991a), seasons (Kamykowski & Zentara 1990), ENSO-associated oceanographic forcing (Tarazona et al. 1988a,b, Gutiérrez et al. 2000), or regime shifts (Stott et al. 2000). An annual cycle of oxygen depletion and replenishment due to seasonal ﬂushing has been proposed to be the primary factor regulating the formation of sedimentary varves in the Santa Barbara Basin (Reimers et al. 1990). Seasonality of diagenetic mobilisation, bacterial mat development and benthic foraminiferal activity are all implicated. Multidecadal increases in water temperature accompanied by a reduction in upwelling may reduce carbon oxidation rates with consequent elevation of dissolved oxygen concentrations (Stott et al. 2000). OMZ intensity and distribution also vary signiﬁcantly on geological timescales. Shifts in productivity or circulation over a few thousands to 10 000 yr (Dansgaard-Oeschger cycles) are thought to drive expansion and contractions of OMZs both vertically and horizontally (Tyson & Pearson 1991, von Rad et al. 1995, Rogers 2000). In the Arabian Sea, signiﬁcant variations in the intensity of the OMZ, related to upwelling intensity and thermocline ventilation, are revealed by studies of benthic Foraminifera, carbonate dissolution, nitrogen isotopes and Cd : Ca ratios (Reichart et al. 1998). Occurrence of the lowest oxygen levels correlates with productivity maxima and shallow winter mixing (Reichart et al. 1998). Millennial-scale ﬂuctuations in the strength of the California margin OMZ, as revealed by shifts in benthic foraminiferal assemblages, are associated with warming and cooling periods (Cannariato et al. 1999, Cannariato & Kennett 1999). Although the oxygen levels throughout most of the California OMZ are not sufﬁcient to prevent bioturbation, annual varve (laminae) preservation occurred during the upper Pleistocene, indicating much lower oxygen levels in the past, possibly due to weakening of ﬂow of Paciﬁc Intermediate Water (Dean et al. 1994). Similar scenarios are described off Peru (Glenn et al. 1993). These changes are not always slow. A shift from varved to bioturbated sediments has been observed within silled basins of the eastern Paciﬁc OMZ in the past few decades (Stott et al. 2000). Organic carbon preservation, laminations, and shifts in C isotopes (thought to reﬂect organic carbon burial rate) have been used to document global-scale anoxic events in the Permo-Triassic, Toaracian, Lower Aptian, Lower and Upper Albian and Cenomanian/ Turonian periods (Jacobs & Lindberg 1998 and references therein). These events largely affected the outer shelves, slopes and basin habitats, but encompassed deeper portions of the Atlantic and the western Paciﬁc at times. Periods of transgression and development of greenhouse conditions in the Mesozoic generated anoxic/dysoxic bottom conditions (Jacobs & Lindberg 1998).

5

LISA A. LEVIN

OMZ sediments and the oceanographic environment The ﬂux of organic matter (OM) sinking to the bottom through oxygen deﬁcient water columns appears to be greater than in oxic waters. Reduced attenuation coefﬁcients for sinking organic matter have been noted for sediment traps deployed beneath OMZs off Peru (Martin et al. 1987), in the Arabian Sea (Haake et al. 1993), and off Mexico (Devol & Hartnett 2001). These are attributed to a decreased oxidation rate of material within the OMZ. Molecular characterisation of sediments underlying the Oman margin OMZ revealed little evidence of zooplankton reworking within the OMZ, but extensive benthic invertebrate reworking at the lower OMZ boundary (Smallwood & Wolff 2000). Thus, the sea bed beneath OMZs receives high inputs of organic matter from the productive overlying waters. Once organic matter reaches the sea ﬂoor, degradation may be reduced in low oxygen conditions. Microbial breakdown is thought to be less efﬁcient when carried out via anaerobic metabolism (Fenchel & Finlay 1995) and anoxia can disrupt the microbial loop by altering bacterial grazing (Lee 1992, Kemp 1990). Experiments by Harvey et al. (1986) indicate that the rate of bacterial degradation of lipids is slower in anoxic, organic-rich sediments (coastal) and that this rate decreases with increasing sediment organic carbon content. Reduced faunal consumption and bioturbation may also slow decomposition (Bianchi et al. 2000). Where oxygen minimum zones impinge on the sea ﬂoor, they create strong gradients in bottom-water oxygen concentration that affect a myriad of habitat properties. As a result of slow decomposition in the overlying water column and on the sea bed, sediment organic carbon and organic nitrogen contents are often very high in OMZ sediments. Phytodetritus layers have been observed on the sea bed beneath OMZ settings in the eastern Paciﬁc and Arabian Sea (Pfannkuche et al. 2000, Beaulieu 2002). There is also a high ratio of organic carbon to mineral surface area associated with low bottom-water oxygen, indicating enhanced preservation of organic matter under OMZ conditions (Kiel & Cowie 1999). Particulate organic carbon (POC) values of 3–6% are typical in many OMZs (Levin & Gage 1998, Cowie et al. 1999), but off Peru they can reach 15–20% (Rosenberg et al. 1983, Neira et al. 2001b). This represents an exceptionally large potential food supply for deposit feeders. Oxygen minima enhance phosphorite deposition by preserving organic sediments and maintaining high levels of dissolved phosphate in porewaters (Manheim et al. 1975). Phosphorites are accumulating beneath OMZs off Peru and southwest Africa (Piper & Codispoti 1975). The enhanced preservation of organic matter in OMZs also yields high hydrocarbon generation potential, and may indicate why upper slope settings often support large reservoirs of petroleum (Demaison & Moore 1980, Paropkari et al. 1993). There is often a rough spatial correspondence between mid slope sedimentary organic C maxima and bottom-water oxygen minima, although Calvert et al. (1992) report no relationship between total organic C contents of sediments and bottom-water oxygen concentrations in the Gulf of California. Considerable debate has ensued about whether factors such as OM supply, sediment texture, dilution and local winnowing may have greater inﬂuence on organic matter accumulation and burial beneath OMZs (Calvert 1987, Pederson et al. 1992). There is generally an inverse relationship between bottom-water oxygen concentration and sediment POC in bathyal sediments (Levin & Gage 1998). Hydrogen sulphide builds up in some OMZ sediments, but often the sulphide is oxidised with iron or is utilised by sulphideoxidising bacterial mats and H2S does not reach high levels.

6

OXYGEN MINIMUM ZONE BENTHOS

Historical aspects of OMZ exploration Although the Challenger expedition documented reduced oxygen in what are now recognised as OMZs (Ditmar 1884), detailed exploration of OMZ biology is relatively new. The Meteor Expedition reported collecting bottom samples with high sulphur content from the shelf near Walvis Bay and attributed these to decomposing wastes of whale processing (Spiess 1928). The discovery of depleted faunas in the Arabian Sea OMZ occurred aboard the RV MABAHISS during the Murray Expedition of 1933–4, but these results were not published and a full understanding of the oxygen minimum layer in this region did not occur until the International Indian Ocean Expedition of 1959–65 (Gage et al. 2000). This was the same period in which Wyrtki published details of anoxia in the world ocean (Wyrtki 1962) and in the eastern Paciﬁc (Wyrtki 1966). Serious deep-water biological studies of OMZ effects on benthic communities began with Sanders’ (1969) landmark transect off Walvis Bay. The ﬁrst benthic studies within the eastern Paciﬁc OMZ took place in the 1960s off northern Chile (Gallardo 1963) and Peru (Frankenberg & Menzies 1968, Rowe 1971).

Biotic response to OMZs Large sulphur bacteria Large, ﬁlamentous, sulphur bacteria in the genera Thioploca (T. chileae and T. araucae) and Beggiatoa (Jorgensen & Gallardo 1999) are often conspicuous features on the surface of OMZ sediments. Thioploca and Beggiatoa spp. typically thrive in dysoxic conditions where waters are rich in nitrate, because they can store and use nitrate as an electron acceptor for sulphide oxidation (Fossing et al. 1995, McHatton et al. 1996). Thioploca can glide in their mucous sheaths, upward into the water to collect nitrate, and down into sediment (to 15 cm) to access sulphide, which is generated by sulphate-reducing bacteria. Dense mats of Thioploca and Beggiatoa spp. have been reported from the Peru–Chile margin (Gallardo 1977) (Fig. 3), the Pakistan margin (Schmaljohann et al. 2001) and the Santa Barbara Basin (Bernhard et al. 2000). Biomass of Thioploca can reach 120 g ww m2 on the mid Chilean shelf (Schulz et al. 1996), comparable to that of the total benthic fauna. Mat biomass varies seasonally, with highest values off Chile in summer when oxygen depletion is greatest, and inter-annually, with reductions during better-oxygenated El Niño years. Tuft formation and thinner, grass-like cover have been observed off northern Chile (L. Levin, unpubl. obs.), Namibia (Gallardo et al. 1998), and the Oman margin (Levin et al. 1997). Inhabited sheaths of Thioploca may be covered by ﬁlamentous sulphate-reducing bacteria in the genus Desulfonema, aiding in the tight recycling of H2S within the mats (Jorgensen & Gallardo 1999). OMZ sediments characteristically support unusually large bacteria. Within the microxic sediments of the Namibian shelf (0.1 ml l1) there are dense populations of a giant spherical sulphur bacterium (Thiomargarita nammibiensis). Cells typically have diameters of 100 m to 300 m, but may attain 750 m, making them the largest known bacteria, with biomass of up to 47 g m2 (Schulz et al. 1999). Although there has been much interest in mat- and chain-forming bacteria within OMZs, the microbiology of other forms in OMZ sediments remains relatively unexplored. Archaea are known to thrive in anoxic conditions 7

LISA A. LEVIN

Figure 3 (a) Mat of Thioploca sp. from the northern Chilean margin (200 m), sampled in March 2000. (b) Thioploca sp. from the Oman margin, 400 m. Note multiple ﬁlaments (39–40 m diameter) in a sheath. Sheath diameter is ⬃200 m.

(Hinrichs et al. 1999, 2000) and could be well represented in OMZ sediments, given the apparent abundance of marine non-thermophilic Archaea in sediments during oceanic anoxic events in the mid-Cretaceous (Kuypers et al. 2001). Filamentous sulphur bacteria mats provide habitat for an abundant protozoan and metazoan meiofauna in the Santa Barbara Basin (Bernhard et al. 2000). Heavy recruitment of the squat lobster Pleuroncodes monodon during the period when Thioploca mats were best developed on the central Chile shelf led Gallardo et al. (1994, 1995, 1996) to suggest that the mats provide a refuge and a food source for new recruits. Several authors have predicted that Thioploca mats represent a signiﬁcant source of food for the OMZ faunas of the Peru–Chile margin (Gallardo 1977, Arntz et al. 1991).

Adaptations to permanent hypoxia Adaptations to low oxygen are relatively well studied in benthic Foraminifera and have been reviewed by Bernhard & Sen Gupta (1999). Among calcareous Foraminifera, test size appears to be smaller under dysaerobic conditions, tests are more porous, and species are sometimes thinner-walled than in oxic sediments (Bradshaw 1961, Phleger & Soutar 1973, Perez-Cruz & Machain-Castillo 1990, Sen Gupta & Machain-Castillo 1993, Gooday et al. 8

OXYGEN MINIMUM ZONE BENTHOS

2000). Small size of Foraminifera in the oxygen minimum zone can facilitate respiration by enhancing surface area to volume ratios. Reduced size may also be related to high reproductive rates (and abundant young individuals) associated with plentiful food (Phleger & Soutar 1973). It has been proposed that test pores in hyaline species may facilitate oxygen exchange, as mitochondria are more abundant near pores in species from low-oxygen environments (Leutenegger & Hansen 1979). However, this idea is not supported by more recent observations (Bernhard & Alve 1996, Bernhard 1996). Taxa lacking these pores (miliolids) are largely absent in low oxygen environments (Gooday et al. 2000). Encystment appears to be one possible response to temporary hypoxia (Linke & Lutze 1993), although a number of species can tolerate anoxic, sulphidic conditions for short periods. Other possible foraminiferal adaptations may include survival of anoxia without oxidative phosphorylation or extension of pseudopodia containing mitochondria across steep oxygen gradients in the sediment (Travis & Bowser 1986). The presence of bacterial symbionts in some of the most abundant species in the nearly anoxic Santa Barbara Basin has prompted Bernhard & Sen Gupta (1999) to propose a H2S detoxiﬁcation function for the symbionts. Sequestration of symbiotic chloroplasts with the potential for oxygen production may represent yet another adaptation to hypoxia in selected Foraminifera from this environment but the chloroplast function remains uncertain (Bernhard & Bowser 1999), since these occur in species living below the photic zone (Cedhagen 1991). Adaptations of shelf and bathyal benthic metazoans to permanent hypoxia have not been summarised previously. A review for planktonic organisms inhabiting OMZs by Childress & Siebel (1998) provides a useful framework for this discussion. These authors emphasise that animals living in OMZs must adapt to limited oxygen availability, not to a complete absence of oxygen. Even at very low oxygen concentrations, there is sufﬁcient oxygen available in the water if organisms can access it; it is the reduced PO2 gradient driving diffusion from the animal exterior to the mitochondria that poses the main problem. Childress & Siebel (1998) proposed that pelagic OMZ species cope with low oxygen by (a) increasing effectiveness of oxygen uptake, (b) lowering metabolic demands, or (c) utilising anaerobic metabolism. They argue that the ﬁrst of these is the most widely encountered approach. While OMZ organisms show lower metabolic oxygen requirements than shallow-water relatives, other deep-water species not living in OMZs do so as well. Anaerobic respiration appears to be used mainly by vertically migrating plankton that can pay back oxygen debts incurred during daily migrations to better-oxygenated water. In general, all of these possible adaptations are little studied in benthic species. However, there is evidence that OMZ benthos maximise oxygen uptake through morphological and physiological adaptation. The effectiveness of oxygen uptake may be increased by raised ventilation rates, increased efﬁciency of O2 removal from the bloodstream, elevated circulation capacity, increased gill surface area, reduced blood to water diffusion distances, and increased blood pigment afﬁnity for oxygen (Childress & Siebel 1998). High gill surface area is evident in many OMZ taxa including amepliscid amphipods, a group that occurs in OMZs off Oman, Chile, Peru and California (Fig. 4A). Elongate, proliferated and numerous branchiae appear to be adaptations to permanent hypoxia in spionid, dorvilleid, and lumbrinerid polychaetes in OMZ sediments (Fig. 4B-D) (Lamont & Gage 2000). Cossurid polychaetes within the Oman margin OMZ have exceptionally long median antennae that are thought to aid respiration (Lamont & Gage 2000); this taxon is well represented in many bathyal OMZ settings. An epsilonematid nematode endemic to microxic sediments within the Peru margin OMZ, Glochinema bathyperuvensis, is covered with dense, hair-like body spines and cuticular 9

LISA A. LEVIN

Figure 4 Examples of enhanced respiratory surface area in OMZ invertebrates. (A) Gill structures of Ampelisca sp. from the Peru margin (562 m). (B) Branchiae on Ninoe sp. (Polychaeta: Lumbrineridae) from Magdalena Bay, Mexico (700 m). (C) Branchiae on Diaphorosoma sp. (Polychaeta: Dorvilleidae) from the northern Chile margin (313 m). (D) Palps and branchiae of Prionospio (Minuspio) sp. (Polychaeta: Spionidae) from the Oman margin (404 m) Photography by P. Lamont, Dunstaffnage Marine Laboratory. (E) Glochinema bathyperuvensis (Nematoda: Epsilonematidae), Peru margin (305 m). (F) Posterior section showing elongate setae from G. bathyperuvensis specimen shown in E. Photographs E and F by C. Neira, Oldenburg University.

protrusions and a large number of modiﬁed somatic setae that may aid oxygen uptake (Fig. 4E,F) (Neira et al. 2001a). Increased gill surface has been documented in mid-water mysids, ﬁshes, and cephalopods from OMZs (Childress & Siebel 1998). However, Young & Vazquez (1997) noted reduced folding and branchial surface area in the ascidian Styela 10

OXYGEN MINIMUM ZONE BENTHOS

gagetyleri from the core of the Oman margin OMZ. They suggest greater importance of branchial complexity for feeding than for gas exchange. The small size of this ascidian (7–11 mm), combined with reduced branchial sac complexity, may lower respiratory requirements for oxygen. Reduced diffusion distances may explain the success of small, thin, elongate taxa such as oligochaetes and nematodes, which thrive at microxic oxygen levels within OMZs (Levin et al. 2002). Development of respiratory pigments (i.e. haemoglobins or haemocyanins) with high afﬁnity for oxygen has been observed in benthic ﬁsh (Sebastolobus alascanus), the bathypelagic mysid Gnathophausia ingens (Sanders & Childress 1990), and pelagic ﬁshes and crustaceans that live in the OMZ (Childress 1975). Haemoglobin is present in a number of bivalves from the Oman margin OMZ (150–1150 m), including the mytilid mussel Amygdalum anoxicolum (Oliver 2001), Pitar sewelli (Veneroidea), Indocrassatella indica (Crassatelloidea), Lucinoma sp. (Lucinoidea), and Propeamussium cf. alcocki (Pectinoidea) (G. Oliver, pers. comm.). The occurrence of haemoglobin has no phyletic association in these groups, as it does not occur in most members of these superfamilies. Thus haemoglobin appears to be a local adaptation to OMZ conditions (Oliver 2001, pers. comm.). Maintenance of low metabolic rate through limited activity and reduced particle sorting (the gut contained a broad range of particle sizes) may also be an adaptation of Amygdalum anoxicolum to the OMZ (Oliver 2001). Nearsurface sediments in many OMZs consist of sloppy, loose mud of high water content and low penetration resistance (Murray et al. 2000, Levin et al. 2002). The unusual morphological features of Glochinema bathyperuvensis (e.g. strong suction apparatus through a well-muscularised pharynx with an elongated posterior bulb as well as enlargement of the body surface area) have been interpreted as an effective adaptation to cope with soupy OMZ sediments (Neira et al. 2001a). Construction of compacted mud dwellings to offer stability may be one adaptation by sessile species to soupy sediments. Within the Oman margin OMZ (700–850 m) these dwellings take the form of nests in Amygdalum anoxicolum (Oliver 2001), mudballs in the cirratulid polychaete Monticellina sp. (Levin & Edesa 1997), tubes in spionid polychaetes (Levin et al. 1997, 2000), and arborescent and mud-walled tests in large Foraminifera (Gooday et al. 2000). Sediments of the Santa Catalina Basin, where O2 ⬃ 0.4 ml1, harbour similar mud-walled cirratulids (Tharyx luticastellus) (Smith 1986) and Foraminifera (Levin et al. 1991b). A variety of behavioural adaptations have been documented for plankton, including vertical migration (Childress & Siebel 1998) and ontogenetic migration (Wishner et al. 1998, 2000). Copepods in the genus Lucicutia occupy the lower OMZ interface in the eastern Paciﬁc Ocean and Arabian Sea (0.07–0.15 ml l1). They inhabit different oxygen zones during different developmental stages, and feed at up to four trophic levels (Wishner et al. 2000). Other zooplankton and mesopelagic ﬁsh migrate from surface waters into the OMZ on a diel basis or during particular life stages. They often escape predators or diapause within the OMZ or below, but feed and grow in oxygenated surface waters (Smith 1982, Smith et al. 1998). The scavenging amphipod, Orchomene obtusus, for example, appears to migrate into the anoxic bottom waters of Saanich Inlet, British Columbia to exploit abundant food and escape from predators and competitors (De Robertis et al. 2001), but will migrate upward into oxygenated waters to recover oxygen debt. Endemic aplacophoran molluscs within the OMZ on the summit of Volcano 7 have their mantle cavity permanently open (even after ﬁxation), exposing the respiratory folds. This appears to be an adaptation to improve respiration, as most neomenioid aplacophorans will 11

LISA A. LEVIN

close the mantle cavity tightly after preservation (A. Scheltema, pers. comm.). High ventilatory ability and circulation capacity have been documented in mid-water crustaceans as a possible adaptation to OMZs (Childress & Siebel 1998) but these have not been studied in benthic species. Enzymatic adaptations associated with anaerobic metabolism in OMZ benthos have received little attention. Recent work by Gonzalez & Quiñones (2000), however, suggest that an important anaerobic pathway involving lactate and pyruvate oxidoreductase, used in maintaining metabolic rate under environmental hypoxic conditions, may evolve adaptively in OMZs despite low yields of ATP per mol of glucose at high ATP production rates (Livingstone 1983). Gonzalez & Quiñones (2000) characterised the enzymes and activity of lactate and opine pathways for nine species of polychaetes inhabiting permanent and seasonally hypoxic zones of the central Chile OMZ. Each species was found to possess a different subset of pyruvate oxidoreductases (LDH, ALPDH, OPPDH and STRDH), with Nephtys ferruginea and Paraprionospio pinnata having all four. ALPDH was present in all species studied. Only those species with two or more pyruvate oxidoreductases were able to occupy the permanently dysoxic stations. Higher numbers of these enzymes may confer metabolic plasticity, and could explain the success of P. pinnata in hypoxic settings around the world. All species had similar Km values for the four enzymes. Enzyme activities of ALPDH and STRDH were strongly correlated with body size. Gonzalez & Quiñones (2000) propose that pyruvate oxidoreductases play a regulatory role in determining rate of pyruvate consumption during transition from dysoxic to anoxic conditions. Whether enzymatic adaptations can improve the efﬁciency of metabolism under conditions of permanent oxygen depletion has yet to be determined. Adaptation to sulphide toxicity in OMZ sediments has not been studied directly for metazoans. Association of metazoans with mat-forming, sulphide-oxidising bacteria may place them in less sulphidic microhabitats (Bernhard et al. 2000). It is likely that ultrastructure studies, as have been conducted for Foraminifera (reviewed in Bernhard & Sen Gupta 1999) will reveal a wealth of adaptive morphologies and symbioses. For example, the abundant lysosomes present in the stomodeal and oesophageal epithelia and head ventral epidermis of the polychaete Xenonerilla bactericola have been proposed to function like the sulphide-oxidising bodies of Urechis caupo, forming a peripheral defence against sulphide toxicity in sediments of the Santa Barbara Basin (Müller et al. 2001). It remains to be seen whether the sulphide-oxidising symbionts of other OMZ species (see later discussion) function in sulphide detoxiﬁcation as well as nutrition. Within OMZs, ﬂatﬁshes exhibit an increase in water content (Hunter et al. 1990) and decline in the metabolic capacity of white muscle (Vetter et al. 1994), probable adaptations to a hypoxic environment and limited food availability. The rockﬁsh Sebastolobus alascanus exhibits a number of respiratory, blood and heart enzyme adaptations in the OMZ off California (Yang et al. 1992). Recent reviews by Grieshaber et al. (1994), Diaz & Rosenberg (1995, 2001), Gray et al. (2002) and Karlson et al. (2002) have synthesised the responses and adaptations of shallowwater benthos to oxygen deﬁciency. Key adaptations described in these reviews are summarised here for comparison with benthos in open-ocean, deep-water OMZs. Shallow-water marine animals may experience hypoxia on a permanent basis, as occurs in enclosed water bodies subject to extreme eutrophication (e.g. Baltic Sea and Black Sea), but more often shallow environmental hypoxia will be seasonal, episodic or short term, resulting when nutrient or sewage loading, drainage, temperature and biotic cycles interact (e.g. Gulf of 12

OXYGEN MINIMUM ZONE BENTHOS

Mexico, Chesapeake Bay, Scandinavian waters) (Diaz & Rosenberg 2001, Gray et al. 2002, Karlson et al. 2002). Cycles of hypoxia may also occur at higher frequency in speciﬁc habitats such as intertidal sediments or rock pools during low tide (Gordon 1960). Shallow-water animals subject to periodic hypoxia may respire aerobically (aerobiosis) during periods of normoxia or moderate hypoxia but during severe hypoxia and anoxia, they may gain energy from anaerobiosis (environmental anaerobiosis), leading to a drastic reduction in energy gain. Anaerobiosis in invertebrates can also result from extreme physical activity, but is not considered in this discussion. Adaptations by shallow-water organisms to environmental oxygen limitation include increased respiratory movements to enhance oxygen uptake, changes in circulation, modulation of oxygen binding by respiratory pigments, reduced oxygen consumption and overall energy expenditure, as well as fermentation pathways for ATP synthesis (Grieshaber et al. 1994). Moderate hypoxia induces physiological responses such as increased ventilation in annelids, molluscs and crustaceans by peristaltic pumping or beating of appendages. Some molluscs and crustaceans (but not annelids) are able to increase their cardiac output to enhance blood ﬂow as well. For most shallow taxa, the ﬁrst responses to severe hypoxia are behavioural. Mobile taxa move away from the affected region or swim to the surface, causing dense clustering of animals in oxygenated waters. Sessile animals on the sea ﬂoor may, through elongation, tube construction, or migration from burrows, raise respiratory structures higher above the sediment/water interface to access faster moving water with more oxygen. Feeding and other activities not related to respiration may decline or cease. Most of these responses will lead to greater exposure to predators (or ﬁshermen) and thus to elevated mortality. These adaptations are unlikely to be helpful to animals subject to permanent severe hypoxia but could potentially be encountered near OMZ boundaries where oxygen concentrations ﬂuctuate due to tidal, seasonal, or climatic forcing. Recent research has revealed that shallow-water invertebrates that can survive prolonged periods of hypoxia or anoxia exploit a variety of anaerobic biochemical pathways to generate energy. These involve synthesis of lactate, acetate, priopionate and succinate. Some taxa also undergo metabolic depression (reduced rate of ATP turnover) to reduce energy requirements. Since hydrogen sulphide often forms under anoxic conditions, adaptations to reduce sulphide toxicity appear to be important to the most tolerant taxa. These adaptations include exclusion of sulphide at the body wall, insensitive cytochrome c oxidase, reversible sulphide binding to blood components, mitochondrial sulphide oxidation to less toxic compounds (e.g. thiosulphate) with ATP synthesis, reliance on anaerobic respiration at high sulphide levels, and oxidation of sulphide by symbiotic bacteria (Grieshaber & Volkel 1998). There have been few investigation of the importance of sulphide tolerance, anaerobic metabolism, or metabolic regulation in OMZ benthos (other than the investigation mentioned above by Gonzalez & Quiñones 2000), thus it is not possible to directly compare physiological responses of deep- and shallow-water to hypoxia. This remains a fertile area for research. A survey of the hypoxia literature reveals that speciﬁc stress responses are typically induced at much higher oxygen concentrations in shallow ecosystems than in OMZs, where animals have evolved to cope with permanent hypoxia. In shallow water, ﬁshes are most sensitive and can exhibit reduced larval growth, production, and feeding at oxygen concentrations below 4.5 mg l1 (3.2 ml l1) with mortality between 1 mg l1 and 2 mg l1. However, mudskippers and other bottom ﬁshes may respond only after oxygen falls below 1 mg l1 O2 (0.7–0.5 ml l1). Shallow-water infaunal species typically exhibit responses at oxygen concentrations below 1 ml l1 O2, but some annelids and molluscs can survive short periods of 13

LISA A. LEVIN

exposure to 0.5 ml l1 O2 or even anoxia. Generally, crustaceans and echinoderms are the next most sensitive, followed by annelids, priapulids and selected molluscs, although there are exceptions (Diaz & Rosenberg 1995, Gray et al. 2002, Karlson et al. 2002). This taxonomic sequence of hypoxia tolerance is similar to that observed in OMZs; however, annelids appear to be more tolerant than most molluscs to extreme oxygen deﬁciency (0.1 ml l1) in OMZs.

Community-level responses to OMZ conditions Oxygen minimum zones support benthic ecosystems that differ fundamentally from those in well-oxygenated environments. Aspects of animal community structure, including body size, abundance, taxonomic composition, diversity and lifestyles, are distinct within sediments intercepted by OMZs. Many of these community issues were ﬁrst explored by Sanders (1969) for the benthos off Walvis Bay, West Africa. Since then, additional research has been carried out in the eastern Paciﬁc Ocean and the Arabian Sea. Recent investigations have expanded the scope of study to include ecosystem-level processes such as bioturbation and energy ﬂow.

Body size and morphology Perhaps the most inclusive system-level response to OMZ conditions is altered size structure. At microxic oxygen levels (0.1 ml l1) the benthic fauna often consists mainly of protozoan and metazoan meiofaunal (or smaller)-size organisms; macrofauna and megafauna are typically rare or absent. On Volcano 7, a seamount off Mexico whose summit protrudes into the eastern Paciﬁc OMZ, O2 increases downslope in a linear fashion. Bacteria are abundant at the uppermost summit where oxygen is lowest, but few eukaryotic organisms occur there except nematodes (Levin et al. 1991a) and Foraminifera (Nienstedt & Arnold 1988). Macrofauna and megafauna appear along the lower summit as oxygen climbs above 0.1 ml l1 (Wishner et al. 1990, 1995). Similarly, Foraminifera, other protozoa, and metazoan meiofauna are the primary inhabitants of the microxic Santa Barbara Basin ﬂoor (Bernhard & Reimers 1991, Bernhard et al. 2000). Macrofauna and megafauna ﬁrst appear at the basin edges where bottom-water oxygen concentrations rise slightly (L. Levin & J. Bernhard, unpubl. obs.). Dissolved oxygen concentration and partial pressure (Spicer & Gaston 1999), through diffusion constraints, have been proposed to inﬂuence maximum body size in invertebrates, including shallow-water amphipods (Chapelle & Peck 1999, Spicer & Gaston 1999) and terrestrial insects during the Carboniferous period (Graham et al. 1995). McClain & Rex (2001) observed a positive relationship between maximum body size and oxygen concentration for nine species of turid gastropods (analysed together) from the western north Atlantic, and within a single species, Benthomangelia antonia. They propose that oxygen may be one factor explaining pervasive size-depth relationships among deep-sea taxa. However, none of these analyses involved animals from oxygen minimum zones. Presumably small-bodied organisms have a true advantage in microxic conditions, by presenting a larger surface area : volume ratio. They may also have greater metabolic ﬂexibility that confers ability to use food resources in the absence of oxygen (Zehnder 1988). Within OMZs, Foraminifera are the only group to exhibit clear within-taxon reductions in 14

OXYGEN MINIMUM ZONE BENTHOS

body size in response to oxygen depletion (reviewed in Bernhard & Sen Gupta 1999). These shifts may, however, reﬂect opportunistic life histories rather than physiological constraints. For macrofauna as a whole, there were no monotonic shifts in body size along oxygen gradients associated with the Oman margin (Levin et al. 2000), Peru margin (Levin et al. 2002), or central Chile margin (V. Gallardo & M. Palma, unpubl. obs.). In these settings, the largest body sizes occur at some stations within and some below the OMZ. Levin et al. (1994) noted increasing body size with decreasing oxygen in three polychaete species (Protodorvillea sp., Tharyx sp., Cirrophorus lyra) along the ﬂanks of Volcano 7. However, it is unclear whether this trend is related to oxygen or to food availability (measured as % org C and chlorophyll a), which varies inversely with oxygen. Among the metazoan meiofauna, nematode body size can be exceptionally large within OMZs (e.g. Neira et al. 2001a), also possibly due to excess food.

Abundance and biomass Foraminifera Foraminifera are perhaps the best-studied taxon within oxygen minimum zones. Detailed reviews of Foraminifera in oxygen-poor habitats have been written by Sen Gupta & Machain-Castillo (1993), Bernhard (1996), Bernhard & Sen Gupta (1999), and Bernhard et al. (2000). A number of experimental studies have addressed foraminiferal responses to oxygen depletion (e.g. Moodley & Hess 1992, Alve & Bernhard 1995, Moodley et al. 1997, 1998a,b). Hyaline calcareous Foraminifera (rotaliids) are broadly tolerant of low oxygen (Bernhard et al. 1997, Jorissen 1999b), although they cannot survive persistent anoxia (Bernhard & Reimers 1991). Foraminifera have been examined in marine sediment cores to provide indicators of environmental factors that are important in palaeoceanography. For example, faunal parameters such as species composition, the relative abundance of different test morphotypes, benthic to planktonic foraminiferan abundance ratios, and benthic Foraminifera accumulation rates, have been used as proxies for surface productivity and its seasonality (Loubere & Fariduddin 1999), organic carbon ﬂux to the sea ﬂoor (Corliss & Chen 1988, Herguera & Berger 1991, van der Zwaan et al. 1999), and bottom water oxygenation (Kaiho 1994, 1999, Baas et al. 1998). However, these relationships may shift within OMZs, where dissolved oxygen exerts stronger control, for example, on foraminiferal accumulation rates (Naidu & Malmgren 1995). Most research has addressed the fossilisable component (with calcareous or multilocular agglutinated tests); only recently have investigations considered the fragile tubular agglutinated and soft-bodied, monothalamous forms. Foraminifera in the Arabian Sea, eastern Paciﬁc, and in dysoxic or microxic basins exhibit parallel ecological responses to oxygen (Gooday et al. 2000, Levin et al. 2002). Foraminifera tend to dominate numerically over metazoan meiofauna and macrofauna in oxygen-deﬁcient settings (Gooday et al. 2000). Surface densities of Foraminifera are typically much higher within than below the OMZ. OMZ densities are 29 820 ind. 50 cm3 at 300 m off Peru (150 m), 16 107 ind. 10 cm2 at 412 m off Oman (63 m), and up to 9626 ind. 10 cm2 at 550 m in the Santa Barbara Basin (63 m) (Gooday et al. 2000, Levin et al. 2002). A benthic transect across the central California OMZ (624–3728 m) revealed an inverse correlation between foraminiferal densities in the upper 0.5 cm and bottom-water oxygen concentration (Bernhard 1992). Release from macrofaunal and megafaunal predation and enhanced food supply are proposed to account for this pattern (Bernhard 1992). 15

LISA A. LEVIN

Metazoan meiofauna Based on OMZ studies in the Arabian Sea (Cook et al. 2000), and in the eastern Paciﬁc Ocean off Mexico (Levin et al. 1991a), Peru (Neira et al. 2001b, Levin et al. 2002) and Chile (Neira et al. 2001c), it appears that total densities of metazoan meiofauna are never reduced within OMZs. Rather, meiofaunal densities reach maximum values at lowest oxygen concentrations within OMZs, presumably due to abundant particulate food and/or reduced predation intensity (Neira et al. 2001b). Meiofaunal densities exhibit a strong positive correlation with indices of food availability on Volcano 7 in the eastern Paciﬁc off Mexico (chlorophyll a; Levin et al. 1991a) and with food quality off Oman (hydrogen index; Cook et al. 2000) and Peru (chloroplastic pigment equivalents and labile organic compounds; Neira et al. 2001b). On the central Chile shelf, meiofaunal abundance was positively correlated with Thioploca biomass, but nutritional relationships have not been established (Neira et al. 2001c). Experiments are needed to explore the signiﬁcance of high food quality or reduced predation in releasing meiofaunal densities within OMZs. Meiofauna appear to be more broadly tolerant of oxygen depletion than are macrofauna (Giere 1993). Because nematodes are the dominant metazoan meiofaunal group present at low oxygen concentrations (typically they are 95–99% of total meiofauna), the abundance patterns described above are driven largely by nematode responses. Nematode densities may be 2–5 times higher within than beneath OMZs (Levin et al. 1991a, Cook et al. 2000, Neira et al. 2001a,b). Gastrotrichs and nerillid polychaetes also can be abundant within OMZs (Todaro et al. 2000, Müller et al. 2001, Levin, unpubl. obs.). In contrast, harpacticoid copepod densities are much reduced within OMZs and exhibit a positive correlation with bottom-water oxygen concentration. Other metazoan meiofaunal taxa that appear to avoid microxic conditions include kinorhynchs, tardigrades, rotifers and non-nerillid polychaetes (Neira et al. 2001b). On the central Chilean shelf (120 m), which experiences persistent OMZ-associated hypoxia, El Niño events signalled signiﬁcant bottom-water oxygenation and reduced organic matter input and quality (Gutiérrez et al. 2000, Neira et al. 2001c, Sellanes 2002). At the end of the 1997–8 El Niño, harpacticoid copepod densities were nine times higher than during the previous year, but total meiofaunal densities were 42% lower (Neira et al. 2001c, Sellanes 2002). These temporal responses to ENSO-related oxygen variation are consistent with oxygen forcing of spatial variation across OMZs. Macrofauna Densities of macrofauna are often depressed within the part of the OMZ where oxygen concentrations are lowest, unlike protozoan and metazoan meiofauna, whose abundances are usually maximal in the OMZ core (Fig. 5). OMZ-related density reductions have been observed on the Walvis Bay margin (Sanders 1969), off central California (Mullins et al. 1985, Hyland et al. 1991), on Volcano 7 off Mexico (Levin et al. 1991a), and on the Oman margin (Levin et al. 2000), but not on the Peru and Chile margins (Levin et al. 2002, V. Gallardo et al., unpubl. obs.). Densities may be 30% to 70% lower in the OMZ core (e.g. 0.15 ml l1) than in other parts of the OMZ. However, there is no clear relationship between bottom-water oxygen concentration and absolute density when OMZ macrofauna are compared globally. Macrofaunal densities estimated for OMZ sediments using a 300-m mesh range from 1854 ind. m2 in the sandy sediments on Volcano 7 (750 m, 0.08 ml l1 O2) to 21 380 ind. m2 on the central Chile margin (200 m, 0.13 ml l1 O2) (see Table 1). Densities up to 60 000 ind. m2 may occur in seasonally dysoxic shelf waters (Gutiérrez et al. 2000). Macrofauna also exhibit reduced biomass where oxygen levels are lowest (e.g. Rowe 1971, Rosenberg et al. 1983, Levin et al. 2000). Low standing stock was observed in the 16

OXYGEN MINIMUM ZONE BENTHOS

(a)

(b)

(c)

(e)

(d)

Figure 5 Vertical transects of macrofaunal density (no. ind. m2) across OMZs at continental margins off (a) Peru, (b) Walvis Bay, (c) Oman, (d) Mexico (Volcano 7) and (e) central California. Note depressed density in OMZ core and maxima near OMZ edges.

OMZ off West Africa (Hart & Currie 1960, Sanders 1969). Rowe (1971) documented a macrofaunal biomass minimum at 329 m off Peru (0.008 g C m2, 15°S), where oxygen concentrations were 0.1 ml l1, and a maximum at 875–1000 m (4.3–5.4 g C m2), below the OMZ where oxygen 1.0 ml l1 (20% saturation). These biomass values exhibit the same 17

LISA A. LEVIN

trend reported for macrofaunal wet weight on the Peru margin by Rosenberg et al. (1983) (3–15°S) and by Levin et al. (2002) at 12°S during the 1998 El Niño. Rowe (1971) suggested reduced biomass within the OMZ is a manifestation of the stress induced by low oxygen but notes that this low biomass involves high macrofaunal densities, indicating small body size. The entire OMZ does not exhibit reduced macrofaunal biomass. Levin et al. (2002) observed a wet weight biomass of 52 g m2 at 562 m on the Peru margin (O2 0.26 ml l1) and V. Gallardo et al. (unpubl. obs.) report biomass of 60.9 g m2 at 100 m on the central Chile margin (O2 0.10 ml l1). Similarly, high biomass values were observed at 700 m (59.7 g m2, O2 0.16 ml l1) and 1000 m (43.5 g m2, O2 0.27 ml l1) on the Oman margin (Levin et al. 2000). Megafauna Megafaunal responses to OMZs mirror those of most macrofauna (Thompson et al. 1985, Wishner et al. 1990). Megafauna appear to be absent or nearly so in the cores of most OMZs where bottom-water oxygen concentration falls below 0.15 ml l1 (Wishner et al. 1990, 1995, Smallwood et al. 1999). Hermit crabs are present in the OMZ core off central California, but oxygen concentrations here are relatively high (0.27 ml l1) (Thompson et al. 1985). Extraordinarily high densities of megafauna can be found near OMZ lower boundaries, where dense aggregations of crustaceans and echinoderms occur (Fig. 6). Dense megafaunal assemblages on Volcano 7 include sponges (11.9 ind. m2) (Fig. 6A), shrimp (Heterocarpus nesisi, Benthescymus altus, 3.1 ind. m2) (Fig. 6C), and galatheid crabs (Munidopsis cf. hystrix, 5.6 ind. m2) (Fig. 6A) (Wishner et al. 1995); galatheids (M. scobina) (Creasey et al. 2000) and spider crabs (Encephaloides armstrongi, 47 ind. m2 (Smallwood et al. 1999)) (Fig. 6D) on the Oman margin; echinoids off northern California (Brissopsis paciﬁca, 14 ind. m2, Thompson et al. 1985), and ophiuroids in many areas (Fig. 6B) (Ophiolymna antarctica off Oman, 51 ind. m2, Smallwood et al. 1999; unidentiﬁed species off Peru, 140 ind. m2, Levin et al., unpubl. obs.; unidentiﬁed species off central California, 50 ind. m2, Thompson et al. 1985; Ophiopthalmus normani, 16.7 ind. m2 in the Santa Catalina Basin where O2 0.4 ml l1, Smith & Hamilton 1983). Feeding by abundant megafauna is proposed to deplete the organic C and the sterol composition of seabed sediments (Smallwood et al. 1999). Because epibenthic megafauna can be quantiﬁed photographically, without destructive sampling, there are transect analyses that provide a much more detailed picture of spatial distribution in OMZs than is available for the infaunal taxa. Photographic analyses of megafauna on the summit of Volcano 7 revealed an apparent zonation across the lower OMZ boundary, much like that which occurs in the intertidal zone (Wishner et al. 1995). Based on temperature records, investigators determined that animals on the summit are exposed to waters whose oxygen content oscillates with internal tides; those on the lower summit (800–850 m) are exposed to higher bottom-water oxygen concentrations over time than benthos on the upper summit at 730 m (Levin et al. 1991a, Wishner et al. 1990, 1995). At water depths 740 m (0.1 ml l1) the fauna was visibly depauperate; only rat-tail ﬁshes (Fig. 6E), sponges, seapens and an unidentiﬁed coelenterate were observed. Just below this (750–770 m), high densities of sponges, crabs, and serpulid polychaetes occurred (Fig. 6B), followed by shrimp (Fig. 6A), ophiuroids (Fig. 6C) and anemones at 770–800 m, and then by brisingid asteroids (925 m) (Wishner et al. 1990, 1995). Similar sharp bands of zonation have been observed for echinoderms on the central California margin. Asteroids (Rathbunaster californicus) dominate at the upper OMZ boundary (400–500 m), hermit crabs are 18

OXYGEN MINIMUM ZONE BENTHOS

Figure 6 Dense megafaunal aggregations at OMZ edges. (A) Sponges, and serpulids (Hyalopomatus merenzelleri), Volcano 7, Mexico, depth 780 m. (B) Ophiuroids Volcano 7, Mexico, depth 780 m. (C) Shrimp (Heterocarpus nesisi and Benthescymus altus) Volcano 7, Mexico, depth 770 m. (D) Spider crabs (Majidae) Encephaloides armstrongi, Oman margin, depth 1000 m. Photograph courtesy of Brian Bett, copyright Southampton Oceanography Centre. (E) Rat-tails Nezumia liolepus, Volcano 7, depth 800 m.

abundant in the OMZ core, the echinoid Brissopsis paciﬁca is most abundant near the lower OMZ boundary (900 m) and ophiuroids attain highest densities below the OMZ) (Thompson et al. 1985). This overview of taxon abundances indicates that protozoan and metazoan meiofauna exhibit similar numerical responses to oxygen gradients (e.g. r2 0.99 on the Peru margin (Levin et al. 2002)). Macrofauna and megafauna exhibit responses that are distinct from those of small organisms. Meiofaunal densities increase with lowered oxygen concentration, macrofaunal densities decrease or stay constant, and megafauna densities decline. Although the macrofauna and megafauna are less tolerant to microxic conditions, in at least one setting symbiosis appears to permit small macrofauna to attain high densities at exceptionally low oxygen concentrations (Levin et al. 2002).

19

LISA A. LEVIN

Edge effects Near the upper or lower boundaries of OMZs, benthic faunal densities often exhibit a maximum at speciﬁc oxygen concentrations that may represent a physiological threshold. This response is believed to reﬂect the interaction of a rich supply of organic matter with release from oxygen limitation. In some regions, and for selected taxa, this occurs within OMZs at extremely low levels of oxygen (0.1–0.2 ml l1). Elsewhere this occurs closer to 0.5 ml l1, near OMZ boundaries. Such density maxima have been observed for macrofauna off central California (Mullins et al. 1985, Thompson et al. 1985), West Africa (Sanders 1969), Mexico (Volcano 7) (Levin et al. 1991a, Wishner et al. 1995), Peru (Rowe 1971) and the Oman margin in the Arabian Sea (Levin et al. 2000) (Fig. 5). Along the central California margin, maximal foraminiferal densities occur near 1000 m at the OMZ edge (0.48 ml l1) (Bernhard 1992). On phosphorite hardgrounds along the Peru margin, agglutinated Foraminifera exhibit maximum densities (up to 97 cm2) at the upper (162 m) and lower (465–620 m) OMZ boundaries (Resig & Glenn 1997). This phenomenon has sometimes been called a boundary or edge effect but, as noted above, these maxima can occur inside OMZ boundaries. Both macrofauna and megafauna exhibit this pattern across a range of taxa, although thresholds are lower for macrofauna, and vary among megafaunal species (Thompson et al. 1985, Wishner et al. 1995). As a result, OMZ edges typically support dense aggregations of asteroids, ophiuroids, shrimp, crabs and sponges (see above discussion and Fig. 6). Similar edge effects are observed for zooplankton near lower OMZ boundaries in the eastern Paciﬁc (Wishner et al. 1995) and Arabian Sea (Morrison et al. 1999, Wishner et al. 2000). Several lines of evidence suggest that upper and lower OMZ boundaries are areas of enhanced biological activity. In the water column, elevated microbial activity, including chemolithic production, has been documented near lower OMZ boundaries (Owen & Zeitzchel 1970, Karl & Knauer 1984, Karl et al. 1984, Mullins et al. 1985) and is reﬂected in plankton diets (Gowing & Wishner 1992, 1998). Some of this production may fuel bottom populations of microbes, protists, and metazoans. However, there are no quantitative studies that examine the fate of in situ OMZ production or rates of sediment microbial activity across oxygen gradients in seaﬂoor OMZs.

Community composition Faunal assemblages within OMZ often exhibit distinct taxonomic trends. Foraminiferan OMZ assemblages consist mainly of calcareous forms; the genera Fursenkoina, Gobobulimina, Nonionella, Bolivina, Bulimina, Cassidulina, Uvigerina, Epistominella and Hoeglundina are particularly well represented. However, agglutinated taxa like Bathysiphon and Trochammina can occur as well (Bernhard et al. 1997, Gooday et al. 2000). Encrusting, agglutinated species are abundant on phosphorite crusts in the Peru margin OMZ, particularly at upper and lower edges (Resig & Glenn 1997). Most infaunal Foraminifera (and other meiofauna) are more likely to be inﬂuenced by the microxic pore-waters in OMZ sediments, than by bottom-water oxygen concentration (e.g. Gooday et al. 2000, Murray 2001, Fontanier et al. 2002), although the two parameters are necessarily related. It is unclear whether infaunal Foraminifera associated with OMZs are responding to elevated organic matter availability (i.e. associated with high sulphate reduction rates) or to low oxygen 20

OXYGEN MINIMUM ZONE BENTHOS

(Jorissen et al. 1995, Jorissen 1999a; see papers reviewed in Gooday & Rathburn 1999). Foraminiferal assemblages in better-oxygenated stations below the OMZ exhibit increased representation of agglutinated and soft-bodied (saccamminid and allogromid) taxa relative to those within the OMZ (Gooday et al. 2000, Levin et al. 2002, A. Rathburn, pers. comm.). Off Peru, the ratio of calcareous to agglutinated Foraminifera shifted from 10.7 and 8.5 within the OMZ (stations with 0.02 ml l1 and 0.26 ml l1 O2, respectively), to 1.73 and 0.13 below the OMZ (0.84 and 1.7 ml l1 O2, respectively) (Levin et al. 2002). Some large, agglutinated Foraminifera, including Pelosina spp. and mud-walled astrorhizinids, may be present in OMZs. They have been recorded at O2 levels 0.5 ml l1 off Oman and Peru, and in the Santa Catalina Basin, but they apparently cannot tolerate O2 concentrations below 0.2 ml l1 (Gooday et al. 2000). It is not clear why, since agglutinated tests should be less energetically expensive to maintain than calcareous tests under conditions of low oxygen and low pH. Diffusion of oxygen could constrain the larger-bodied protozoans, but even small agglutinates are not common under conditions of severe oxygen depletion. Among the metazoans, soft-bodied forms dominate in the core of OMZs (e.g. Thompson et al. 1985, Levin et al. 1997). In contrast to the Foraminifera, calciﬁed invertebrates are absent or only weakly calciﬁed where oxygen levels fall below 0.3 ml l1 (Rhoads and Morse 1971, Thompson et al. 1985). Rhoads and Morse (1971) state that “low calcium concentration or carbonate dissolution are not responsible for the relative absence of living calcareous forms in low oxygen marine basins”. Rather, they propose that periodic anaerobic metabolism leads to production of lactic and pyruvic acids; these are buffered by dissolution of calciﬁed structures like shells. Many bivalves living in reducing environments display a dull inner shell surface indicative of dissolution. Gastropods (Astyris permodesta) and bivalves (Amygdalum anoxicolum) that occur where O2 0.15 ml l1 have very thin shells. However, as indicated above, massive densities of calciﬁed echinoderms are commonly found at OMZ lower boundaries (Thompson et al. 1985, Wishner et al. 1995, Levin et al. 2002), so oxygen concentrations of 0.3 ml l1 to 0.5 ml l1 appear manageable for these taxa (J. B. Thompson et al. 1985, B. Thompson et al. 1993). Nematodes are the metazoan meiofaunal group most tolerant of low oxygen (Levin et al. 1991a, Moodley et al. 1997, Cook et al. 2000, Gooday et al. 2000, Neira et al. 2001a,b). Kinorhynchs, rotifers, and ostracods are present in OMZs but are poorly represented. Harpacticoid copepods are among the meiofauna most sensitive to low oxygen and are consistently rare or absent within well-developed OMZs (Levin et al. 1991a, Neira et al. 2001a,b). These taxonomic responses to oxygen are reﬂected in changing ratios of nematodes to harpacticoid copepods. On the summit of Volcano 7 and on the Peru margin nematode : harpacticoid ratios are very high (500 : 1 and 65 : 1, respectively) within the OMZ, but much lower beneath the OMZ (Levin et al. 1991a, Neira et al. 2001b). Among macrofauna, annelids dominate, typically accounting for 90% or more of fauna in muddy OMZ sediments (Table 1). Although these are often polychaetes, tubiﬁcid oligochaetes are common as well. In a nearly anoxic basin off Peru, gutless tubiﬁcid oligochaetes comprised 83% of the macrofauna (Levin et al. 2002). Within foraminiferal sands on the summit of Volcano 7, polychaetes (including pogonophorans) comprised only 39% of the fauna and an aplacophoran species was dominant (47%) (Levin et al. 1991a). Among the Annelida, patterns of family representation within OMZs may reﬂect hydrodynamic, bathymetric or geochemical conditions more than oxygen concentration. On the Oman margin, upper slope OMZ sediments (400–700 m) are dominated by spionid and cirratulid polychaetes, whereas lower OMZ sediments (850–1000 m) support mainly 21

LISA A. LEVIN

Table 1

Density, biomass, and diversity estimates for macrobenthos in OMZ sediments.

Location

Position

Oman margin Oman margin Oman margin Oman margin Central Chile margin Central Chile margin N. Chile margin N. Chile margin N. Chile margin Peru margin

19°22N 58°15E 19°19N 58°16E 19°14N 58°23E 19°16N 58°31E 36°26S 73°24W 35°44S 73°04W 20°15S 70°12W 20°16S 70°14W 20°17S 70°15W 6°14S 81°05W

N. Peru margin Central Peru margin Peru margin Peru margin Volcano 7 off Mexico Volcano 7 off Mexico Magdalena Bay, Baja CA

Water depth (m)

O2 concentration (ml l1)

Mesh size (microns)

400 700 850 1000 120 206 100 200 300 126

0.13 0.16 0.20 0.27 0.10 0.13 0.24 0.24 0.24 ?

300 300 300 300 300 300 300 300 300 500

6°33S 80°58W 11°57S 77°21N 12°22S 77°29W 12°32S 77°29W 13°25N 102°35W 13°25N 102°35W 23°26N 111°34W

380 107 305 562 745–767 788–857 713

0.15 0.20 0.02 0.26 0.09 0.13 0.05

500 500 300 300 300 300 300

Santa Catalina Basin Santa Barbara Basin Edge

32°58N 118°22W 34°13N 120°2W

1130 555

0.41 0.05

420 300

Central California

40°47N 124°36W

500

0.24–0.48

300

* based on number of individuals. ** includes pogonophorans.

ampharetid and paranoid polychaetes (Levin et al. 1997, 2000). Off Central Chile, cossurid and cirratulid polychaetes dominate in the OMZ core (100–200 m), paranoid and amphinomid polychaetes dominate at the OMZ edge (364 m), and maldanid polychaetes appear as dominants beneath the OMZ (810 m) (V. A. Gallardo et al., unpubl. obs.). Spionid, parnaoid, magelonid and cirratulid polychaetes are usually the dominant macrofauna in shallow OMZ sediments on the Peru and Chile margins (Gallardo et al. 1995, Tarazona et al. 1988a,b, Gutiérrez et al. 2000, D. Gutiérrez & V. A. Gallardo, unpubl. obs.). Frankenberg & Menzies (1968) reported an assemblage comprised of 96% cirratulid polychaetes off Peru at 126 m. Ampharetid polychaetes comprise most of the polychaete assemblage at 700 m off central California (0.3 ml l1; Mullins et al. 1985), but limited sampling suggests that cossurid polychaetes form over half of the OMZ assemblage at 700 m near Magdalena Bay, Mexico (L. Levin et al., unpubl. data). In the deeper sandy OMZ sediments on Volcano 7 (750–850 m), pogonophoran, cirratulid, dorvilleid, parnaoid and ampharetid polychaetes were dominant (Levin et al. 1991a). With the exception of the dorvilleids and pogonophorans, however, these taxa are also common in well-oxygenated bathyal environments (Jumars & Gallagher 1982). 22

OXYGEN MINIMUM ZONE BENTHOS

Density (ind. m2)

Biomass (g m2)

% Annelid*

Rank 1 dominance (taxon)

H (log base 2)

Reference

12 362 19 183 16 383 5818 16 478 21 381 1342 141 1384 2373

14.2 59.7

96 90 90 90 97% 70% 46 100 98 97

66 (Prionospio sp.) 40 (Prionospio sp.) 23 (Aricidea sp.) 24 (Lysippe sp.) 48 (Cossura chilensis) 23 (Tubiﬁcidae) magelonid polychaete spionid polychaete 73 Diaphorosoma sp. 96 (cirratulid sp.)

1.45 2.74 3.53 4.07 2.07 2.89

Levin et al. 1997, 2000 Levin et al. 2000 Levin et al. 2000 Levin et al. 2000 Gallardo et al., in prep. Gallardo et al., in prep. Gallardo et al., unpubl. Gallardo et al., unpubl. Gallardo et al., unpubl. Frankenberg & Menzies 1968 Gutierrez, pers. comm. Gutierrez, pers. comm. Levin et al. 2002 Levin et al. 2002 Levin et al. 1991a Levin et al. 1991a Levin et al., unpubl. obs. Jumars 1974 Beaudreau & Levin, unpubl. obs. Levin et al., unpubl. obs.

890 13 16 233 9151 1854 8457 8877 1880 1691 16 552

43.5 60.9 16.9 14.5 0.13 2.5 49.5 38.2 0.01 8.6 52

5.1

68.5 100 86 50 39** 71 98

65 Amphicteis sp. 1.02 100 Paraprionospio pinnata 83 (Olavius crassitunicatus) 2.28 37 (Ampelisca sp.) 4.52 47 (Lepidomeniidae) 21 (anemone) 58 (Cossura sp.)

83 44

20 Paraonis gracilis 44 (Tubiﬁcidae)

53

19 (tanaid sp.)

4.46

Echinoderms, crustaceans and molluscs are much less tolerant to hypoxia than annelids (Diaz & Rosenberg 1995, Levin & Gage 1998, Levin et al. 2000). On average, echinoderm tolerance is less than that of crustaceans, which is less than that of molluscs. The rarity of these taxa contributes to notable declines in species richness within OMZ sediments (Levin & Gage 1998). However, there are certain taxa common or endemic to OMZs that represent clear exceptions. Among these are ampeliscid amphipods, the gastropod Astyris permodesta, ophiuroids, and the endemic mussel Amygdalum anoxicolum. A number of ﬁsh species exploit oxygen minimum zones. On the California margin these include commercially valuable species such as thornyhead rockﬁshes (Sebastolobus altivelis and S. alascanus) (Jacobson & Vetter 1996), Dover sole (Microstomus paciﬁcus) (Hunter et al. 1990, Jacobson & Hunter 1993), sableﬁsh (Anaplopoma ﬁmbria) (Adams et al. 1995), and grenadier (Macrouridae) (P. Smith, pers. comm.). On Volcano 7 off Mexico, the rat-tail Nezumia liolepus was the only ﬁsh abundant under dysoxic conditions (Wishner et al. 1990, Fig. 6E). These species either feed on bottom fauna or feed on other ﬁshes that do, thereby coupling pelagic and benthic resources within OMZs (e.g. Wakeﬁeld & Smith 1990). Dover 23

LISA A. LEVIN

sole, like a number of ﬁshes off California, exhibit ontogenetic migrations, with juveniles living on the shelf and 98% of its spawning biomass located deeper, within the OMZ (Hunter et al. 1990). Oxygen interfaces may be important to these species as sites of aggregation (e.g. Wishner et al. 1995) or as refugia from predation or ﬁshing (Jacobsen & Hunter 1993, Jacobsen & Vetter 1996).

Are there OMZ endemics? If OMZs comprise a broad, continuous habitat, OMZ taxa might be expected to have global distributions, whereas if OMZs function as isolated habitats, a high degree of endemism should develop. Among the OMZs that have been studied, there is signiﬁcant variability in the species-level identity of the dominant macrofaunal species. At 400 m on the Oman margin, most individuals belong to two species, a spionid Prionospio sp. and a cirratulid, Aphelochaeta sp. (Levin et al. 1997). These appear to be new species and may be endemic to the region (A. Mackie, pers. comm.), as is the co-occurring mytilid, Amygdalum anoxicolum (Oliver 2001) and ascidian Styela gagetyleri (Young & Vàzquez 1997). The gutless, tubiﬁcid oligochaete, Olavius crassitunicatus (Phallodrilinae), appears limited to hypoxic settings on the Peru and Chile margin (Finogenova 1986, Levin et al. 2002, L. Levin, unpubl. obs.). The nematode Glochinema bathyperuvensis (Epsilonematidae) is also endemic to the Peru margin OMZ (Neira et al. 2001a). The gastropod Astyris permodesta (formerly Mitrella permodesta) is common to microxic settings off California (Bernhard & Reimers 1991) and Peru (Levin et al. 2002, L. Levin, pers. obs.), but also occurs at methane seeps off California and Oregon where waters have more oxygen (0.4 ml l1 O2 to 0.8 ml l1 O2) (unpubl. obs.). In coarser foraminiferal sands on the microxic summit of Volcano 7, an undescribed lepidomeniomorph aplacophoran and pogonophorans dominate (Levin et al. 1991a). On the Santa Barbara Basin ﬂoor a meiofaunal polychaete, Xenonerilla bactericola (Müller et al. 2001) and a gastrotrich, Urodasys anorektoxys (Todaro et al. 2000) occur at high densities and appear to be endemic species. Tubiﬁcid oligochaetes are the most abundant taxon at the edge of the Santa Barbara Basin (0.05 ml l1 at 555 m), but there is surprising macrofaunal diversity that includes crustaceans, echinoderms and aplacophorans (Beaudreau 1999). Two taxa that seem particularly widespread at OMZs are ampeliscid amphipods and lucinid bivalves. They are abundant at numerous OMZ sites within the eastern Paciﬁc and in the Arabian Sea. It is likely that OMZs support many more characteristic or endemic species, but many remain undescribed, and most have yet to be sampled.

Species diversity within OMZs Careful diversity studies have yet to be conducted for microbes, metazoan meiofauna, megafauna or ﬁshes within OMZs. Most diversity research has focused on Foraminifera and macrofauna; these groups tend to show similar responses to oxygen depletion (Levin et al. 2001). Foraminifera Species richness of Foraminifera is characteristically reduced and dominance is high in oxygen-depleted basins and margins (Phleger & Soutar 1973, Douglas 1981, Sen Gupta & Machain-Castillo 1993, Bernhard et al. 1997). There is lower diversity 24

OXYGEN MINIMUM ZONE BENTHOS

under microxic than dysoxic conditions (Gooday et al. 2000). Neinstedt & Arnold (1988) reported very low diversity on the microxic summit of Volcano 7 where rotaliid forms were dominant and 50–61% of the assemblage consisted of a single species. Similarly, high rank 1 dominance has been reported by Perez-Cruz & Machain-Castillo (1990) in the Gulf of Tehuantepec, Mexico (64% Epistominella bradyana), by Bernhard et al. (1997) and Phleger & Soutar (1973) in the Santa Barbara Basin (80–86% Nonionella stella), and by Gooday et al. (2000) on the Oman margin (44% Bolivina seminuda). Diversity shifts in the foraminiferan fossil record apparently reﬂect changes in the intensity of the oxygen minimum. Recognition of low-diversity assemblages extends back into the Cretaceous, where they were associated with widespread anoxia in continental margin basins (Koutsoukos et al. 1990). Fluctuations in foraminiferan diversity are often linked to temporal shifts in productivity or on shorter timescales, to circulation changes. Oscillations or shifts between low diversity assemblages with high dominance and high diversity assemblages with low dominance have been observed over 23 000-yr cycles in the northern Arabian Sea (den Dulk 1998), on the Pakistan margin during the Younger Dryas (11.5–12.7 103 yr BP), and in the Santa Barbara Basin on decadal to millennial timescales (Cannariato et al. 1999). Macrofauna The taxonomic shifts described previously for macrofauna are translated into changes in both dominance and species richness. Levin et al. (2001) review the inﬂuence of hypoxia on diversity of OMZ macrofauna. In all OMZ regions studied, macrofaunal species richness is reduced and dominance is extraordinarily high. Sanders (1969) was the ﬁrst to note this pattern off Walvis Bay, West Africa, and it has been documented since in the OMZs of the eastern Paciﬁc on Volcano 7 (Levin et al. 1991a), the Peru and Chile margin (Levin et al. 2002, V. Gallardo et al., unpubl. obs.) and in the Arabian Sea (Levin et al. 2000). Macrofaunal richness of 12 species or lower has been recorded in some OMZ settings (Levin et al. 1997, 2002, Bernhard et al. 2000). Where richness is low, the top-ranked species may comprise 47–87% of the total macrofauna (Table 1). Measures of rank 1 dominance and rarefaction richness (Es100), evaluated as a function of oxygen for bathyal sites around the world, indicate that the oxygen inﬂuence on macrofaunal diversity is greatest at oxygen concentrations below about 0.3 ml l1 or 0.4 ml l1 (Levin & Gage 1998). A major cause of reduced species richness within OMZs is the loss of taxa that are intolerant of low oxygen, for example, most echinoderms, crustaceans and molluscs. However, reduction in richness also occurs within tolerant taxa such as the annelids (Levin et al. 1994, Levin & Gage 1998). Organic enrichment may contribute to reduced diversity within OMZs, independent of oxygen. Separating the effects of oxygen availability from those of organic enrichment within OMZs is difﬁcult, as these factors vary inversely. Multiple regression analyses of diversity data for polychaetes from the eastern Paciﬁc and Indian Oceans suggest that oxygen exerts greatest control on species richness, but organic matter availability has more inﬂuence on measures of dominance and evenness (Levin & Gage 1998). In the Arabian Sea, sediment pigment concentration and oxygen concentration together explained 99% of variation in two measures of macrofaunal diversity (Es[100] and H) at stations between 400 m and 3400 m (Levin et al. 2000).

25

LISA A. LEVIN

Spatial heterogeneity Where horizontal spatial pattern has been studied, it appears that oxygen-stressed macrofaunal communities lack the structural heterogeneity and biotic complexity of oxygenated systems. On the Oman margin, within-station core similarity of communities was much greater inside the OMZ (multivariate index of dispersion (MID, 0.39–0.96) than beneath it (1.49–1.67) (Levin et al. 2000). A similar result was obtained for macrofauna on the central Chile margin, with greatest homogeneity (MID, 0.59) where oxygen concentration was lowest (V. Gallardo et al., unpubl. obs.). The increased homogeneity of assemblages within the OMZ may be due to an absence of large burrowers and associated biogenic structures that function to create considerable environmental patchiness for macrofauna (Jumars & Eckman 1983). The vertical distribution of infauna within the OMZ sediment column is not related to oxygenation of overlying water in a straightforward manner. One might expect infauna to dwell closer to the sediment surface under conditions of bottom-water oxygen depletion, but the reverse is sometimes observed. In the microxic sediments of the Peru OMZ, most meiofauna are found 1–5 cm into the sediment column; vertical distributions shallow as oxygen concentration increases (Neira et al. 2001b). The same is true for macrofauna on the Peru margin (Levin et al. 2002), in the core of the Oman margin OMZ (Smith et al. 2000), and on the northern Chile margin OMZ (L. Levin, unpubl. data). However, on the central Chilean shelf, nematodes within the OMZ are concentrated near the surface, and deepen their vertical distribution during El Niño-induced oxygenation events (Neira et al. 2001c). Benthic Foraminifera have shallower vertical distributions within than below the OMZ on the northern California margin, but subsurface abundance peaks were evident in some OMZ samples (Bernhard 1992). The vertical distribution of various sediment biochemical components might affect infaunal distributions and feeding within OMZ sediments (Neira et al. 2001c); carbohydrates and proteins could play contrasting roles in different sediment layers (Danovaro et al. 2001). Thus, faunal vertical distributions may be modulated by trophic processes as well as sulphide and oxygen. Further studies are required to understand the mechanisms and factors controlling the horizontal and vertical distributions of infauna in OMZs.

Reproductive modes There have been few inclusive studies of reproduction in OMZ faunas. As with most environments, it appears that a range of development modes are represented, in part due to taxonomic constraints. At 400 m on the Oman margin (0.13 ml l1 O2), there are abundant species with planktotrophic larval development (e.g. Prionospio (Minuspio) sp. and Amygdalum anoxicolum) and with lecithotrophic development (Aphelochaeta sp.) (Levin et al. 1997, Oliver 2001). Some of the most abundant megafauna, such as the spider crab Encephaloides armstrongi, have planktotrophic development. Spider crabs exhibit a highly skewed sex ratio in the Oman margin OMZ, suggestive of reproductive migration (Creasey et al. 1997). The galatheid crab Munidopsis scobina exhibits bimodal egg sizes on the Oman margin, with increasing occurrence of small eggs near the OMZ (Creasey 1998). Macrofauna along the ﬂanks of Volcano 7, whose summit penetrates the eastern Paciﬁc OMZ, exhibited an increase in reproductive activity and per cent brooding with decreasing

26

OXYGEN MINIMUM ZONE BENTHOS

oxygen upslope (Levin et al. 1994). At the microxic summit of this seamount, over 63% of the polychaetes collected during November–December 1998 were reproductive; many appeared to be opportunistic brooders (e.g. Protodorvillea sp., Cirrophorus lyra and Tharyx sp.) with lecithotrophic or direct development. On Volcano 7, brooding was observed among 86% of the polychaetes on the summit (0.08 ml l1) and 81% on the lower summit (0.13 ml l1) but in only 27% and 15% of polychaete individuals on the better oxygenated ﬂank and base, respectively (Levin et al. 1994). In this setting the high frequency of brooding at the volcano summit may be an adaptation to the isolated, island-like conditions or to high food availability rather than to hypoxia.

Interannual and seasonal variability Strong evidence of the overriding importance of oxygen in controlling benthic communities within OMZs is provided by climatic events that alter oxygenation, for example, seasonal monsoons in the Arabian Sea, El Niño events in the eastern Paciﬁc, and upwelling events off West Africa. There is a monsoon-related shift in the position of the OMZ off the west coast of India, with the OMZ moving up the shelf during the SW monsoon (Banse 1984). This results in notably diminished catches of ﬁshes and prawns between June and September (Sankaranarayanan & Qasim 1968). During El Niño events, normally dysoxic waters on the shelf and upper slope of the Peru and Chile margin experience an infusion of warmer, higher-salinity, oxygenated water ﬂowing poleward from the Peru undercurrent. This infusion is greatest on the Peru margin and lessens to the south off Chile (Arntz et al. 1991). At a 34 m station in the Bay of Ancon, Peru that is normally hypoxic (0.5 ml l1), the 1982–3 El Niño caused a tripling of macrofaunal species diversity, increased biomass, increased density (32-fold), and the appearance of echinoderms, molluscs and crustaceans (previously absent) (Tarazona et al. 1988a,b). The nearly immediate response to increased oxygenation of bottom waters suggests the new colonists were highly opportunistic species (Tarazona et al. 1988b). In slightly deeper waters off Peru (40–110 m, from 5°S to 17°S), mean macrofaunal density and biomass did not increase during the 1982–3 El Niño, despite a dramatic rise in oxygen to 2–4 ml l1. Arntz et al. (1991) suggest there was increased predation pressure on deep-water assemblages during the 1982–3 El Niño, resulting from downward migration of benthic predators such as crabs and hake (to escape high temperatures in shallow water) and from a proliferation of other predators such as nemerteans. The normally dominant polychaetes, Paraprionospio pinnata and Magelona phyllisae, maintained dominance during this El Niño, while the reef building Sabellaria bella invaded at 60–80 m. In deeper waters (110–400 m) increased density and biomass were reported from 5°30S to 10°S (Arntz et al. 1991). On the Chile margin, Gallardo (1985) reported reduced biomass and density of macrobenthos during the 1983 El Niño, but an increased proportion of molluscs relative to other taxa. Detailed studies of infauna on the central Chile shelf during the 1997–8 El Niño revealed abrupt shifts in composition of both macrofauna (Gutiérrez et al. 2000) and meiofauna (Neira et al. 2001c) in response to increased oxygenation of overlying water. There was a reduction in biomass and reduced dominance of nematodes and spionid polychaetes, accompanied by increased representation of copepods (Neira et al. 2001c) and subsurface burrowing annelids (e.g. Cossura, Nephtys, Nereis) (D. Gutiérrez & V. A. Gallardo, unpubl.

27

LISA A. LEVIN

obs.). Effects of El Niño on the prokaryotic component of the benthos are also evident. The biomass of Thioploca spp. and Beggiatoa spp. ﬁlaments dropped or disappeared during El Niño events off central Chile in 1982–3 (Gallardo 1985, Arntz et al. 1991) and 1997–8 (Gutiérrez et al. 2000). El Niño events off Peru typically generate positive responses by megafauna such as the scallop Argopecten purpuratus, the shrimp Xiphopenaeus riveti, and the galatheid crab Pleuroncodes monodon, particularly in shallower waters. Post El Niño increases in Peruvian hake (Merluccius gayi peruanus) and barnacles (Pollicipes elegans) have also been observed (Arntz 1986, Tarazona et al. 1988a). December 1992 trawls off Peru recovered a number of subtropical species from the Panamanian province (north Peru and off Ecuador and Columbia) including shrimp and swimming crabs (Arntz et al. 1991). Many inshore demersal ﬁshes migrated to the edge of the Peru shelf, their diets became diversiﬁed, and catches of selected species on the shelf and upper slope generally increased during the 1983 El Niño (Arntz et al. 1988). Taxa related to those responding off Peru, including Callinectes arcuatus, Sicyona spp., Penaeus spp. and Pleuroncodes planipes (tuna crabs), appear or increase in abundance by several orders of magnitude on the shelf and upper slope off California during El Niño events (Thompson et al. 1993, Montagne & Cadien 2001). Episodic vertical mixing events in OMZ waters off Walvis Bay, West Africa are notable for release of noxious hydrogen sulphide gas into the surrounding atmosphere (Weeks et al. 2002). These events create anoxic waters that cause mass mortalities of ﬁshes, crustaceans and molluscs. These mortalities, in conjunction with redistribution and crowding of mobile species like rock lobsters, cause signiﬁcant economic impacts in the region (Bailey et al. 1985).

Ecosystem-level responses to OMZ conditions Bioturbation and animal lifestyles Animal activities such as bioturbation (the mixing of sediment particles by animals) and bioirrigation (advection of solutes by animals) enhance oxygenation and solute transport, speed the remineralisation of organic matter, and mix sediment horizons, obliterating laminae that might otherwise form. It is generally believed that these activities should be reduced when oxygen becomes limiting (e.g. Pearson & Rosenberg 1978, Rhoads et al. 1978), a paradigm adopted by palaeoecologists to generate ichnofacies and biofacies models for reconstruction of past oxygen and productivity regimes (Savrda & Bottjer 1991). The inﬂuence of OMZ-associated oxygen depletion on bioturbation has been examined with x-radiography (e.g. Smith et al. 2000), counts of tubes and burrows (e.g. Thompson et al. 1985; Meadows et al. 2000), analysis of animal lifestyles (Gutiérrez et al. 2000) and activities of radionuclides to estimate depths and rates of particle mixing (for example, Smith et al. 2000). Bioturbation is generally reduced within OMZs. Under extreme hypoxia or anoxia, all bioturbating organisms are absent and laminae or varves often form. These are well documented from some OMZs (Emeis et al. 1991) and silled basins with restricted circulation (Dean et al. 1994). However, these sediments are not inert biologically. Infaunal Foraminifera can burrow and could have a role in bioturbation (e.g. Gross 2000). There are dense assemblages of meiofauna (mainly nematodes) that produce extensive open burrow networks to 3 cm depth in the 28

OXYGEN MINIMUM ZONE BENTHOS

Santa Barbara Basin sediment column where O2 0.3 M to 2.5 M (Pike et al. 2001). These facilitate microbioirrigation, which signiﬁcantly increases solute transport across the sediment/water interface. Levin et al. (2002, L. Levin, unpubl. obs.) report the presence of high densities (16 000 ind. m2) of burrowing (bioturbating) gutless oligochaetes in a 6-cm deep homogenised layer (O2 levels of 0.02 ml l1) above laminated sediments, suggesting recent colonisation. Sampling occurred during maximum intensity of the 1997–8 El Niño when temporary oxygenation may have permitted faunal colonisation (Levin et al. 2002). Bioturbation of modern OMZ sediments has been evaluated on the California, Peru, Chile and Oman margins. For the central California margin, Thompson et al. (1985) reported maximum numbers of polychaete tubes at the upper OMZ boundary (500 m) with declines towards the OMZ core, few burrows in the upper or core OMZ, and increasing burrow numbers towards the lower OMZ boundary (900 m). Meadows et al. (2000) also documented reduced numbers of tubes and burrows (they did not distinguish them) at the core of the Oman OMZ (391–406 m) and highest values from 840–1265 m, encompassing higher oxygenation at the lower OMZ boundary. They found a positive linear relationship between burrow number and bottom-water oxygen near the sediment surface. Higher burrow density in the Oman margin OMZ was associated with decreased penetration resistance of the sediment, and decreased variability in these measures (Murray et al. 2000). Off Oman, modal burrow diameter and diversity of burrow types were positively correlated with oxygen concentration (Smith et al. 2000). On the central California margin, where laminations are evident only in subsurface parts of cores, all modern sediments are at least partially bioturbated. Thompson et al. (1985) state that environments having oxygen values between 0.3 ml l1 and 0.5 ml l1 contain abundant animal populations capable of completely homogenising bottom sediments and that preservation of laminations requires O2 0.1 ml l1. Calvert (1964) found laminated sediments only below 0.2 ml l1. Savrda et al. (1984) suggest bioturbation halts only below 0.15 ml l1. X-radiographs from the Oman margin suggest faint laminations present at the lowest oxygen levels (0.13 ml l1) but fairly extensive bioturbation at 0.16 ml l1 and higher (Smith et al. 2000). The mixed layer depth, determined with 210Pb and 234Th proﬁles, is much lower within than beneath the OMZ on the Oman (Smith et al. 2000) and Peru margins (Levin et al. 2002). The mean 210Pb mixed layer of 4.6 cm (100-yr timescale) within the Oman margin OMZ (0.1 ml l1 to 0.3 ml l1 O2) was half the mean measured on oxygenated slopes of the Paciﬁc and Atlantic Oceans (11.1 cm) (Smith et al. 2000) and much less than the global mean for bioturbated sediments reported by Boudreau (1994). Kim & Burnett (1988) also reported a thinning of 210Pb mixed layers at bathyal depths within the Peru margin OMZ. Particle mixing rates (D6), estimated by 234Th using non-local mixing models, are reduced within the Peru margin OMZ (14 cm2 yr1 at 0.26 ml l1 O2) relative to stations beneath the OMZ (80–100 cm2 yr1) (Levin et al. 2002). No reduction in 210Pb-based particle mixing rates (per unit volume) was observed within (0.15 cm2 yr1 to 2.9 cm2 yr1) compared with that beneath the Oman margin OMZ (1.1 0.62 cm2 yr1), but the edge station (0.52 ml l1 O2) had an unusually high value (40 cm2 yr1) (Smith et al. 2000). A longer-lived tracer (Pb-210) was used in this study than in the Peru study and it may be tracking lower quality organic matter (Smith et al. 1993) or may have different particle reactivities (Shull & Mayer 2002). The reduced mixed-layer depth without a change in overall mixing intensity suggests that the Oman OMZ sediments must experience only half the bioturbation energy of the oxygenated slope sediments (Smith et al. 2000). Temporal changes associated with El Niño-induced oxygenation provide further evidence of the regulatory potential of oxygen. 29

LISA A. LEVIN

On a normally dysoxic outer shelf off central Chile, Gutiérrez & Gallardo (submitted) documented substantial deepening of the mixed layer and a 3-fold increase in particle mixing (Db) during the 1997–8 El Niño, relative to conditions before and after. In general, bioturbation patterns appear to be driven by a shift from tube-building, interface-feeding taxa in the core of the OMZ to burrowing, subsurface-feeding species above or below the OMZ (Thompson et al. 1985, Gutiérrez et al. 2000, Levin et al. 2000), whereas changes in animal density and biomass have little effect on bioturbation patterns within OMZs (Smith et al. 2000, Levin et al. 2002). These same lifestyle shifts are mirrored on the central Chile shelf, when OMZ-associated hypoxia (0.3 ml l1) lessens during El Niño events (Gutiérrez et al. 2000). There, increased oxygenation (to 1 ml l1) was accompanied by a reduction in the quality of organic matter deposited, and the fauna shifted from surfacefeeding, tube-building spionids (Paraprionospio pinnata) to deep-dwelling, diffusive bioturbators (Gutiérrez et al. 2000, D. Gutiérrez & V. A. Gallardo, unpubl. obs.). Similar responses were observed on the central Peru shelf (D. Gutiérrez, pers. comm.).

Trophic pathways OMZ inhabitants reside beneath the most productive waters in the world. Thus, the fauna might be expected to rely exclusively on heterotrophic consumption of this production. However, recent ﬁndings suggest that chemosynthesis-based nutrition plays an important role in OMZ systems, either through symbiosis or through heterotrophic consumption of chemosynthetically ﬁxed carbon. Numerous OMZ heterotrophs consume free-living, chemosynthetic bacteria or prey on those species that do. Mats of ﬁlamentous sulphur bacteria offer an abundant food supply for species that can tolerate the sulphur, and Gallardo et al. (1994, 1995) have proposed that these bacteria may form the base of food chains on the Peru–Chile margin. Carbon and nitrogen isotopic signatures of macrofauna collected from the northern Chile margin suggest that these animals have chemosynthetically-derived food sources within the 200–300 m OMZ core (avg 13C 23.2, 15N 8.6). The 13C values for macrofauna in this part of the OMZ are 5–7‰ lighter and the 15N signatures are 8–10‰ lighter than macrofaunal signatures of animals from above the OMZ (100 m), near the OMZ edge (500 m), or below the OMZ (800 m). The OMZ macrofaunal signatures more closely resemble those of Thioploca from the region (avg 13C 22.9, 15N 10.6) than phytodetritus (avg 13C 18.1,

15N 13.8) (L. Levin et al., unpubl. data); however, several of the OMZ taxa (Olavius crassitunicatus and pogonophorans) are known to possess chemoautotrophic symbionts. The presence of endosymbiotic, sulphide-oxidising bacteria that ﬁx and translocate carbon to the host appears to be widespread in OMZs. Within the Santa Barbara Basin there are at least four Foraminifera, one euglenoid ﬂagellate, ﬁve ciliates, and one bivalve (Lucinoma aequizonata) that exhibit endobiotic bacteria (Cary et al. 1989, Bernhard et al. 2000). In addition, there are four euglenozoan ﬂagellates, 10 ciliates, a nematode (Desmodora masira) and a polychaete (Xenonerilla bactericola) with ectobiotic bacteria. For most of these taxa, the host-symbiont metabolic interactions have not been determined, but a nutritional role for the symbionts has been documented in Lucinoma aequizonata (Cary et al. 1989, Hentschel et al. 1993), and is likely in many of the other Santa Barbara Basin taxa (Bernhard et al. 2000). The gutless oligochaete, Olavius crassitunicatus, the dominant taxon at 300 m off Peru, possesses three or more types of subcuticular bacteria (Fig. 7), at least one 30

OXYGEN MINIMUM ZONE BENTHOS

Figure 7 Subcuticular, symbiotic bacteria in the gutless oligochaete, Olavius crassitunicatus (Tubiﬁcidae), Peru margin OMZ, depth 305 m. The large globular bacteria are sulphide oxidising proteobacteria. Also present are smaller, rod-shaped bacteria immediately under the cuticle, and elongate, ﬁlamentous, spirochaete-like forms. See Giere & Krieger (2001) for a detailed description.

of which oxidises sulphide (Giere & Krieger 2001). Recently, Dubilier et al. (2001) have shown that a congener, O. algarvensis, possesses sulphate-reducing and sulphide-oxidising bacteria in a conﬁguration similar to what is seen beneath the cuticle of O. crassitunicatus. Other examples of symbiont-bearing taxa within OMZs include pogonophorans on Volcano 7 (Levin et al. 1991a) and on the northern Chile margin (L. Levin et al., unpubl. obs.), and lucinid clams on the Oman margin (Levin et al. 2000), in the Gulf of California (L. Levin, unpubl. obs.) and in the Santa Barbara Basin (Cary et al. 1989). Further investigation will almost certainly lengthen this list and reveal complex interactions where multiple symbionts interact with each other and with their hosts. The full extent to which chemoautotrophic symbioses or free-living chemosynthetic bacteria provide nutrition to OMZ benthos remains unknown. Ward et al. (1989) suggest that chemoautotrophy in the water column off Peru, from the activities of nitrifying and denitrifying bacteria, could be a signiﬁcant source of organic matter to deep, low-oxygen waters. Gut analysis of copepods from the eastern Paciﬁc and Arabian Sea reveal extensive consumption of bacterial aggregates in the lower OMZ; these aggregates may reﬂect in situ chemoautotrophic production (Gowing & Wishner 1992, 1998, Wishner et al. 1995). A thorough investigation of food sources for OMZ benthos might include gut content analyses, evaluation of natural isotopic signatures and lipid biomarkers characteristic of different food sources, use of molecular sequencing tools to identify the presence and functions of symbionts, gut ﬂora and gut contents, tracer experiments to track the assimilation of 31

LISA A. LEVIN

experimentally labelled food items or precursors, and use of naturally occurring radioisotopes (14C and 234Th) to examine particle selectivity by deposit feeders.

Palaeoceanographic and evolutionary implications of OMZ benthos OMZ sediments are of great value to palaeoecologists. High accumulation rates and reduced bioturbation in microxic OMZs generate extraordinarily high-resolution historical records of planktonic and benthic organisms and their activities, allowing reconstruction of hydrographic and productivity regimes (e.g. Behl & Kennett 1996, Cannariato et al. 1999). The modern biota of dysoxic oxygen minimum zones and silled basins has also played a major role in shaping oxygen-based biofacies models. Palaeoecologists predict shifts in body fossils, trace fossils and sedimentary structures as a function of oxygen concentration (reviewed in Savrda et al. 1984, Savrda & Bottjer 1987, 1991, Rhoads et al. 1991). Key features associated with low oxygen include increasing dominance by softbodied annelids, reduced species richness and evenness, and reduced incidence of calcified species. Dwelling-habit shifts are more complicated (Ekdale & Mason 1988, 1989,Wheatcroft 1989a,b). Burrowing by oligochaetes and nematodes typically occurs under microxic conditions (0.1 ml l1), tube building by polychaetes is common from 0.1 ml l1 to 0.3 ml l1, and larger burrowers may appear at oxygen concentrations 0.3 ml l1. Other links between modern OMZ faunas and biofacies models are reviewed in Levin et al. (2000). A difficulty in relating modern and ancient seafloor OMZ assemblages is lack of understanding about which OMZ features are preserved in the historical sediment record, particularly because the modern sediments are often soupy, with high water content and low shear strength. The oxygen minimum zone is recognised as a key oceanographic boundary for animals intolerant of hypoxia. OMZs inﬂuence evolution of slope populations and species by (a) creating strong gradients in oxygen over relatively short distances and (b) by forming barriers that block gene ﬂow or reduce migrations, thereby isolating populations and promoting allopatric speciation (White 1987, Jacobs & Lindberg 1998, Rogers 2000). OMZ isolation effects are believed to act on both benthic and pelagic species. Maximum species richness and genetic diversity of benthos often occur at bathyal depths (500–2500 m) on continental margins (Haedrich et al. 1980, Rex 1981, R. Etter, pers. comm.), where OMZs may have played a role in shaping the faunas. The taxonomic composition of the abyssal faunas and differences between oceans may in fact reﬂect cumulative inﬂuence of past oxygen depletion or anoxia (Rogers 2000). Indirect effects of hypoxia on life-history traits and on favourable biochemical traits can also inﬂuence rates and direction of speciation. Creasey et al. (2000) observed reduced heterozygosity in the squat lobster Munidopsis scobina near the Oman margin OMZ. This might be related to loss of poorly adapted individuals or to lower effective population size. Spider crabs (Encephaloides armstrongi) also exhibited genetic differentiation (of AAT and PGI) between individuals above (juveniles) and within the OMZ (adults), possibly as a result of selection, but the interpretations are complicated (Creasey et al. 1997). Expansions and contractions of OMZs over geologic time are proposed to inﬂuence the rate and alter the location of species evolution in the ocean (White 1987, Jacobs & Lindberg 1998, Rogers 2000). During anoxic events, deep-sea and vent/seep species will die out or 32

OXYGEN MINIMUM ZONE BENTHOS

seek refuge in shallow water. During these periods, the shelves may become sites of faunal diversiﬁcation; shelf faunas may have recolonised deep water after the anoxic events in the Cenomanian/Turonian and late Palaeocene periods (Jacobs & Lindberg 1998). Where modern OMZ faunas ﬁt into this picture and whether they are ancient or relatively recently evolved, are issues that have not been addressed. Not only can modern OMZs provide clues about the past, but they may also tell us how shallow-water systems could change should they shift from episodic to permanent hypoxia. These vast, relatively unexplored ecosystems may preview the types of adaptations, species, and processes that will prevail with increasing hypoxia over ecological and evolutionary time.

OMZ frontiers The OMZ ecosystem represents one of the ocean’s oldest and most extreme environments. Studies of OMZ benthos to date have been hampered largely by limited access to deep-water systems in remote parts of the world. The OMZ in the Bay of Bengal, for example, remains largely unexplored. As scientists begin to understand the importance of these ecosystems for nutrient cycling, as incubators of evolutionary novelty, and even for ﬁsheries production, our knowledge of OMZs should increase tremendously. Below are unanswered questions that reﬂect major gaps in our understanding of seaﬂoor OMZ ecosystems. (1)

(2) (3) (4)

(5)

What really controls standing stock in OMZs? Do oxygen and organic matter interact in determining abundances? What is the role of sulphides? Food supply cannot be the sole determinant of community structure because the most organic-rich sediments in the world support small-bodied organisms with low biomass. What supports dense animal aggregation at OMZ boundaries? What are the physiological, enzymatic, metabolic, reproductive and molecular adaptations of benthic animals to OMZs? Much of the OMZ research on these topics to date has focused on plankton and nekton. What are the ecosystem-level manifestations of extreme oxygen depletion? Do carbon cycling pathways differ? Is there a shift to chemosynthesis-based nutrition? What are the functional consequences of low species diversity in OMZs? OMZ assemblages may exhibit some of the lowest metazoan species diversities on earth, as well as some of the sharpest diversity gradients. How does low animal diversity inﬂuence microbial production, remineralisation processes and the nature of species interactions? What evolutionary constraints and adaptations are associated with permanent hypoxia that are not evident under conditions of seasonal or episodic hypoxia? Bathyal animals within OMZs clearly tolerate much lower oxygen concentrations than those in shallow, temporally variable settings. Many faunal shifts in community structure occur when oxygen falls below 0.45 ml l1 in OMZs, (e.g. Rosenberg et al. 1983, Levin & Gage 1998), but under conditions of seasonal hypoxia in shallow water, these shifts occur at 2 ml l1 (⬃3 mg l1) in motile species and 1.0–0.7 ml l1 (1.5–1.0 mg l1) in sessile taxa (Rabalais et al. 2001, Gray et al. 2002, Karlson et al. 2002). What is the basis for this difference? 33

LISA A. LEVIN

To answer the questions above, the next generation of seaﬂoor OMZ investigations must integrate biogeochemical, microbiological, molecular and ecological tools. New instrumentation, the use of biomarkers, and the potential for in situ, shipboard and laboratory experimentation under controlled oxygen conditions offer exciting means of advancing our mechanistic understanding of ecological processes in OMZs. This improved understanding of the extensive areas impacted by modern OMZs will substantially enhance our ability to interpret the nature of palaeoceanographic and palaeoclimatic changes from the sediment record and to predict the impacts of global climate change and anthropogenic activities on continental margins. Our knowledge of the OMZ ecosystems is truly in its infancy, and many surprises surely lie ahead.

Acknowledgements I dedicate this paper to the memory of Howard Sanders, whose pioneering work in Walvis Bay during the 1960s woke the scientiﬁc community to the considerable ecological impact of oxygen minima in the world ocean. The author’s oxygen minimum zone research has been supported by grants from the Ofﬁce of Naval Research, the National Science Foundation INT94-14397, INT02-27511, OCE 98-03861, NATO, and the University of California Ship Funds. Collaboration and discussions with Joan Bernhard, Adam Cook, John Gage, Victor Gallardo, Dimitri Gutiérrez, Andrew Gooday, Peter Lamont, Arnold Mantyla, Lauren Mullineaux, Carlos Neira, Graham Oliver, Anthony Rathburn, Alex Rogers, Javier Sellanes, Craig Smith, Paul Tyler and Karen Wishner have been particularly helpful. I thank Anne Beaudreau, Leslie Harris, Joshua Hillman, Cindy Huggett, David James, Larry Lovell, Chris Martin, Guillermo Mendoza, and Katherine Wilson for assistance in the laboratory and with identiﬁcations, John Helly for providing hydrocast data, and J. Gonzalez and S. Challeripe for help with manuscript preparation. Special thanks are extended to A. Gooday, C. Neira, A. Rathburn, J. Atkinson, and an anonymous external reviewer for helpful comments on an earlier draft of this paper.

References Adams, P. B., Butler, J. L., Baxter, C. H., Laidig, T. E., Dahlin, K. A. & Wakeﬁeld, W. W. 1995. Population estimates of Paciﬁc coast groundﬁshes from video transects and swept-area trawls. Fishery Bulletin 93, 446–455. Alve, E. & Bernhard, J. M. 1995. Vertical migratory response of benthic foraminifera to controlled oxygen concentrations in an experimental mesocosm. Marine Ecology Progress Series 116, 137–151. Anderson, G. C. 1972. Double oxygen minimum in the southeastern Paciﬁc Ocean. Sears Foundation: Journal of Marine Research 30, 275–280. Arntz, W. E. 1986. The two faces of El Nino 1982–83. Commercially exploited invertebrates (mariscos) of the Peruvian shore. Meeresforschungen 31, 1–46. Arntz, W. E., Tarazona, J., Gallardo, V., Flores, L. & Salzwedel, H. 1991. Benthos communities in oxygen deﬁcient shelf and upper slope areas of the Peruvian and Chilean Paciﬁc coast, and changes caused by El Niño. In Modern and ancient continental shelf anoxia, R. V. Tyson & T. H. Pearson (eds). London: Geological Society Special Publications No. 58, 131–154.

34

OXYGEN MINIMUM ZONE BENTHOS

Arntz, W. E., Valdivia E. & Zeballos, J. 1988. Impact of El Niño 1982–83 on the commercially exploited invertebrates (mariscos) of the Peruvian shore. Meeresforschungen 32, 3–22. Baas, J. H., Schönfeld, J. & Zahn, R. 1998. Mid-depth oxygen drawdown during Heinrich events: evidence from benthic foraminiferal community structure, trace-fossil tiering, and benthic 13C at the Portuguese Margin. Marine Geology 152, 25–55. Bailey, G. W., Beyers, C. J. De. B. & Lipschitz, S. R. 1985. Seasonal variations of oxygen deﬁciency in the waters off south west Africa in 1975 and 1976 and its relation to the catchability and distribution of the cape rock lobster Jasus lalandii. South African Journal of Marine Science 3, 197–214. Banse, K. 1984. Overview of the hydrography and associated biological phenomena in the Arabian Sea off Pakistan. In Marine geology and oceanography of Arabian Sea and coastal Pakistan, B. U. Hag & J. D. Milligan (eds). New York: Van Nostrand Reinhold, 271–304. Beaudreau, A. 1999. Macrofaunal community responses to hypoxia in the Santa Barbara Basin. UCSD SURF Program undergraduate report, 44 pp. Beaulieu, S. E. 2002. Accumulation and fate of phytodetritus on the sea ﬂoor. Oceanography and Marine Biology: an Annual Review 40, 171–232. Behl, R. J. & Kennett, J. P. 1996. Brief interstadial events in the Santa Barbara Basin, NE Paciﬁc, during the past 60 kyr. Nature 379, 243–246. Bernhard, J. M. 1992. Benthic foraminiferal distribution and biomass related to pore-water oxygen content: central California continental slope and rise. Deep-Sea Research 39, 585–605. Bernhard, J. M. 1996. Microaerophilic and facultative anaerobic benthic foraminifera: a review of experimental and ultrastructural evidence. Revue de Paléobiologie 15, 261–275. Bernhard, J. M. & Alve, E. 1996. Survival, ATP pool, and ultrastructural characterization of foraminifera from Drammensfjord (Norway): response to anoxia. Marine Micropaleontology 28, 5–17. Bernhard, J. M. & Bowser, S. S. 1999. Benthic foraminifera of dysoxic sediments: chloroplast sequestration and functional morphology. Earth-Science Reviews 46, 149–165. Bernhard, J. M., Buck, K. R., Farmer, M. A. & Bowser, S. S. 2000. The Santa Barbara Basin is a symbiosis oasis. Nature 403, 77–80. Bernhard, J. & Reimers, C. 1991. Benthic foraminiferal population ﬂuctuations related to anoxia: Santa Barbara Basin. Biogeochemistry 15, 127–149. Bernhard, J. & Sen Gupta, B. K. 1999. Foraminifera of oxygen depleted environments. In Modern Foraminifera, B. K. Sen Gupta (ed.). Dordrecht: Kluwer, 201–216. Bernhard, J. M., Sen Gupta, B. K. & Borne, P. F. 1997. Benthic foraminiferal proxy to estimate dysoxic bottom-water concentrations, Santa Barbara Basin, U.S. Paciﬁc continental margin. Journal of Foraminiferal Research 27, 301–310. Bianchi, T. S., Johansson, B. & Elmgren, R. 2000. Breakdown of phytoplankton pigments in Baltic sediments: effects of anoxia and loss of deposit-feeding macrofauna. Journal of Experimental Marine Biology and Ecology 251, 161–183. Boudreau, B. P. 1994. Is burial velocity a master parameter for bioturbation? Geochimica et Cosmochimica Acta 58, 1243–1249. Bradshaw, J. S. 1961. Laboratory experiments of the ecology of foraminifera. Contributions from the Cushman Foundation for Foraminiferal Research 12, 31–71. Calvert, S. E. 1964. Factors affecting distribution of laminated diatomaceous sediments in the Gulf of California. In Marine geology of the Gulf of California, T. H. Van Andel & G. G. Shor (eds). Memoirs of the American Association of Petrologists and Geologists. Tulsa, Oklahoma: American Association of Petroleum Geologists, 311–330. Calvert, S. E. 1987. Oceanographic controls on the accumulation of organic matter in marine sediments. In Marine petroleum source rock, J. Brooks & A. J. Fleet (eds). Geological Society Special Publication No. 26. London: Geological Society 137–151. Calvert, S. E., Bustin R. M. & Pederson, T. F. 1992. Lack of evidence for enhanced preservation of sedimentary organic matter in the oxygen minimum of the Gulf of California. Geology 20, 757–760.

35

LISA A. LEVIN

Cannariato, K. G. & Kennett, J. P. 1999. Climatically related millennial-scale ﬂuctuations in strength of California margin oxygen minimum zone during the past 60 k.y. Geology 27, 975–978. Cannariato, K. G., Kennett, J. P. & Behl, R. J. 1999. Biotic response to late Quaternary rapid climate switches in Santa Barbara Basin: ecological and evolutionary implications. Geology 27, 63–66. Cary, S. C., Vetter, R. D. & Felbeck, H. 1989. Habitat characterization and nutritional strategies of the endosymbiont-bearing bivalve Lucinoma aequizonata. Marine Ecology Progress Series 55, 31–45. Cedhagen, T. 1991. Retention of chloroplasts and bathymetric distribution in the sublittoral foraminiferan Nonionella labradorica. Ophelia 33, 17–30. Chapelle, G. & Peck, L. S. 1999. Polar gigantism dictated by oxygen availability. Nature 399, 114–115. Childress, J. J. 1975. The respiratory rates of midwater crustaceans as a function of depth of occurrence and relation to oxygen minimum layer off southern California. Comparative Biochemistry and Physiology 50, 787–799. Childress, J. J. & Seibel, B. A.1998. Life at stable low oxygen levels: adaptations of animals to oceanic oxygen minimum layers. Journal of Experimental Biology 201, 1223–1232. Codispoti, L. A. & Christiansen, J. P. 1989. Nitriﬁcation, denitriﬁcation and nitrous oxide cycling in the eastern tropical South Paciﬁc Ocean. Marine Chemistry 16, 277–300. Cook, A. A., Lambshead, P. J., Hawkins, L. E., Mitchell, N. & Levin, L. A. 2000. Nematode abundance at the oxygen minimum zone in the Arabian Sea. Deep-Sea Research, Part II 47, 75–85. Corliss, B. H. & Chen, C. 1988. Morphotype patterns of Norwegian Sea deep-sea benthic foraminifera and ecological implications. Geology 16, 716–719. Cowie, G. L., Calvert, S. E., Pedersen, T. F., Schulz, H. & von Rad, U. 1999. Organic content and preservational controls in surﬁcial shelf and slope sediments from the Arabian Sea (Pakistan margin). Marine Geology 161, 23–38. Creasey, S. S. 1998. Population genetics and biology of deep-sea Crustacea. PhD thesis, University of Southampton, United Kingdom. Creasey, S., Rogers, A., Tyler, P., Gage, J. & Jollivet, D. 2000. Genetic and morphometric comparisons of squat lobster, Munidopsis scobina (Decapoda: Anomura: Galatheidae) populations, with notes on the phylogeny of the genus Munidopsis. Deep-Sea Research, Part II 47, 87–118. Creasey, S., Rogers, A. D., Tyler, P., Young, C. & Gage, J. 1997. The population biology and genetics of the deep-sea spider crab, Encephaloides armstrongi Wood-Mason 1891 (Decapoda: Majidae). Philosophical Transactions Royal Society London B 352, 365–379. Danovaro, R., Dell’Anno, A. & Fabiano, M. 2001. Bioavailability of organic matter in the sediments of the Porcupine Abyssal Plain, northeastern Atlantic. Marine Ecology Progress Series 220, 25–32. Dean, W. E., Gardner, J. V. & Anderson, R. Y. 1994. Geochemical evidence for enhanced preservation of organic matter in the oxygen minimum zone of the continental margin of northern California during the late Pleistocene. Paleoceanography 9, 47–61. Demaison, G. J. & Moore, G. T. 1980. Anoxic environments and oil source bed genesis. American Association of Petrologists and Geologists Bulletin 64, 1179–1209. den Dulk, M., Reichart, G. J., Memon, G. M., Roelofs, E. M. B., Zachariasse, W. J. & van der Zwaan, G. J. 1998. Benthic foraminiferal response to variations in surface water productivity and oxygenation in the northern Arabian Sea. Marine Micropaleontology 35, 46–66. De Robertis, A., Eiane, M. K. & Rau, G. 2001. Eat and run: anoxic feeding and subsequent aerobic recovery by Orchomene obtusus in Saanich Inlet, British Columbia, Canada. Marine Ecology Progress Series 219, 221–227. Devol, A. H. & Hartnett, H. E. 2001. Role of the oxygen-deﬁcient zone in transfer of organic carbon to the deep ocean. Limnology and Oceanography 46, 1684–1690. Diaz, R. J. & Rosenberg, R. 1995. Marine benthic hypoxia, A review of its ecological effects and the behavioural responses of benthic macrofauna. Oceanography and Marine Biology: an Annual Review 33, 245–303. Diaz, R. J. & Rosenberg, R. 2001. Overview of anthropogenically-induced hypoxic effects on marine

36

OXYGEN MINIMUM ZONE BENTHOS

benthic fauna. In Coastal hypoxia: consequences for living resources and ecosystems, N. Rabalais & R. Turner (eds). Washington DC: American Geophysical Union, 129–146. Ditmar, W. 1884. Report into the composition of seawater. HMS Challenger Reports 1 (section 2), 1–251. Douglas, R. G. 1981. Paleoecology of continental margin basins: a modern case history from the borderland of southern California. In Depositional systems of active continental margin basins, R. G. Douglas et al. (eds). SEPM Paciﬁc Section Short Course Notes. Paciﬁc Section, Society of Economic Paleontologists and Mineralogists, Los Angeles, 121–156. Dubilier, N., Mulders, C., Ferdelman, T., de Beer, D., Pernthaler, A., Klein, M., Wagner, M., Erseus, C., Thiermans, F., Krieger J., Giere, O. & Amann, R. 2001. Endosymbiotic sulphate-reducing and sulﬁde-oxidizing bacteria in an oligochaete worm. Nature 411, 298–302. Ekdale A. A. & Mason, T. R. 1988. Characteristic trace-fossil associations in oxygen-poor sedimentary environments. Geology 16, 720–723. Ekdale, A. A. & Mason, T. R. 1989. Commentary and reply on “Characteristic trace-fossil associations in oxygen-poor sedimentary environments”. Geology 17, 674–676. Emeis, K.-C., Whelan, J. K. & Tarafa, M. 1991. Sedimentary and geochemical expressions of oxic and anoxic conditions on the Peru shelf. In Modern and ancient continental shelf anoxia, R. V. Tyson & T. H. Pearson (eds). Geological Society Special Publication No. 58. London: Geological Society, 155–170. Fenchel, T. & Finlay, B. J. 1995. Ecology and evolution in anoxic worlds. New York: Oxford University Press. Finogenova, N. 1986. Six new species of marine Tubiﬁcidae (Oligochaeta) from the continental shelf off Peru. Zoologica Scripta 15, 45–51. Fontanier, C., Jorissen, F. J., Licari, L., Alexandre, A., Anschutz, P. & Carbonel, P. 2002. Live benthic foraminiferal faunas from the Bay of Biscay: faunal density, composition, and microhabitats. DeepSea Research 49, 751–785. Fossing, H., Gallardo, V. A., Jørgensen, B. B., Hüttel, M., Nielsen, L. P., Schulz, H., Canﬁeld, D. E., Forster, S., Glud, R. N., Gundersen, J. K., Küver, J., Ramsing, N. B., Teske, A., Thamdrup, B. & Ulloa, O. 1995. Concentration and transport of nitrate by the mat-forming sulphur bacterium Thioploca. Nature 374, 713–715. Frankenberg, D. & Menzies, R. J. 1968. Some quantitative analysis of deep-sea benthos off Peru. Deep-Sea Research 15, 623–626. Gage, J. D., Levin, L. A. & Wolff, G. A. 2000. Benthic processes in the deep Arabian Sea. Deep-Sea Research, Part II 47, 1–7. Gallardo, V. A. 1963. Notas sobre la densidad de la fauna bentonica en el sublitoral del norte de Chile. Gayana (Zool.) 10, 3–15. Gallardo, V. A. 1977. Large benthic microbial communities in sulﬁde biota under Peru–Chile subsurface countercurrent. Nature 268, 331–332. Gallardo, V. A. 1985. Efectos del fenómeno de El Niño sobre el bentos sublitoral frente a Concepción, Chile. In El Niño y su impacto en la fauna marina, W. Arntz et al. (eds). Callao: Inst. Mar Perú – Callao (Boletín Extraordinario), 79–85. Gallardo, V. A., Canete, J. I., Roa, R., Enriquez-Briones, S. & Baltazar, M. 1994. Recruitment of the squat lobster Pleuroncodes monodon on the continental shelf off Central Chile. Journal of Crustacean Biology 14, 665–669. Gallardo, V. A., Carrasco, F. D., Roa, R. & Cañete, J. I. 1995. Ecological patterns in the benthic macrobiota across the continental shelf off Central Chile. Ophelia 40, 167–188. Gallardo, V. A., Carrasco, F. D., Roa, R., Quiñones, R., Ulloa, O., Cañete, J. I., Mesías, J., Sobarzo, M., Pineda, V. & Baltazar, M. 1996. Benthic habitat detoxiﬁcation by a bacterial biomat (Thioploca spp.)? Observations on the recruitment of the squat lobster (Pleuroncodes monodon) off central Chile. In 9th Southern African Marine Science Symposium, Marine science in southern Africa. Past perspectives and future challenge. South Africa: University of Cape Town, 165 only.

37

LISA A. LEVIN

Gallardo, V. A., Klingelhoeffer, E., Arntz, W. & Graco, M. 1998. First report of the bacterium Thioploca in the Benguela ecosystem off Namibia. Journal of the Marine Biological Association of the United Kingdom 78, 1007–1010. Giere, O. 1993. Meiobenthology, Berlin: Springer-Verlag. Giere, O. & Krieger, J. 2001. A triple bacterial endosymbiosis in a gutless oligochaete (Annelida). Ultrastructural and immunocytochemical evidence. Invertebrate Biology 120, 41–49. Glenn, C. R., Arthur, M. A., Resig, J. M., Burnett, W. C., Dean, W. E. & Jahnke, R. A. 1993. Are modern and ancient phosphorites really so different? In Siliceous, phosphatic and glauconitic sediments of the Tertiary and Mesozoic, B. A. Iijima & R. E. Garrison (eds). Zeist, Netherlands: VSP International Science, 159–188. Gonzalez, R. R. & Quiñones, R. A. 2000. Pyruvate oxidoreductases involved in glycolytic anaerobic metabolism of polychaetes from the continental shelf off central-south Chile. Estuarine Coastal and Shelf Science 51, 507–519. Gooday, A. J., Bernhard, J. M., Levin, L. A. & Suhr, S. 2000. Foraminifera in the Arabian Sea OMZ and other oxygen deﬁcient settings: taxonomic composition, diversity and relation to metazoan faunas. Deep-Sea Research, Part II 47, 25–54. Gooday, A. J. & Rathburn, A. E. 1999. Temporal variability in living deep-sea benthic foraminifera: a review. Earth-Science Reviews 46, 187–212. Gordon, M. S. 1960. Anaerobiosis in marine sandy beaches. Science 132, 616–617. Gowing, M. M. & Wishner, K. F. 1992. Feeding ecology of benthopelagic zooplankton on an eastern tropical Paciﬁc seamount. Marine Biology 112, 451–467. Gowing, M. M. & Wishner, K. F. 1998. Feeding ecology of the copepod Lucicutia aff. L. grandis near the lower interface of the Arabian Sea oxygen minimum zone. Deep-Sea Research 45, 2433–2459. Graham, J. B., Dudley, R., Aguilar, N. M. & Gans, C. 1995. Implications of the late Palaeozoic oxygen pulse for physiology and evolution. Nature 375, 117–120. Gray, J. S., Wu, R. S. & Or, Y. Y. 2002. Effects of hypoxia and organic enrichment on the coastal marine environment. Marine Ecology Progress Series 238, 249–279. Grieshaber, M. K., Hardewig, I., Kreutzer, U. & Portner, H. O. 1994. Physiological and metabolic responses to hypoxia in invertebrates. Reviews of Physiology, Biochemistry, and Pharmacology 125, 43–147. Grieshaber, M. K. & Volkel, S. 1998. Animal adaptations for tolerance and exploitation of poisonous sulﬁde. Annual Review of Physiology 60, 33–53. Gross, O. 2000. Inﬂuence of temperature, oxygen and food availability on the migrational activity of bathyal benthic foraminifera: evidence by microcosm experiments. Hydrobiologia 426, 123–137. Gutiérrez, D., Gallardo, V. A., Mayor, S., Neira, C., Vásquez, C., Sellanes, J., Rivas, M., Soto, A., Carrasco, F. & Baltazar, M. 2000. Effects of dissolved oxygen and fresh organic matter on the bioturbation potential of macrofauna in sublittoral bottoms off central Chile, during the 1997–98 El Niño. Marine Ecology Progress Series 202, 81–99. Haake, B., Ittekkott, V., Rixen, T., Ramaswamy, V., Nair, R. R. & Curry, W. B. 1993. Seasonality and interannual variability of ﬂuxes to the deep Arabian Sea. Deep-Sea Research 40, 1323–1344. Haedrich, R. L., Rowe, G. T. & Polloni, P. T. 1980. The megabenthic fauna in the deep sea south of New England. Marine Biology 57, 165–179. Hart, T. J. & Currie, R. I. 1960. The Benguela current. Discovery Report 31, 123–298. Harvey, H. R., Fallon, R. D. & Patton, J. S. 1986. The effect of organic matter and oxygen on the degradation of bacterial membrane lipids in marine sediments. Geochemica Cosmochimica Acta 50, 795–804. Hentschel, U., Cary, S. & Felbeck, H. 1993. Nitrate respiration in chemoautotrophic symbionts of the bivalve Lucinoma aequizonata. Marine Ecology Progress Series 94, 35–41. Herguera, J. C. & Berger, W. H. 1991. Paleoproductivity from benthic foraminifera abundance: glacial to post-glacial change in west-equatorial Paciﬁc. Geology 19, 1173–1176. Hinrichs, K. U., Hayes, J. M., Sylva, S. P., Brewer, P. G. & DeLong, E. F. 1999. Methane-consuming archaebacteria in marine sediments. Nature 398, 802–805.

38

OXYGEN MINIMUM ZONE BENTHOS

Hinrichs, K. U., Summons, R. E., Orphan, V. J. H., Sylva, S. P. & Hayes, J. M. 2000. Molecular and isotopic analysis of anaerobic methane-oxidizing communities in marine sediments. Organic Geochemistry 31, 1685–1701. Hunter, J. R., Butler, J. L., Kimbrell, C. & Lynn, E. A. 1990. Bathymetric patterns in size, age, sexual maturity, water content and caloric density of Dover sole, Microstomus paciﬁcus. CalCOFI Report 31, 132–144. Hyland, J., Baptiste, E., Campbell, J., Kennedy, J., Kropp, R. & Williams, S. 1991. Macroinfaunal communities of the Santa Maria Basin on the California outer continental shelf and slope. Marine Ecology Progress Series 78, 147–161. Jacobs, D. K. & Lindberg, D. R. 1998. Oxygen and evolutionary patterns in the sea: Onshore/offshore trends and recent recruitment of deep-sea faunas. Proceedings of the National Academy of Sciences USA 95, 9396–9401. Jacobson, L. D. & Hunter, J. R. 1993. Bathymetric demography and movement of Dover sole. North American Journal of Fisheries Management 13, 405–420. Jacobson, L. D. & Vetter, R. D. 1996. Bathymetric demography and niche separation of thornyhead rockﬁsh: Sebastolobus alascanus and Sebastolobus altivelis. Canadian Journal of Fisheries and Aquatic Sciences 53, 600–609. Jørgensen, B. B. & Gallardo, V. A. 1999. Thioploca spp: ﬁlamentous sulfur bacteria with nitrate vacuoles. FEMS Microbiology Ecology 28, 301–313. Jorissen, F. 1999a. Benthic foraminiferal microhabitats below the sediment-water interface. In Modern Foraminifera, B. K. Sen Gupta (ed.). Dordrecht: Kluwer Academic Publishers, 161–179. Jorissen, F. 1999b. Benthic foraminiferal successions across late Quaternary sapropels. Marine Geology 153, 91–101. Jorissen, F. J., de Stigter, H. C. & Widmark, J. G. V. 1995. A conceptual model explaining benthic foraminiferal microhabitats. Marine Micropaleontology 26, 3–15. Jumars, P. A. 1974. Dispersion patterns and species diversity of macrobenthos in two bathyal communities. PhD thesis, University of California, San Diego. Jumars, P. A. & Eckman, J. A. 1983. Spatial structure within deep-sea benthic communities. In The sea., G. T. Rowe (ed.). New York: John Wiley, 399–451. Jumars, P. A. & Gallagher, E. D. 1982. Deep-sea community structure: three plays on a benthic proscenium. In The environment of the deep sea, W. G. Ernst & J. G. Morin (eds). New Jersey: Prentice-Hall, 217–255. Kaiho, K. 1994. Benthic foraminiferal dissolved-oxygen index and dissolved-oxygen levels in the modern ocean. Geology 22, 719–722. Kaiho, K. 1999. Effect of organic carbon ﬂux and dissolved oxygen on the benthic foraminiferal oxygen index (BFOI). Marine Micropaleontology 37, 67–76. Kamykowski, D. & Zentara, S. J. 1990. Hypoxia in the world ocean as recorded in the historical data set. Deep-Sea Research 37, 1861–1874. Karl, D. M. & Knauer, G. A. 1984 Vertical distribution, transport and exchange of carbon in the northeast Paciﬁc Ocean. Evidence for multiple zones of biological activity. Deep-Sea Research 31, 221–243. Karl, D. M., Knauer, G. A., Martin, J. H. & Ward, B. B. 1984. Bacterial chemolithotrophy in the ocean is associated with sinking particles. Nature 309, 54–56. Karlson, K., Rosenberg, R. & Bonsdorf, E. 2002. Temporal and spatial large-scale effects of eutrophication and oxygen deﬁciency on benthic fauna in Scandinavian and Baltic waters – a review. Oceanography and Marine Biology: an Annual Review 40, 427–489. Keil, R.G. & Cowie, G.L. 1999. Organic matter preservation through the oxygen-deﬁcient zone of the NE Arabian Sea as discerned by organic carbon: mineral surface area ratios. Marine Geology 161, 13–22. Kemp, P. F. 1990. The fate of bacterial production. Reviews in Aquatic Sciences 2, 109–124. Kim, K. H. & Burnett, W. 1988. Accumulation and biological mixing of Peru margin sediments. Marine Geology 80, 181–194.

39

LISA A. LEVIN

Koutsoukos, E. A. M., Leary, P. N. & Hart, M. B. 1990. Latest Cenomanian–earliest Turonian lowoxygen tolerant benthonic foraminifera: a case study from the Sergipe basin (N.E. Brazil) and the western Anglo-Paris basin (southern England). Palaeogeography, Palaeoclimatology, Palaeoecology 77, 145–177. Kuypers, M. M. M., Blokker, P., Erbacher, J., Kinkel, H., Pancost., R. D., Schouten, S. & Damste, J. S. S. 2001. Massive expansion of marine archaea during a mid-Cretaceous oceanic anoxic event. Science 293, 92–94. Lamont, P. A. & Gage, J. D. 2000. Morphological responses of macrobenthic polychaetes to low oxygen on the Oman continental slope, NW Arabian Sea. Deep-Sea Research, Part II 47, 9–24. Lee, C. 1992. Controls on carbon preservation: the use of stratiﬁed water bodies to compare intrinsic rates of decomposition in oxic and anoxic systems. Geochimica Cosmochimica Acta 56, 3323–3335. Leutenegger, S. & Hansen, H. J. 1979. Ultrastructural and radiotracer studies of pore function in foraminifera. Marine Biology 54, 11–16. Levin, L. A., Childers, S. E. & Smith, C. R. 1991b. Epibenthic, agglutinating foraminiferans in the Santa Catalina Basin and their response to disturbance. Deep-Sea Research 38, 465–483. Levin, L. A. & Edesa, S. 1997. The ecology of cirratulid mudballs on the Oman margin, northwest Arabian Sea. Marine Biology 128, 671–678. Levin, L. A., Etter, R. J., Rex, M. A., Gooday, A. J., Smith, C. R., Pineda, J., Stuart, C. T., Hessler, R. R. & Pawson, D. 2001. Environmental inﬂuences on regional deep-sea species diversity. Annual Review of Ecology and Systematics 132, 51–93. Levin, L. A. & Gage, J. D. 1998. Relationships between oxygen, organic matter and the diversity of bathyal macrofauna. Deep-Sea Research 45, 129–163. Levin, L. A., Gage, J., Lamont, P., Cammidge, L., Martin, C., Patience, A. & Crooks, J. 1997. Infaunal community structure in a low-oxygen, organic rich habitat on the Oman continental slope, NW Arabian Sea. In Responses of marine organisms to their environments, L. Hawkins & S. Hutchinson (eds). Proceedings of the 30th European Marine Biology Symposium, United Kingdom: University of Southampton, 223–230. Levin, L. A., Gage, J. D., Martin, C. & Lamont, P. A. 2000. Macrobenthic community structure within and beneath the oxygen minimum zone, NW Arabian Sea. Deep-Sea Research, Part II 47, 189–226. Levin L. A., Plaia, G. R. & Huggett, C. L. 1994. The inﬂuence of natural organic enhancement on life histories and community structure of bathyal polychaetes. In Reproduction, larval biology, and recruitment of the deep-sea benthos, C. M. Young & K. J. Eckelbarger (eds). Columbia, SC: Columbia University Press, 261–283. Levin, L. A., Rathburn, A. E., Neira, C., Sellanes, J., Munoz, P., Gallardo, V. & Salamanca, M. 2002. Benthic processes on the Peru margin: A transect across the oxygen minimum zone during the 1997–1998 El Niño. Progress in Oceanography 53, 1–27. Levin, L. A., Thomas, C. L. & Wishner, K. 1991a. Control of deep-sea benthic community structure by oxygen and organic-matter gradients in the eastern Paciﬁc Ocean. Journal of Marine Research 49, 763–800. Linke, P. & Lutze, G. F. 1993. Microhabitat preferences of benthic foraminifera – a static concept or a dynamic adaptation to optimise food acquisition? Marine Micropaleontology 20, 215–234. Livingstone, D. R. 1983. Invertebrate and vertebrate pathway of anaerobic metabolism: evolutionary considerations. Journal of the Geological Society (London) 140, 27–37. Loubere, P. & Fariduddin, M. 1999. Benthic foraminifera and the ﬂux of organic carbon to the seabed. In Modern Foraminifera, B. K. Sen Gupta (ed.). Dordrecht: Kluwer Academic Publishers, 181–199. Manheim, F., Rowe, G. T. & Jipa, D. 1975. Marine phosphorite formation off Peru. Journal of Sedimentary Petrology 45, 243–251. Martin, J. H., Knauer, G. A., Karl, D. M. & Broenkow, W. W. 1987. VERTEX: carbon cycling in the northeast Paciﬁc. Deep-Sea Research 34, 267–285.

40

OXYGEN MINIMUM ZONE BENTHOS

McClain, C. R. & Rex, M. A. 2001. The relationship between dissolved oxygen concentration and maximum size in deep-sea turrid gastropods: an application of quantile regression. Marine Biology 13, 681–685. McHatton, S. C., Barry, J. P., Jannasch, H. W. & Nelson, D. C. 1996. High nitrate concentrations in vacuolated, autotrophic marine Beggiatoa spp. Applied Environmental Microbiology 62, 954–958. Meadows, A., Meadows, P. S., West, F. J. C. & Murray, J. M. H. 2000. Bioturbation, geochemistry and geotechnics of sediments affected by the oxygen minimum zone on the Oman continental slope and abyssal plain, Arabian Sea. Deep-Sea Research 47, 259–280. Montagne, D. E. & Cadien, D. B. 2001. Northern range extensions into the southern California Bight of ten decapod Crustacea related to the 1991/2 and 1997/98 El Niño events. Bulletin of the Southern California Academy of Sciences 100, 199–211. Moodley, L. & Hess, C. 1992. Tolerance of infaunal benthic foraminifera for low and high oxygen concentrations. Biological Bulletin 183, 94–98. Moodley, L., Schaub, B. E. M., van der Zwaan, G. J. & Herman, P. M. J. 1998a. Tolerance of benthic foraminifera (Protista: Sarcodina) to hydrogen sulphide. Marine Ecology Progress Series 169, 77–86. Moodley, L., van der Zwaan, G. J., Herman, P. M. J., Kempers, L. & van Breugel, P. 1997. Differential response of benthic meiofauna to anoxia with special reference to the Foraminifera (Protista: Sarcodina). Marine Ecology Progress Series 158, 151–163. Moodley, L., van der Zwaan, G. J., Rutten, G. M. W. Boom, R. C. E. & Kempers, A. J. 1998b. Subsurface activity of benthic foraminifera in relation to porewater oxygen content: laboratory experiments. Marine Micropaleontology 34, 91–106. Morrison, J. M., Codispoti, L. A., Smith, S. L., Wishner, K., Flagg, C., Gardner, W. D., Gaurin, S., Naqvi, S. W. A., Manghnani, V., Prosperie, L. & Gundersen, J. S. 1999. The oxygen minimum zone in the Arabian Sea during 1995. Deep-Sea Research 46, 1903–1931. Müller, M. C., Bernhard, J. M. & Jouin-Toulmond, C. 2001. A new member of Nerillidae (Annelida: Polychaeta), Xenonerilla bacteriicola gen. et sp. nov., collected off California, USA. Cahiers de Biologie Marine 42, 203–217. Mullins, H. T., Thompson, J. B., McDougall, K. & Vercoutere, T. L. 1985. Oxygen-minimum zone edge effects: evidence from the central California coastal upwelling system. Geology 13, 491–494. Murray, J. M. H., Meadows, P. S., Meadows, A., West, F. J. C. & Wood, D. M. 2000. Micro-scale geotechnical variability in continental slope and abyssal sediments inﬂuenced by the oxygen minimum zone in the Arabian Sea. Deep-Sea Research, Part II. 47, 281–301. Murray, J. W. 2001. The niche of benthic foraminifera, critical thresholds and proxies. Marine Micropaleontology 41, 1–7. Naidu, P. D. & Malmgren, B. A. 1995. Do benthic foraminifer records represent a productivity index in oxygen minimum zone areas – an evaluation from the Oman margin, Arabian Sea. Marine Micropaleontology 26, 49–55. Neira, C., Gad, G., Arroyo, N. L. & Decraemer, W. 2001a. Glochinema bathyperuvensis sp. n. (Nematoda, Epsilonematidae): a new species from Peruvian bathyal sediments, SE Paciﬁc Ocean. Contributions to Zoology 70, 147–159. Neira, C., Sellanes, J., Levin, L. A. & Arntz, W. E. 2001b. Meiofaunal distributions on the Peru margin: relationship to oxygen and organic matter availability. Deep-Sea Research 48, 2453–2472. Neira, C., Sellanes, J., Soto, A., Gutiérrez, D. & Gallardo, V. A. 2001c. Meiofauna and sedimentary organic matter off central Chile: response to changes caused by the 1997–98 El Niño. Oceanologica Acta 24, 313–328. Nienstedt, J. C. & Arnold, A. J. 1988. The distribution of benthic foraminifera on seamounts near the East Paciﬁc Rise. Journal of Foraminiferal Research 18, 237–249. Oliver, P. G. 2001. Functional morphology and description of a new species of Amygdalum (Mytiloidea) from the oxygen minimum zone of the Arabian Sea. Journal of Molluscan Studies 67, 225–241.

41

LISA A. LEVIN

Owen, R. W. & Zeitzschel, B. 1970. Phytoplankton production: seasonal change in the oceanic eastern tropical Paciﬁc. Marine Biology 7, 32–36. Paropkari, A. L., Prakash Babu, C. & Mascarenhas, A. 1993. New evidence for enhanced preservation of organic carbon in contact with oxygen minimum zone on the western continental slope of India. Marine Geology 111, 7–13. Pearson, T. H. & Rosenberg, R. 1978. Macrobenthic succession in relation to organic enrichment and pollution in the marine environment. Oceanography and Marine Biology: an Annual Review 16, 229–311. Pederson, T. F., Shimmield, G. B. & Price, N. B. 1992. Lack of enhanced preservation of organic matter in sediments under the oxygen minimum on the Oman margin. Geochimica Cosmochimica Acta 56, 454–551. Perez-Cruz, L. L. & Machain-Castillo, M. L. 1990. Benthic foraminifera of the oxygen minimum zone, continental shelf of the Gulf of Tehuantepec, Mexico. Journal of Foraminiferal Research 20, 312–325. Pfannkuche, O., Sommer, S. & Kahler, A. 2000. Coupling between phytodetritus deposition and the small-sized benthic biota in the deep Arabian Sea: analyses of biogenic sediment compounds. Deep-Sea Research Part II 47, 2805–2833. Phleger, P. B. & Soutar, A. 1973. Production of benthic foraminifera in three east Paciﬁc oxygen minima. Micropaleontology 19, 110–115. Pike, J., Bernhard, J. M., Moreton, S. G., Butler, I. B. 2001. Microbioirrigation of marine sediments in dysoxic environments: implication for early sediment fabric formation and diagenetic processes. Geology 29, 923–926. Piper, D. Z. & Codispoti, L. A. 1975. Marine phosphorite deposits and the nitrogen cycle. Science 188, 15–18. Rabalais, N. N., Harper, D. E. & Turner, R. E. 2001. Responses of nekton and demersal and benthic fauna to decreasing oxygen concentrations. In Coastal hypoxia: consequences for living resources and ecosystems, N. Rabalais & R. Turner (eds). Washington, DC: American Geophysical Union, 115–128. Rabalais, N. N. & Turner, R. E. (eds). 2001. Coastal hypoxia: consequences for living resources and ecosystems. Washington DC: American Geophysical Union,. Reichart, G. L., Lourens, L. J. & Zachariasse, W. J. 1998. Temporal variability in the northern Arabian Sea oxygen minimum zone (OMZ) during the last 225,000 years. Paleoceanography 13, 607–621. Reimers, C., Lange, C., Tabak, M. & Bernhard, J. 1990. Seasonal spillover and varve formation in the heart of the Santa Barbara Basin, California. Limnology and Oceanography 35, 1577–1585. Resig, J. M. & Glenn, C. R. 1997. Foraminifera encrusting phosphatic hardgrounds of the Peruvian upwelling zone: taxonomy, geochemistry, and distribution. Journal of Foraminiferal Research 27, 133–150. Rex, M. A. 1981. Geographic patterns of species diversity in the deep-sea benthos. In The sea V.8, G. T. Rowe (ed.). New York: John Wiley, 453–472. Rhoads, D. C., McCall, P. L. & Yingst, J. Y. 1978. Disturbance and production on the estuarine seaﬂoor. American Scientist 66, 577–586. Rhoads, D. C. & Morse, J. W. 1971. Evolutionary and ecologic signiﬁcance of oxygen-deﬁcient marine basins. Lethaia 4, 413–428. Rhoads, D. C., Mulssow, S. G., Gutschick, R., Baldwin, C. T. & Stolz, J. F. 1991. The dysaerobic zone revisited: a magnetic facies? In Modern and ancient continental shelf anoxia, R. V. Tyson & T. H. Pearson (eds). Geological Society Special Publication No. 58. London: Geological Society, 187–199. Rogers, A. D. 2000. The role of oceanic oxygen minimum zones in generating biodiversity in the deep sea. Deep-Sea Research 47, 119–148. Rosenberg, R., Arntz, W. E., Chuman de Flores, E., Flores, L. A., Carbajal, G., Finger, G. & Tarazona, J. 1983. Benthos biomass and oxygen deﬁciency in the upwelling system off Peru. Journal of Marine Research 41, 263–279.

42

OXYGEN MINIMUM ZONE BENTHOS

Rowe, G. T. 1971. Benthic biomass in the Pisco, Peru upwelling. Investigación Pesquera 35, 127–135. Sanders, H. 1969. Benthic marine diversity and the stability-time hypothesis. Brookhaven Symposium in Biology 22, 71–81. Sanders, N. K. & Childress, J. J. 1990. Adaptations to the deep-sea oxygen minimum layer: oxygen binding by the hemocyanin of the bathypelagic mysid, Gnathophausia ingens Dohrn. Biological Bulletin 178, 286–294. Sankaranarayanan, V. N. & Qasim, S. Z. 1968. The inﬂuence of some hydrographical factors on the ﬁsheries of the Cochin area. Bulletin of the National Institute of Sciences of India 38, 846–853. Sarmiento, J. L., Herbert, T. D. & Toggweiler, J. R. 1988. Causes of anoxia in the world ocean. Global Biogeochemical Cycles 2, 115–128. Savrda, C. E. & Bottjer, D. J. 1987. Trace fossils as indicators of bottom-water redox conditions in ancient marine environments. In New concepts in the use of biogenic sedimentary structures for paleoenvironmental interpretation, D. J. Bottjer (ed.). Society of Economic Paleontologists and Mineralogists, Paciﬁc Section, Volume and Guidebook 52, 3–26. Savrda, C. E. & Bottjer, D. J. 1991. Oxygen-related biofacies in marine strata, An overview and update. In Modern and ancient continental shelf anoxia, R. V. Tyson & T. H. Pearson (eds). Geological Society Special Publication No. 58. London: Geological Society, 201–219. Savrda, C. E., Bottjer, D. J. & Gorsline, D. S. 1984. Development of a comprehensive oxygen-deﬁcient marine biofacies model: evidence from Santa Monica, San Pedro, and Santa Barbara Basins, California Borderland. American Association of Petroleum Geologists Bulletin 68, 1179–1192. Schmaljohann, R., Drews, M., Walter, S., Linke, P., von Rad, U. & Imhoff, J. F. 2001. Oxygenminimum zone sediments in the northeastern Arabian Sea off Pakistan: a habitat for the bacterium Thioploca. Marine Ecology Progress Series 200, 27–42. Schulz, H. N., Brinkhoff, T., Ferdelman, T. G., Hernandez, M., Teske, A. & Jørgensen, B. B. 1999. Dense populations of a giant sulfur bacterium in Namibian shelf sediments. Science 284, 493–95. Schulz, H. N, Jørgensen, B. B., Fossing, H. & Ramsing, R. 1996. Community structure of ﬁlamentous, sheath-building sulfur bacteria, Thioploca spp., off the coast of Chile. Applied Environmental Microbiology 62, 1855–1862. Sellanes, J. 2002. Efectos de El Niño 1997–98 sobre los indicadores bioquímicos de la material orgánica sedimentaria y la meiofauna en un area de surgencia costera de Chile central. PhD thesis, Universidad de Concepcion, Chile. Sen Gupta, B. K. & Machain-Castillo, M. L. 1993. Benthic foraminifera in oxygen-poor habitats. Marine Micropaleontology 20, 183–201. Shull, D. H. & Mayer, L. M. 2002. Dissolution of particle-reactive radionuclides in deposit-feeder digestive ﬂuids. Limnology and Oceanography 47, 1530–1536. Smallwood, B. J. & Wolff, G. A. 2000. Molecular characterization of organic matter in sediments underlying the oxygen minimum zone at the Oman margin, Arabian Sea. Deep-Sea Research, Part II 47, 353–375. Smallwood, B. J., Wolff, G. A., Smith, C. R., Hoover, D., Gage, J. D. & Patience, A. 1999. Megafauna can control the quality of organic matter in marine sediments. Naturwissenschaften 86, 320–324. Smith, C. R. 1986. Nekton falls, low intensity disturbance and community structure of infaunal benthos in the deep sea. Journal of Marine Research 44, 567–600. Smith, C. R. & Hamilton, S. C. 1983. Epibenthic megafauna of a bathyal basin off southern California: patterns of abundance, biomass and dispersion. Deep-Sea Research 30, 907–928. Smith, C. R., Pope, R. H., DeMaster, D. J. & Magaaard, L. 1993. Age-dependent mixing of deep-sea sediments. Geochimica et Cosmochimica Acta 57, 1473–1488. Smith, C. R., Levin, L. A., Hoover, D. J., Cremer, M., McMurtry, G. & Gage, J. D. 2000. Bioturbation across the oxygen minimum zone on the Arabian-Sea slope. Deep-Sea Research, Part II 47, 227–257. Smith, S. L. 1982. The Northwestern Indian Ocean during the monsoons of 1979: distribution, abundance and feeding of zooplankton. Deep-Sea Research 29, 1331–1353.

43

LISA A. LEVIN

Smith, S. L., Roman, M., Wishner, K., Gowing, M. M., Codispoti, L., Barber, R. Marra, J., Prusova, I. & Flagg, C. 1998. Seasonal response of zooplankton to monsoonal reversals in the Arabian Sea. Deep-Sea Research 45, 2369–2404. Spicer, J. I. & Gaston, K. J. 1999. Amphipod gigantism dictated by oxygen availability. Ecology Letters 2, 397–401. Spiess, F. 1928. The Meteor expedition. Berlin: Vertag von Dietrich Reimer. Stott, L. D., Berelson, W., Douglas, R. & Gorsline, D. 2000. Increased dissolved oxygen in Paciﬁc intermediate waters due to lower rates of carbon oxidation in sediments. Nature 407, 367–370. Tarazona, J., Salzwedel, H. & Arntz, W. 1988a. Positive effects of “El Niño” on macrobenthos inhabiting hypoxic areas of the Peruvian upwelling system. Oecologia 76, 184–190. Tarazona, J., Salzwedel, H. & Arntz, W. E. 1988b. Oscillations of macrobenthos in shallow waters of the Peruvian central coast induced by El Niño 1982–83. Journal of Marine Research 46, 593–611. Thiel, H. 1978. Benthos in upwelling regions. In Upwelling systems, R. Boje & M. Tomczak (eds). Berlin: Springer, 124–138. Thompson, B., Tsukada, D. & Laughlin, J. 1993. Megabenthic assemblages of coastal shelves, slopes and basins off southern California. Bulletin of the Southern California Academy of Sciences 92, 25–42. Thompson, J. B., Mullins, H. T., Newton, C. R. & Vercoutere, T. 1985. Alternative biofacies model for dysaerobic communities. Lethaia 18, 167–179. Todaro, M. A., Bernhard, J. M. & Hummon, W. D. 2000. A new species of Urodasys (Gastrotricha, Macrodasyida) from dysoxic sediments of the Santa Barbara Basin (California, USA). Bulletin of Marine Science 66, 467–476. Travis, J. L. & Bowser, S. S. 1986. Microtubule-dependent reticulopodial motility. Is there a role for actin? Cell Motility and the Cytoskeleton 6, 146–152. Tyson, R. V. & Pearson, T. H. 1991. Modern and ancient continental shelf anoxia: an overview. In Modern and ancient continental shelf anoxia, R. V. Tyson & T. H. Pearson (eds). Geological Society Special Publication No. 58. London: Geological Society, 1–24. van der Zwaan, G. J., Duijnstee, I. A. P., den Dulk, M., Ernst, S. R., Jannik, N. T. & Kouwenhoen, T. J. 1999. Benthic foraminifers: proxies or problems? A review of paleoecological concepts. Earth Sciences Review 46, 213–236. Vetter, R. D., Lynn, E. A., Garza, M. & Costa, A. S. 1994. Depth zonation and metabolic adaptation in Dover sole, Microstomus paciﬁcus and other deep-living ﬂatﬁshes: factors that affect the sole. Marine Biology 120, 145–159. von Rad, U., Schulz, H. & Sonne 90 Scientiﬁc Party. 1995. Sampling the oxygen minimum zone off Pakistan: glacial-interglacial variations of anoxia and productivity (Preliminary results, SONNE 90 cruise). Marine Geology 125, 7–19. Wakeﬁeld, W. W. & Smith, K. L. 1990. Ontogenetic vertical migration in Sebastolobus altivelis as a mechanism for transport of particulate organic matter at continental slope depths. Limnology & Oceanography 35, 1314–1328. Ward, B. B., Glover, H. E. & Lipschultz, F. 1989. Chemoautotrophic activity and nitriﬁcation in the oxygen minimum zone off Peru. Deep-Sea Research 36, 1031–1051. Weeks, S. J., Currie, B. & Bakun, A. 2002. Satellite imaging – massive emissions of toxic gas in the Atlantic. Nature 415, 493–494. Wheatcroft, R. A. 1989a. Characteristic trace-fossil associations in oxygen-poor sedimentary environments; discussion. Geology 17, 674. Wheatcroft, R. A. 1989b. Comment and Reply on “Characteristic trace-fossil associations in oxygenpoor sedimentary environments”. Geology 17, 674. White, B. N. 1987. Oceanic anoxic events and allopatric speciation in the deep sea. Biological Oceanography 5, 243–259. Wishner, K. F., Ashjian, J., Gelfman, C., Gowing, M. M., Kann, L., Levin, L. A., Mullineaux, L. & Saltzman, J. 1995. Pelagic and benthic ecology of the lower interface of the Eastern Tropical Paciﬁc oxygen minimum zone. Deep-Sea Research 42, 93–115.

44

OXYGEN MINIMUM ZONE BENTHOS

Wishner, K. F., Gowing, M. M. & Gelfman, C. 1998. Mesozooplankton biomass in the upper 1000 m in the Arabian Sea: overall seasonal and geographic patterns, and relationship to oxygen gradients. Deep-Sea Research, Part II 45, 2405–2432. Wishner, K. F., Gowing, M. M. & Gelfman, C. 2000. Living in suboxia: ecology of an Arabian Sea oxygen minimum zone copepod. Limnology & Oceanography 45, 1576–1593. Wishner, K., Levin, L., Gowing, M. & Mullineaux, L.1990. Involvement of the oxygen minimum in benthic zonation on a deep seamount. Nature 346, 57–59. Wyrtki, K. 1962. The oxygen minima in relation to ocean circulation. Deep-Sea Research 9, 11–23. Wyrtki, K. 1966. Oceanography of the eastern Paciﬁc Ocean. Oceanography and Marine Biology: an Annual Review 4, 33–68. Wyrtki, K. 1973. Physical oceanography of the Indian Ocean. In The biology of the Indian Ocean, B. Zeitzschel (ed.). Berlin: Springer, 18–36. Yang, T. H., Lai, N. C., Graham, J. B. & Somero, G. N. 1992. Respiratory, blood and heart enzymatic adaptations of Seabastolobus alascanus (Scorpaenidae: Teleostei) in the oxygen minimum zone: a comparative study. Biological Bulletin 183, 490–499. Young, C. M. & Vàzquez, E. 1997. Agnezia monnioti and Styela gagetyleri, new deep-sea ascidians specialized for life within and below the oxygen minimum layer in the Arabian Sea. Invertebrate Biology 116, 252–276. Zehnder, A. 1988. Biology of anaerobic organisms. New York: John Wiley.

45

Oceanography and Marine Biology: an Annual Review 2003, 41, 47–114 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

ANTARCTIC MARINE BENTHIC DIVERSITY ANDREW CLARKE & NADINE M. JOHNSTON British Antarctic Survey, High Cross, Madingley Road, Cambridge CB3 0ET, UK e-mail: [email protected]

Abstract Species lists have been compiled for all the major groups of Southern Ocean benthic marine invertebrates, eliminating synonymies where possible and providing a subjective estimate of completeness and reliability for each group. Antarctic marine diversity (pelagic and benthic) is relatively high at the phylum and class level, with the gaps mostly comprising minor, meiofaunal or parasitic groups. Most benthic diversity data come from the continental shelves, with relatively few samples from deeper water. Even for the continental shelves, however, sampling is highly patchy with some areas hardly investigated at all. Over 4100 benthic species have been reported from the Southern Ocean, with the most speciose groups being polychaetes, gastropods and amphipods. Comparison with tropical and temperate regions suggest that decapods, bivalves and teleost ﬁshes are poorly represented in the Southern Ocean benthic marine fauna, whereas pycnogonids, echinoderms and many suspension feeding groups are rich and diverse. Some groups that are currently low in diversity were previously well represented in the Antarctic shallow water marine fauna, notably decapods and many ﬁshes. Other groups have undergone marked radiations in the Southern Ocean, including pycnogonids, amphipods, isopods and teleost ﬁshes; in all cases, however, it is only some lineages that have diversiﬁed. This indicates that evolutionary questions concerning the origin, diversiﬁcation or extinction of the Southern Ocean marine fauna will have no single answer; the evolutionary history of each group appears to reﬂect a different response to the tectonic, climatic and oceanographic changes to which they have been subject through history. The disposition of southern hemisphere continents makes it difﬁcult to assess whether there is a latitudinal cline in shallow-water marine diversity to mirror that known from the northern hemisphere. Within Antarctica, many species appear to have circumpolar distributions, and the long established biogeographical division into continental Antarctic, Antarctic Peninsula and sub-Antarctic regions have not been challenged by recent sampling. For most groups the frequency distribution of species per genus ratios is typical, though none is well described by the predictions from current evolutionary or null models. Where data are available, size spectra indicate that many Southern Ocean taxa are small, a few spectacular examples of gigantism notwithstanding, and species abundance plots are normal. Knowledge of the Southern Ocean benthic marine fauna has reached a stage where we can now ask powerful evolutionary questions, and the development of new molecular techniques provides the mechanism for answering them.

Introduction In the past two or three decades there has been an increasing recognition of the loss of species through the activities of man. This has lead to a resurgence of interest in biological diversity, both in its purely intellectual aspects and in terms of its relevance to conservation, management and environmental issues. Although there was considerable theoretical interest 47

ANDREW CLARKE & NADINE M. JOHNSTON

in the measurement of biological diversity in the 1960s and 1970s, such work then fell out of fashion. This was partly because an understanding of the underlying processes which produced the observed differences in diversity between habitats or areas had proved somewhat elusive. The theoretical framework provided by the work of MacArthur & Wilson (1963, 1967) on island biogeography did not translate easily to the more complex mainland, and no general theory of biological diversity emerged. The present revival of interest in biological diversity can be traced back to a seminal conference, the National Forum on BioDiversity, held in Washington, DC, on 21–24 September 1986. This conference was the birthplace of the neologism biodiversity and resulted in a highly inﬂuential proceedings (Wilson & Peter 1988). There followed an almost exponential rise in the number of scientiﬁc papers and other literature concerned with biological diversity. Particularly inﬂuential amongst these was E. O. Wilson’s lucid call to arms The diversity of life (Wilson 1992). In recent years interest in biological diversity has remained strong, with the appearance of a number of important text books and edited volumes (Magurran 1988, Huston 1994, Rosenzweig 1995, Hawksworth 1995, Gaston 1996) and compendious reviews of data (World Conservation Monitoring Centre 1992, Heywood 1995).

What is biological diversity? There are currently well over a dozen formal deﬁnitions of biological diversity or biodiversity. The deﬁnition which has come to be most universally accepted is that enshrined within the Convention on Biological Diversity. This far-reaching treaty was signed by 156 nations on 5 June 1992 at the United Nations Conference on Environment and Development (UNCED) in Rio de Janeiro. The Convention on Biological Diversity was perhaps the most important outcome of UNCED, and it came into force approximately 18 months later. The UNCED deﬁnition of biodiversity is laid out in Article 2 of the Convention on Biological Diversity. It is: “Biological diversity” means the variability among living organisms from all sources including, inter alia, terrestrial, marine and other aquatic ecosystems and the ecological complexes of which they are part; this includes diversity within species, between species and of ecosystems. This deﬁnition provides a broad conceptual framework for any consideration of biological diversity. It also creates a problem in that it includes everything from the genome to the ecosystem. Although such all-encompassing deﬁnitions may have value in a political or management context, they tend not to help the development of science. More recent deﬁnitions have tended to be more focused, and a typical one is that of Hubbell (2001) who deﬁnes biological diversity as . . . synonymous with species richness and relative species abundance in space and time. Species richness is simply the total number of species in a deﬁned space at a given time, and relative species abundance refers to their commonness or rarity. This deﬁnition, and many others like it, are far more useful in being more clearly expressed and restrictive. They also highlight a major concern of the early studies of biological diver48

ANTARCTIC MARINE BENTHIC DIVERSITY

sity, which was the problem of how best to express the number of taxa (usually species) in an area or sample. The challenge was to devise a statistic which simultaneously expressed the number of species, and the distribution of individuals within those species. This proved both difﬁcult and contentious, but these early theoretical studies did clarify the important distinction between local and regional taxonomic richness (MacArthur 1965, 1972, Whittaker 1972, 1977). In attempting to develop a pragmatic approach to quantifying biological diversity, ecologists have recently developed three main themes. The ﬁrst has stemmed from the recognition of the role of evolution and the intuitive feel that taxonomic distinctiveness is an important criterion in diversity: an assemblage of six grass species and an assemblage composed of a single grass, a moss, two species of shrub and two trees contain the same number of species but clearly differ greatly in some intuitive concept of diversity. Clarke & Warwick (1998, 1999, 2001) have developed a number of innovative new diversity indices which incorporate taxonomic distinctiveness or the evolutionary relatedness of component taxa, and a similar approach has been developed in parallel in terrestrial ecology (Webb 2000). Warwick & Clarke (2001) provide a recent review of this topic. The second theme has been a move to studies of diversity at the molecular level. These have been important in emphasising the enormous taxonomic and functional diversity in organisms previously regarded as somewhat more uniform (particularly Archaea and Eubacteria). These approaches provide a valuable intellectual challenge to preconceived notions of metazoan importance, but they also emphasise both the limited value of a single measure of biological diversity and the danger of limiting studies of biological diversity to the well known macroscopic metazoans. The third theme in recent work on biological diversity has been the return to simple taxon richness as a straightforward and informative quantitative measure of diversity. It had long been recognised that most intuitive concepts of diversity somehow merged the simple number of taxa present with the distribution of individuals amongst those taxa: a given number of equally common species form an assemblage that is in some way more diverse than one comprised of the same number of species but which vary in abundance. Unfortunately, attempts to incorporate both species number and species abundance into a single diversity index have resulted in no single agreed approach, and taxon richness has once again emerged as the most useful simple measure. Taxon richness does, of course, run the severe risk of sampling bias but when used with care it has the value of being quick, simple and informative, particularly when assessing diversity over large spatial scales where relative abundance data are rarely available. Furthermore, rare species can be as informative for evolutionary or biogeographic considerations as common ones. Generally, species are regarded as the fundamental unit of diversity (Claridge et al. 1997). In some cases species level data are not available or are unreliable; here genus or family level richness can provide a good indication of underlying species richness. Examples are rapid assessments of the diversity of a previously unsurveyed area or, in palaeobiology, where sampling error is too great at the species level (Roy et al. 1996). A useful brief review of the use of higher taxonomic level richness measures is given by Lee (1997).

49

ANDREW CLARKE & NADINE M. JOHNSTON

Marine diversity The sea covers two-thirds of the earth’s surface, making sea water the single largest habitat there is. Life originated in the sea and today marine metazoan diversity at higher taxonomic levels (class, phylum) signiﬁcantly exceeds that on land (Table 1). The precise number of classes and phyla depends on the taxonomy used, as shown here by the contrast between the compilation by May (1994) and this study. Also listed in Table 1 are the number of major groups (classes and phyla) reported so far from the Southern Ocean; these numbers will almost certainly change as the Antarctic marine fauna becomes better known and taxonomies are revised. At lower taxonomic levels the contrast between land and sea is very much reversed: the vast majority of described species are terrestrial. It is not at all straightforward to estimate the number of species described to date; Wilson (1988) provided a remarkably precise estimate of 1 392 485, May (1988) one of 1.8 million, and Stork (1988) a similar value of 1.82 million. Minelli (1993) has contributed an insightful review of the nature of the data on which such estimates are based, and provided his own estimate of 1.8 million described species. All authors recognise the many difﬁcult problems inherent in making such estimates, notably the need to add together very different types of “species” but there is general agreement that at present about 1.8 million species have been described. Of this total, only about 200 000 species are marine (Grassle 2001) whereas approximately 1 million are insects (an almost exclusively terrestrial group). This striking difference is driven in large part by the intense species richness of some insect groups (notably Hymenoptera, Diptera, Coleoptera, Lepidoptera and Homoptera), but it may also be a reﬂection of our ignorance of the sea. In a highly inﬂuential paper, Grassle & Maciolek (1992) suggested that there may be a vast number of undescribed species in the deep sea. There followed an intense, and as yet unresolved, debate over the extent to which the deep sea contains a largely undescribed fauna (May 1993, 1994, Poore & Wilson 1993, Boucher & Lambshead 1995, Lambshead et al. 2000, Snelgrove & Smith 2002). Knowlton (1993) has drawn attention to the possibility that undescribed sibling species may harbour an immense amount of marine diversity, and Reaka-Kudla (1997) has estimated that coral reefs alone may contain over 600 000 species (though perhaps only 35 000 to 60 000 have been described so far). It is clear that we really have no idea how many species there may be in the sea. A series of recent estimates for individual higher taxa are shown in Table 2. It is striking how much variation there is even for apparently well known taxa such as molluscs. Table 1 Two estimates of marine diversity at higher taxonomic levels. Data are restricted to metazoan animal phyla. Previous data are from Nicol (1971) as summarised by May (1994). Data for this study are based on taxonomy of Barnes (1998). The data for the Southern Ocean are based on all occurrences known to the authors; the number of marine phyla and classes reported for Antarctica will undoubtedly increase with further sampling and taxonomic work. nd no data. Total

May (1994) Land Sea

Phylum

33

12

32

Class

nd

33

73

This study Sea

Southern Ocean

38

36

28

100

90

58

Total

50

51

150 10 240

Kinorhyncha Loricifera Nematomorpha

Rotifera Acanthocephala Chaetognatha Onychophora Tardigrada

2000 600 – 80 –

80 000

400 8

Gastrotricha Priapula

Nematoda

80

Gnathostomula

–

15 000

Platyhelminthes

Orthonecta

9500 90

10 000 –

Margulis

Cnidaria Ctenophora Rhombozoa

Placozoa

Porifera Symplasma

Phylum

– – – – –

12 000

– – –

– –

–

–

–

–

5000 –

2000 750 110 – 531

20 000

74 – 230

– 9

–

–

–

50

15 000

2

6000 –

1800 900 200 80 600

20 000

150 100 325

430 17

–

–

20 000

8000 80

1

5000 –

Approximate number of species Wilson Minelli Nielsen

1800 1000 200 70 600

20 000

150 100 325

450 17

100

10

25 000

10 000 100 75

1

10 000 500

Barnes

Almost exclusively pelagic but a few benthic taxa known Problematic obscure organisms, once classiﬁed as Mesozoa and sometimes treated as two phyla, Dicyemida and Heterocyemida, the latter with only two known species Predominantly parasitic but includes free-living Turbellaria with planktonic and benthic species Obscure phylum of parasitic organisms, previously grouped with Rhombozoa in Mesozoa Meiofaunal group of uncertain afﬁnity; included with annelids by Nielsen Meiofaunal; marine and freshwater A small group of infaunal worms with a fossil record extending back to the Cambrian in the Burgess Shale Meiofaunal; marine First described in 1983; meiofaunal; marine Juveniles parasitic in arthropods; adult stages free-living but short-lived and non-feeding; mostly freshwater Many parasitic forms; a major component of the meiofauna and possibly harbouring an enormous number of undescribed marine species Mostly freshwater Exclusively parasitic Pelagic marine; higher level taxonomy (class, order) unresolved Exclusively terrestrial Meiofaunal

Hexactinellids grouped with true sponges (Porifera) by all authors except Barnes Enigmatic phylum comprising Trichoplax adhaerans and possibly one other species

Comments

Table 2 A classiﬁcation of metazoan animal phyla with the approximate number of described species. The taxonomy and sequence of phyla follows Barnes (1998), with species numbers taken from Margulis & Schwartz (1982), Wilson (1988), Minelli (1993), Nielsen (1995) and Barnes (1998).

ANTARCTIC MARINE BENTHIC DIVERSITY

continued

Margulis

Approximate number of species Wilson Minelli Nielsen Barnes

Comments

Pentastoma 70 – – – 100 A small phylum of uncertain afﬁnities; parasitic Crustacea – – – – 40 000 Chelicerata – – – – 63 000 Includes pycnogonids (sometimes treated as a separate phylum) Uniramia – – – – 1 000 000 Insects and allies; almost exclusively terrestrial Crustacea, Chelicerata, Uniramia and sometimes Pentastoma often combined as a single phylum: Arthropoda “Arthropoda” – 874 161 – 1 000 000 – Nemertea 900 – 950 900 900 Mollusca 110 000 50 000 130 000 100 000 100 000 Sipuncula 300 – – 320 350 Echiura 140 – 140 – 150 Usually viewed as a separate phylum, but grouped by Nielsen with annelids Annelida 8800 12 000 18 600 15 000 15 000 Pogonophora 100 – 100 – 150 Two classes (Perviata, Vestimentifera) sometimes regarded as separate phyla; grouped by Nielsen with annelids Entoprocta 150 – 150 150 150 A small phylum of marine lophophorates (also known as Kamptozoa) Cycliophora – – – – 1 First described 1995; only one species known, Symbion pandora (from mouthparts of Nephrops) Phorona – – 10 12 20 A small phylum of marine lophophorates Brachiopoda 335 – 335 300 350 A once abundant group of sessile lophophorates Bryozoa 5000 – 5000 4000 4300 The most diverse lophophorate phylum; also called Ectoprocta (Nielsen) or Polyzoa (defunct) Hemichordata – – 100 70 100 Includes Pterobranchia and Enteropneusta, regarded as separate phyla by Nielsen Echinodermata 6000 6100 6700 7000 7000 Most taxonomists regard the enigmatic Xyloplax (2 species) as an echinoderm Chordata – 43 000 48 000 – 43 000 A single phylum comprising Urochordata (Tunicata), Cephalochordata and Vertebrata, sometimes each given individual status as phyla (e.g. by Nielsen) Incertae sedis Senoturbella bocki Buddenbrockia plumatellae Lobatocerebrum (3 or 4 species) (Nielsen regards these as specialised annelids) Salinella salva (known only from a single report from saline lagoons in Argentina)

Phylum

Table 2

ANDREW CLARKE & NADINE M. JOHNSTON

52

ANTARCTIC MARINE BENTHIC DIVERSITY

The Southern Ocean The Southern Ocean comprises all waters south of the Polar Front (referred to as the Antarctic Convergence in the earlier literature). This well-deﬁned circum-Antarctic oceanographic feature marks the northernmost extent of cold surface water. The total area of the Southern Ocean is thus about 34.8 million km2. Of this, up to 21 million km22 is covered by ice at the winter maximum but only about 7 million km2 is covered at the summer minimum (Gloersen et al. 1992). Two inﬂuential schemes for subdividing this vast area of ocean are those of Tréguer & Jacques (1992) based predominantly on ice and nutrient dynamics, and Longhurst (1998) based on upper water column structure and remotely-sensed phytoplankton pigments. These schemes are based essentially on surface processes but they are relevant to the benthos in that almost all life on the sea bed depends on the ﬂux of material from surface waters for its energy and nutrients. Patterns of surface production will therefore inﬂuence the diversity, abundance and ecology of benthic organisms, albeit modiﬁed by advective processes. Much of the Southern Ocean overlies deep sea ﬂoor (Fig. 1). Relatively little of the Southern Ocean sea bed is continental shelf, and much of this shelf is unusually deep as a result of scouring from ice shelves and depression by the enormous mass of continental ice (Clarke 1996). Continental shelves elsewhere in the world are typically 100–200 m deep and 75 km wide (Walsh 1988); those around Antarctica average over 450 m deep, and in places they extend to over 1000 m depth. The deepest areas are trenches and basins, with the edge of the shelf usually being somewhat shallower. Although in some areas the continental shelves around Antarctica are narrow, and in many places are overlain extensively by ice shelves, their average width of 125 km is almost twice that of continental shelves elsewhere. This large average width is caused predominantly by the inﬂuence of the Ross Sea and the Weddell Sea. In both of these vast embayments much of the continental shelf is covered by ﬂoating ice shelf (Fig. 1). In other areas, for example, Dronning Maud Land between 0ºE and 70ºE, the continental shelf is relatively narrow. For the purpose of this review we have taken the 1000 m isobath to mark the edge of the continental shelf. The transition from the continental slope to the continental rise is less clear cut, and we have used the 3000 m isobath (Snelgrove 2001). It is likely (but not known for certain) that there is little benthic life beneath the permanent ice shelves of the Ross and Weddell Seas, and other less extensive ice shelves (Lipps et al. 1977, 1979). The area of Antarctic continental shelf that is not beneath permanent ice totals about 3 million km2 (Table 3). This is 8.5% of the Southern Ocean, and about 11.4% of the world’s total continental shelf area, which is estimated to be 31.1 million km2 (Walsh 1988). The great depths of the Antarctic continental shelves means that the fauna living there may be expected to show physiological adaptations similar to (but perhaps not as marked as) those in the deep sea. Like that of the deep sea, the Antarctic continental shelf fauna must also be adapted to low temperatures and a marked seasonality of food (Clarke 1996).

53

ANDREW CLARKE & NADINE M. JOHNSTON

Figure 1 Map of the Southern Ocean showing the mean position of the Polar Front (the Antarctic Convergence in earlier literature) and broad scale bathymetry around the Antarctic continent. The 1000 m isobath marks the edge of the continental shelf around Antarctica, and the 3000 m isobath is taken to mark the transition from the continental slope to the deep sea. For reasons of clarity areas of depth 3000 m associated with mid-ocean ridges are not shown.

54

ANTARCTIC MARINE BENTHIC DIVERSITY

Table 3 Some characteristics of the Southern Ocean benthic environment. The fractal nature of coastlines and depth contours mean that all data depend on the length scale used. Deep-sea area calculated by subtraction, so ignores mid-ocean ridges. All data from latest edition of Antarctic Digital Database (ADD Consortium 2002), with mean position of the Polar Front from Trathan et al. (1997, 2000), Orsi et al. (1995). Coastline Length (km 103) Ice coastline (%) Ice front (%) Rock coastline (%) Continental shelf (1000 m depth) Area not beneath ice shelves in 2002 (km2 106) Area beneath ice shelves in 2002 (km2 106) Total area (km2 106) Continental slope (area between 1000 m and 3000 m isobaths) Area (km2 106) Deep Sea (3000 m depth) Area (km2 106) Total area of Southern Ocean (km2 106) Permanently open water Seasonal ice cover (max) Seasonal ice cover (min)

39.2 39.9 46.4 13.7 2.97 1.63 4.59 2.35 27.9 34.8 14 21 7

Aims of this study The questions we have attempted to answer in this review were: (1) (2) (3)

How well do we know the Southern Ocean benthic marine fauna? Which groups of benthic marine organisms are well represented in the Southern Ocean, and which are not? To what extent does the Southern Ocean marine fauna match, or deviate from, established macroevolutionary patterns such as the latitudinal cline in diversity?

This required the compilation of validated species lists for all the major benthic taxa of the Southern Ocean. These data were then analysed for temporal trends in species description and the distribution of species amongst higher taxa (genus, family). Some preliminary macroecological analyses were also undertaken, and comparisons made with data from other geographical areas.

55

ANDREW CLARKE & NADINE M. JOHNSTON

Methods Data compilation The initial task for this study was to assemble a species list for each of the major benthic taxa of the Southern Ocean. The approach taken was to combine a study of the most recent taxonomic review of a particular group with a thorough search of both the earlier literature and all of the literature published since that review that we could locate. Extensive searches of the modern literature were undertaken to locate newly described taxa, and the general reviews most important as a starting point were those of Dell (1972, 1990) and Arntz et al. (1994, 1997). Our deﬁnition of the Southern Ocean means that we have excluded taxa found only at Tristan da Cunha, Gough Island, Prince Edward and Marion Islands, Macquarie Island, Îles Crozet, Îles Kerguelen and the Magellan region of South America. Data were stored in simple spreadsheet databases. For each major group these spreadsheets listed (where known), species, taxonomic authority and date, higher level taxonomy and references. The use of established numerical taxonomic codes was investigated but none proved suitable for use with the Southern Ocean fauna without extensive additional work. The higher level taxonomy (especially at the family level) of many phyla varies widely between different monographs. We have therefore generally utilised the higher level taxonomy given in the most recent taxonomic revision. Where a particular group has been recently revised or reviewed in a monograph then synonymy is generally minimal. Where no recent revision is available we have not always been able to eliminate synonymies, and our compilation of the literature may thus overestimate species richness. In these groups we have also been unable to eliminate all inconsistencies in the taxonomic authorities (names, dates and requirement for parentheses). In a few cases we were unable to establish taxonomic authorities at all.

Distributional analyses For plotting sampling data an established Geographic Information System (GIS) was used. The outline of the Antarctic continent was provided by Version 3.0 of the Antarctic Digital Database (ADD Consortium 2000), and the mean position of the Polar Frontal Zone digitised from previous work (Trathan et al. 1997, 2000, Orsi et al. 1995). Bathymetric data proved far more problematical. Hydrographic maps of varying degrees of precision and accuracy exist for many areas of Antarctica but almost none of these is in digital form. Work by various nations has led to a small number of high resolution digital bathymetric datasets, but these generally cover only a small geographic area of the Southern Ocean. The best general bathymetric data for the Southern Ocean come from the General Bathymetric Chart for the Oceans (GEBCO) which is available in digital form on CD-ROM. It was decided to utilise this as the basis for plotting Southern Ocean marine diversity data, but unfortunately the GEBCO bathymetric contours contained many gaps where data were non-existent. In order to render such contours compatible with a GIS all these gaps were ﬁlled by interpolation. It was originally planned to record sampling location data for all taxa but it was quickly realised that this would take several years of work. Instead we chose to concentrate on two 56

ANTARCTIC MARINE BENTHIC DIVERSITY

well-studied taxa in the expectation that these would reveal general patterns for Antarctic marine benthos; the taxa chosen were gastropod and bivalve molluscs. For these we recorded the geographical co-ordinates (latitude, longitude) for every occurrence recorded in the literature. Much of the literature provides only a general sampling location (for example, South Georgia, or Weddell Sea) and for these no location data could be entered. This compilation has subsequently been extended and validated, forming the Southern Ocean Molluscan Database (SOMBASE: Crame, Grifﬁths & Linse, unpublished British Antarctic Survey data).

Historical and macroecological analyses For each major group (phylum, class) the data were examined to establish the historical pattern of species description. The temporal trend in species description when coupled with knowledge of recent taxonomic monographs or revisions can provide an indication as to how well known the group is (Costello et al. 1996). A recent thorough revision of a well collected group, in which few new species have been described recently, indicates a well documented fauna. Conversely, a cumulative species description which is still climbing rapidly indicates an incompletely documented fauna. We have also undertaken some preliminary evolutionary and macroecological analyses of the major groups. These have included a compilation of the number of species for each family, as a ﬁrst order indication of which clades have radiated in the Southern Ocean, and an initial analysis of species to genus ratios within the major taxa.

Marine benthic diversity in the Southern Ocean: an inventory The species lists compiled for this study total over 4000 and are therefore not reproduced here. Rather, for each major group we list all of the important taxonomic synopses, indicating whether these are complete or cover only a taxonomic or geographic subset. The most speciose families are listed, and for most taxa the historical pattern of species description is presented. For most major groups of Antarctic benthic organisms we have been able to provide at least an estimate of known species richness in the Southern Ocean. It is not, however, possible to put a formal conﬁdence estimate on these estimates, for we cannot judge how many species remain to be discovered. Instead we have assigned each group to one of three classes as an indication of how well that group is known. The criteria were: (1) (2) (3)

Well known: group relatively well sampled; taxonomy reasonably stable with recent monograph or review; relatively few species described in the past decade. Moderately well known: group fairly well sampled; taxonomic review or monograph published in past ﬁfty years; species still being described regularly. Poorly known: group poorly sampled; no taxonomic revision since early in the century; group not currently receiving signiﬁcant taxonomic attention in the Southern Ocean.

This is a purely subjective assessment based on the most recent taxonomic revision, the history of species description, and personal knowledge of taxonomic work underway. There 57

ANDREW CLARKE & NADINE M. JOHNSTON

may be, of course, active taxonomic work of which we are unaware, and these assessments cannot take into account the very large amount of work by scientists from the former Soviet Union which is currently unavailable to western scientists. There is little agreement amongst systemacists or evolutionary biologists as to the higher level classiﬁcation of living organisms, and even the number of phyla depends on the taxonomy in use (Table 1). As we are not taxonomists ourselves and cannot make meaningful judgements concerning higher level systematics, for this compilation and analysis of the inventory of the Southern Ocean benthic marine fauna we have therefore followed the classiﬁcation scheme of Barnes (1998), noting where there are signiﬁcant differences from older or competing arrangements. The sequence of phyla used here also follows Barnes (1998).

Macroalgae Macroalgae (seaweeds) are a convenient grouping of three quite different taxa, Rhodophyta (red algae), Chlorophyta (green algae) and Phaeophyta (brown algae). These are generally treated as separate phyla within the Kingdom Protoctista (Margulis & Schwartz 1982, Barnes 1998). The Antarctic marine ﬂora also contains a single macroscopic chrysophyte and Wiencke & Clayton (2002), in their review of Southern Ocean seaweeds, group the Chrysophyceae and the brown algae (Phaeophyceae) as subtaxa within the phylum Heterokontophyta. The Southern Ocean seaweed ﬂora has a low species richness compared with temperate and tropical regions, with a total of only 119 species described to date (Clayton 1994, Clayton et al. 1997, Wiencke & Clayton 2002). Rhodophyta contribute 75 species to this total, Phaeophyta 26 and Chlorophyta 17, though Wiencke & Clayton (2002) comment that the rhodophytes are almost certainly underestimated. Endemism is high, ranging from 18% (Chlorophyta) to 42% (Phaeophyta). Macroalgal species richness decreases markedly with increasing latitude along the Antarctic Peninsula (Moe & DeLaca 1976), and the seaweed ﬂora of continental Antarctica is very sparse (Wiencke & Clayton 2002).

Phylum Porifera and Phylum Symplasma Sponges are a very difﬁcult group taxonomically. The four classes (Calcarea, Demospongiae, Sclerospongiae and Hexactinellida) are fairly well deﬁned and have traditionally been grouped together in a single phylum (Porifera). It has now become clear that glass-sponges (Hexactinellida) differ from the others in important aspects of internal anatomy and cellular organisation, and are increasingly regarded as a separate phylum (Symplasma). The two phyla, Porifera and Symplasma, are then grouped in the superphylum Parazoa (Barnes 1998). We have followed this arrangement here. The most important synopses for each class of the Antarctic sponge fauna are: • • •

Calcarea: Burton (1929), Brondsted (1931) and Koltun (1964, 1976) document particular material but none provides a complete revision of the entire class. Demospongiae: Sarà et al. (1992) provides a full revision. Sclersopongiae: Sclerosponges tend to occur in association with coral reefs or in caves and tunnels; they appear not to have been reported from the Southern Ocean. 58

ANTARCTIC MARINE BENTHIC DIVERSITY

•

Hexactinellida: A very difﬁcult group, some requiring microscopic examination of isolated spicules to conﬁrm identiﬁcation. Barthel & Tendal (1994) provide a full review for the Southern Ocean taxa.

Important systematic or faunal literature published since the above revisions include Gutt & Koltun (1995) for the Calcarea, Battershill (1989), Pansini et al. (1994), Gutt & Koltun (1995), Kunzmann (1996) and Thomas & Mathew (1986a) for the Demospongiae, and Battershill (1989), Gutt & Koltun (1995), Kunzmann (1996) and Barthel (1997) for the Hexactinellida. The systematics of the Calcarea is unstable and there appears to be no accepted classiﬁcation system at the level of family or order in the literature. The demosponges also have an uncertain systematics and Sarà et al. (1992) presented their species list down to order only. The systematics of the hexactinellids follows that adopted by Barthel & Tendal (1994), who themselves followed Ijima (1927), although somewhat simpliﬁed and with minor changes from Bergquist (1978), Burton (1929), Hartman (1982) and Lévi (1964a,b). The time-course of species description (Fig. 2a) shows clearly how the groundwork for our current knowledge of the Antarctic sponge fauna was laid between the late 1880s and the early 1930s, with important contributions later from Koltun in the 1960s and 1970s. The pattern of species per genus values (Fig. 2b) shows a typical hollow curve distribution (Willis 1922). The most speciose orders are all demosponges: Poecilosclerida (132 species), Haplosclerida (36 species) and Hadromerida (22 species). The Calcarea have only 14 species recorded from Antarctica, and the Sclerospongiae appear to be absent. The glasssponges (Hexactinellida) are a very important group ecologically, and there are 29 species reported from the Southern Ocean.

(a)

(b)

Figure 2 Sponges (true sponges, Porifera and glass sponges, Symplasma, combined). (a) Time course of description of Southern Ocean sponge taxa; data presented in percentage terms (total 279 species). (b) Distribution of species amongst genera for sponges. Data presented as a frequency distribution of species per genus values, with data for Porifera and Symplasma shown separately.

59

ANDREW CLARKE & NADINE M. JOHNSTON

Phylum Cnidaria Cnidarians comprise a varied collection of organisms united by their possession of nematocysts, and containing many difﬁcult groups. The ﬁve classes are generally grouped into two subphyla, the Medusozoa containing the Hydrozoa, Scyphozoa and Cubozoa, and the Anthozoa containing the Alcyonaria and Zoantharia. The life cycle typically contains both a benthic and a pelagic phase. Whereas in anthozoans the larval phase is usually a simple planula larva and the medusa stage lost from the life cycle, in medusozoans the planktonic phase can be dominant (as in jellyﬁsh). This can make the decision as to what constitutes a benthic taxon somewhat arbitrary. Within the Hydrozoa, the dominant stage in the hydroid life cycle is typically benthic; for hydromedusans or leptomedusans, the distinction between truly benthic or plankton taxa is more difﬁcult. The Scyphozoa are predominantly pelagic and are not considered further here. The Cubozoa (box-jellyﬁsh) are pelagic and exclusively tropical. Within the Medusozoa, the most recent synopsis for Anthomedusae, Leptomedusae, Limnomedusae and Narcomedusae is O’Sullivan (1982) and for Stylasterina it is Cairns (1983). Subsequent references consulted were Browne (1902), Kramp (1957), Blanco & Belluscí de Miralles (1972), Blanco (1977, 1984), El Beshbeeshy (1991), Jarms & Tiemann (1996), Peña Cantero (1997a,b, 1998a,b,c), Peña Cantero & Garcia Carrascosa (1991), Peña Cantero & Vervoort (1995, 1998), Peña Cantero et al. (1995, 1996, 1997a,b,c). For Stylasterina there were no references after Cairns (1983) containing new species, or species not previously recorded from the Southern Ocean. As the hydromedusan groups were only reviewed in part by O’Sullivan (1982), a thorough literature search was undertaken. The systematics used here is that of O’Sullivan (1982) for all except the Stylasterina, for which we used Cairns (1983). For Anthozoa the most recent synopses of the Antarctic fauna have been partial reviews of the gorgonians by Molander (1929), Thomson & Rennet (1931), Broch (1965), Dell (1972) and Bayer & Stefani (1987), a complete review of scleractinians by Cairns (1990), partial reviews of the actinarians by Dell (1972) and Dunn (1983, 1984), and a review of pennatularians by Broch (1959). References consulted in addition to these synopses were Pasternak (1961, 1975, 1993), Bayer (1980, 1996a,b), Williams (1981), Thomas & Mathew (1986a,b) and Keller (1990). The intermediate level systematics (order, family) of the Anthozoa used here follows that of the synopses listed above, with the addition of Pasternak (1975) for gorgonians. Cairns (1990) and Dunn (1983, 1984) have developed valuable keys for scleractinians and actinarians. For both hydrozoans and anthozoans, the key period of taxonomic work was from the 1870s to the mid-1920s as the material collected during the early expeditions was worked up (Fig. 3a). Since then there has been a signiﬁcant increase in taxonomic work in the 1990s, associated principally with the SCAR (Scientiﬁc Committee on Antarctic Research) programme on Ecology of the Antarctic Sea Ice Zone (EASIZ). The Southern Ocean cnidarian fauna is badly in need of a thorough taxonomic overhaul. This, coupled with investigation of cryptic species using modern molecular techniques, would almost certainly result in a drastic revision of Antarctic cnidarian diversity. The distribution of species to genus values is normal, though there is a notable outlier for the hydroid genus Oswaldella (Fig. 3b). Three hydrozoan and one alcyonarian anthozoan family contain more than ten species (Table 4).

60

ANTARCTIC MARINE BENTHIC DIVERSITY

(a)

(b)

Figure 3 Cnidaria (Anthozoa and benthic Medusozoa). (a) Time course of description of Southern Ocean benthic cnidarian taxa; data presented in percentage terms (total 272 species). (b) Distribution of species amongst genera for benthic cnidarians. Data presented as a frequency distribution of species per genus values. In both plots data for Anthozoa and benthic Medusozoa are shown separately.

Table 4 The most speciose families of Southern Ocean cnidarians. The threshold for inclusion was ten species, and species numbers can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. The higher level taxonomy follows Barnes (1998) and within classes, families are listed in order of richness. Subphylum

Class

Family

Approximate number of species

Medusozoa

Hydrozoa

Anthozoa

Alcyonaria

Sertulariidae Plumulariidae Stylasteridae Haleciidae Primnoidae

42 32 13 10 28

Phylum Ctenophora Although ctenophores are almost exclusively pelagic, a few species are benthic and one large benthic species, Lyrocteis ﬂavopallidus, has been described from McMurdo Sound (Robilliard & Dayton 1972). These are delicate organisms, easily damaged by trawling and dredging, so any subsequent reports will likely come from photographic or direct SCUBA observation.

61

ANDREW CLARKE & NADINE M. JOHNSTON

Phylum Priapula This tiny phylum has only three representative in the Southern Ocean. The most recent synopsis is that of van der Land (1970), whose systematics we follow. Recent references consulted were Murina (1975) and Wu & Zhao (1986).

Phylum Crustacea The arthropods have frequently been regarded as a single phylum comprising the insects and their close relatives, arachnids and crustaceans. Here we follow Barnes (1998) in regarding crustaceans, chelicerates (arachnids with pycnogonids) and uniramians (insects and relatives) as separate phyla, grouped with onychophorans and possibly tardigrades and pentastomans in the superphylum Panarthropoda. The crustaceans are now one of the best-known marine invertebrate groups in Antarctica. We will therefore deal with the major orders individually.

Order Amphipoda (Class Malacostraca) The peracarid crustaceans are among the most intensely studied marine taxa in the Southern Ocean. The amphipods have been thoroughly revised by De Broyer & Jazdzewski (1993). More recently De Broyer & Jazdzewski have collected new material from the Weddell Sea in particular, and this has added significantly to the previously published list. These studies have also shown that there are likely to be many more new species to be described from the very small elements of the fauna, previously under-sampled and poorly known. The most recent estimate of benthic gammaridean species richness for waters south of the Subtropical Front is 692 (De Broyer & Jazdzewski 1996). This is not directly comparable with the data analysed here, since our compilation is for the Southern Ocean sensu stricto, as deﬁned by the Polar Front. We have therefore excluded species found only at Tristan da Cunha, Gough Island, Prince Edward and Marion Islands, Îles Crozet, Îles Kerguelen, the Magellan region of South America and the sub-Antarctic islands of New Zealand. Recent literature we have included in our compilation is De Broyer (1985a,b), Rauschert & Andres (1993), Coleman (1994), Coleman et al. (1994), De Pina (1995), Jazdzewski et al. (1995), Wakabara et al. (1995) and Kunzmann (1996). The systematics of the Amphipoda used here is generally that of De Broyer & Jazdzewski (1993). For the Gammarellidae we have followed Barnard (1969) and Barnard & Karaman (1991), for the Valettidae, Thurston (1989), for the iphimedid group, Coleman & Barnard (1991), and for the Carenioidae, retained in the Synopiidae, Jazdzewski & De Broyer (1990). For the Orchomene complex, still under revision, De Broyer (1984, 1985a) has been followed. For the Caprellidae, we have followed Laubitz (1993), and for the Clarenciidae, Zeidler (1994). The time-course of species description (Fig. 4a) indicates an important phase of taxonomic work in the early decades of the last century, but also that ongoing work (most notably by Belgian and Polish taxonomists) is continuing to add new species. There is no indication that we have reached an asymptote in the description of new amphipod taxa for 62

ANTARCTIC MARINE BENTHIC DIVERSITY

(a)

(b)

Figure 4 Amphipoda. (a) Time course of description of Southern Ocean amphipod taxa; data presented in percentage terms (total 496 species). (b) Distribution of species amongst genera for amphipods, presented as a frequency distribution of species per genus values.

the Southern Ocean and the true species richness is very likely signiﬁcantly greater than the current ﬁgure. The frequency distribution of species per genus values (Fig. 4b) exhibits a normal pattern, and there are eight families with over 20 species (Table 5).

Table 5 The most speciose families of Southern Ocean crustaceans. The threshold for inclusion was 20 species, and species richness can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. The higher level taxonomy follows Barnes (1998) and, within orders, families are listed in order of richness. Class

Subclass

Order

Family

Amphipoda

Lysianassoidae Eusiridae Stenothoidae Iphimediidae Corophiidae Epimeriidae Isochyroceridae Phoxocephalidae Arcturidae Munnopsidae Serolidae Anarthruridae Scalpellidae

Approximate number of species

Malacostraca Peracarida

Isopoda

Cirripedia

Tanaidacea Thoracica

63

92 64 50 36 22 21 20 20 60 40 31 39 34

ANDREW CLARKE & NADINE M. JOHNSTON

Order Isopoda (Class Malacostraca) The isopod fauna of the Southern Ocean is now well known particularly through the extensive taxonomic work of Brandt (1988) and Wägele (1989). Key recent references are Brandt (1991b, 1992a,b, 1999), Brandt & Janssen (1994), Kunzman (1996), and Pirez & Sumida (1997). For systematics we have followed that used by Brandt (1991a) in her review of Southern Ocean isopods; this follows Wägele (1989), except where new genera have been erected. The time-course of species description for isopods (Fig. 5a) shows the importance of descriptive work in the ﬁrst two decades of the last century, based on material collected in the early expeditions, but also the continuing description of new taxa. As with amphipods there is no indication of an asymptote in species description for the Southern Ocean. The frequency distribution of species per genus values shows a typical hollow curve shape (Fig. 5b). There are three speciose families (Table 5), two of which (Serolidae and Arcturidae) have clearly radiated within the Southern Ocean (Brandt 1991a).

Order Tanaidacea (Class Malacostraca) The most recent synopses of the Southern Ocean tanaid fauna are those of Sieg (1983, 1984a,b, 1986a,b,c). Also consulted was Blazewicz & Jazdzewski (1996). The systematics followed Sieg (1986c, 1988). The inﬂuence of this recent attention is shown in the shape of the species accumulation curve (Fig. 5a). There are no unusual features in the frequency distribution of species per genus values (Fig. 5b), and only one family, Anarthruridae, contains more than 20 species (Table 5). (a)

(b)

Figure 5 Isopoda and Tanaidacea. (a) Time course of description of Southern Ocean isopod taxa; data presented in percentage terms (total 257 species) with data for Isopoda and Tanaidacea shown separately. (b) Distribution of species amongst genera for isopods and tanaids. Data presented as a frequency distribution of species per genus values, with data for Isopoda and Tanaidacea shown separately.

64

ANTARCTIC MARINE BENTHIC DIVERSITY

Order Decapoda (Class Malacostraca) The eucarid crustaceans are strikingly low in diversity in the Southern Ocean. The most recent review of the Southern Ocean decapod fauna is that by Kirkwood (1984). More recent references consulted were Tiefenbacher (1990) and Klages et al. (1995); the systematics follows Kirkwood (1984). The total fauna numbers only a dozen or so species, including six lithodid crabs from sub-Antarctic waters. Brachyuran crabs and lobsters are now completely absent from the Southern Ocean, although fossil evidence indicates a rich fauna which has since become extinct (Feldmann & Wilson 1988, Feldmann & Tshudy 1989, Feldmann & Quilty 1997, Feldmann & Crame 1998)

Class Cirripedia For cirripedes the most recent synopsis is the partial review of Newman & Ross (1971). Since this latter review does not cover the complete fauna, the primary literature was searched as far back as possible. Key references were Hoek (1907), Gruvel (1907, 1910), Borradaile (1916), Nilsson-Cantell (1930a,b, 1939), Zevina (1964), Utinomi (1965), Zevina (1968), Grygier (1981, 1984, 1987), Kuznetsova & Neurova (1986), Ren & Huang (1989), Zevina (1990) and Young & Leta (1996). The systematics used here follows Newman & Ross (1971). Ice-scour means that the intertidal barnacles so characteristic of temperate and tropical rocky shores are completely absent from the Southern Ocean. The total fauna is about 50 species, with only one family (Scalpellidae) containing more than 20 species (Table 5). It also includes the ascothoracians, parasitic in starﬁsh.

Phylum Chelicerata The chelicerates comprise the horseshoe crabs (merostomatans), the arachnids (spiders, mites, scorpions and allies) and the pycnogonids (also referred to as pantopods). There is no consistent view as to whether the pycnogonids should be included within the chelicerates, and they are sometimes regarded as a phylum in their own right (Minelli 1993). Here we follow Barnes (1998) in regarding them as a class within the chelicerates. There are no horseshoe crabs in the Southern Ocean. Marine mites appear to occur widely, but we could locate no taxonomic review although Pugh (1993) lists 45 species of the family Rhagidiidae from a variety of Southern Ocean locations.

Class Pycnogonida Although the phylogenetic status of this enigmatic group is not at all clear, they have most frequently been regarded as a relatively ancient lineage of marine chelicerates (Nielsen 1995). The most recent taxonomic revision are those of Munilla León (2001a), Child (1994a,b, 1995a,b,c) and the partial review of Fry & Hedgpeth (1969). More recent references consulted were Arnaud (1972a,b), Turpaeva (1974), Pushkin (1976), Child (1987), Munilla León (1989, 2001b), Turpaeva (1990), Stiboy-Risch (1992), Jaya Sree et al. (1993), Bamber 65

ANDREW CLARKE & NADINE M. JOHNSTON

(1995) and Kunzmann (1996). Unfortunately, Fry & Hedgpeth (1969) never completed their synopsis of the whole Antarctic pycnogonid fauna, and only one part was ever published. For systematics we follow Fry & Hedgpeth (1969), who commented that . . . the unsatisfactory state of the taxonomy of the higher taxa of the Pycnogonida has prevented us from attempting to devise keys to genera or families, which, in their layout reproduce the classiﬁcation of the Pycnogonida, and throughout this work the keys are entirely artiﬁcial in their design. Our compilation of the Southern Ocean pycnogonid fauna totals 175 species, in comparison with the 180 reported for the Antarctic plus sub-Antarctic by Munilla León (2001a). This discrepancy is related in part to a difference in the deﬁnition of the Southern Ocean, and partly to recent description of new taxa. The species accumulation curves devised from our data and the decadal summary provided by Munilla León (2001a) are broadly similar, though the recent description of new taxa is apparent in the different trajectories in the upper third of the two curves (Fig. 6a). The shape of the frequency distribution of species per genus values is highly skewed and the data are unusually patchy (Fig. 6b). This, together with the striking outliers for Nymphon (55 species), Colossendeis (29), Ammothea (19) and Pallenopsis (17) suggests an immature taxonomy. The large number of species currently assigned to the genus Nymphon indicates either that this represents a clade in the process of active speciation, or that different taxonomic characters are required. This genus would seem to be a prime candidate for cladistic and/or molecular phylogenetics work. Four families contain more than 20 species: Colossendeidae (33), Ammotheidae (39), Nymphonidae (58) and Callipallenidae (29).

(a)

(b)

Figure 6 Pycnogonida. (a) Time course of description of Southern Ocean pycnogonid taxa; data presented in percentage terms (total 175 species). (b) Distribution of species amongst genera for pycnogonids, presented as a frequency distribution of species per genus values.

66

ANTARCTIC MARINE BENTHIC DIVERSITY

Phylum Nemertea Benthic nemertean worms are relatively well known from the Southern Ocean, with the most recent taxonomic revisions being that of Gibson (1985) for the heteronemerteans and Dawson (1957) for the hoplonemerteans, both of whose systematics we follow here. We have been unable to locate any subsequent taxonomic or faunal work on Southern Ocean benthic nemerteans. The overall diversity of nemertean is low at just over 30 species divided more or less evenly between the heteronemerteans (Class Anopla) and the hoplonemerteans (Class Enopla). The Southern Ocean is, however, unusual in that one very large species, Parborlasia corrugatus, is extremely common in shallow waters, where it is a major predator and scavenger (Gibson 1983).

Phylum Mollusca The Mollusca are a well-described group in most seas of the world, and the Southern Ocean bivalve and gastropod fauna is fairly well known. Although the various molluscan classes differ strongly in species richness, many taxonomic monographs treat some or all of them together. The major taxonomic synopses are those of Powell (1960) and Dell (1990), although the latter only covers species found in the Ross Sea. Taxonomic works since Powell’s review are Nicol (1966a), Ponder (1971, 1983), Arnaud (1972a,b, 1974), Arnaud & van Mol (1979), Arnaud et al. (1986), Cantera & Arnaud (1984), Warén et al. (1986), Dell (1964a,b, 1972), Powell (1973), Egorova (1982), Oliver & Picken (1984), Mühlenhardt-Siegel (1989), Numanami et al. (1996), Hain (1990), Voss (1988), Numanami (1996) and Cattaneo-Vietti et al. (2000). We have also cross-checked our species list with that for the most recent comprehensive benthic sampling in the Weddell Sea (Gutt et al. 2000). The systematics used here follows Dell (1990) and Numanami et al. (1996), with Hain (1990) for new species. Dell (1990) does not cite Hain (1990), and vice versa; as Dell (1990) is the larger and more complete work the allocation of genera to families generally follows Dell (1990). A few additions follow Numanami et al. (1996). The phylogeny of gastropod and bivalve molluscs is currently an area of active research and the results of molecular and cladistic analyses are leading to fundamental revisions of molluscan taxonomy (Ponder & Lindberg 1996, 1997). Changes to intermediate level taxonomy (orders, families) will have no immediate effect on diversity measures at the species level but they are important in generating evolutionary hypotheses. Continuing systematic work reﬁning the composition of families and genera will, however, affect measures such as species to genus ratios or family richness. We can expect a period of continuing change in this area and the data presented here can be regarded only as an interim picture. The time-course of species description for gastropods and bivalves is similar, with a major period of systematic work in the ﬁrst two decades of the last century (Fig. 7a). There were signiﬁcant additions to both the gastropod and bivalve faunas from systematic work in the ﬁnal decade of the last century. Overall the bivalve and gastropod fauna of the Southern Ocean is probably described more completely than for any other major group of benthic marine invertebrates. The frequency distribution of species per genus values was normal, although there were a number of outliers for gastropod genera containing very high numbers of species (Fig. 7b). 67

ANDREW CLARKE & NADINE M. JOHNSTON

(a)

(b)

Figure 7 Gastropod and bivalve molluscs. (a) Time course of description of Southern Ocean gastropod and bivalve taxa; data presented in percentage terms (total 640 species) with data for gastropods and bivalves shown separately. (b) Distribution of species amongst genera for molluscs. Data presented as a frequency distribution of species per genus values, with data for gastropods and bivalves shown separately.

These included Prosipho (37 species), Trophon (22), Onoba (17) and Eatoniella (15). The most species-rich families of gastropods include the predatory and scavenging buccinids, naticids and muricids (Table 6), reﬂecting the radiation of these groups at high latitudes (Crame 1996). Other molluscan groups are not well known in the Southern Ocean. The primitive Solenogastres are an obscure group often placed with other shell-less molluscs in the Aplacophora. The few Antarctic Solenogastres have been reviewed by von Salvini-Plawen (1979), with new species reported by García-Alvarez et al. (2000, 2001). Scaphopoda (tuskshells) appear to be represented by fewer than ten species (Dell 1964b, Steiner & Linse 2000, Katrin Linse, pers. comm.); this is a fairly typical diversity for a group that is nowhere very speciose. Polyplacophora (chitons) appear to be similarly lacking in diversity in the Southern Ocean. A taxonomic review is badly needed but it is likely that the total fauna amounts to fewer than a dozen species. The benthic octopods have recently been examined in detail by Allcock (1997), who reports 34 species from benthic samples taken in the Weddell Sea and the Antarctic Peninsula. Of these some 18 are new to science and await formal description. The most striking feature is the number of species of Pareledone, a genus which appears to have undergone a radiation in the Southern Ocean (Allcock et al. 2001).

Phylum Sipuncula This is a small but relatively well-described phylum. The most recent Antarctic monograph is Saiz-Salinas (1995), but little work has been undertaken since. Currently, 16 species are 68

ANTARCTIC MARINE BENTHIC DIVERSITY

Table 6 The most speciose families of Southern Ocean gastropod and bivalve molluscs. The threshold for inclusion was ten species, and species richness can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. Gastropod taxonomy in particular is in a state of ﬂux with families being renamed and subdivided. Families are listed within classes in order of richness. Class

Family

Gastropoda

Buccinidae Turridae Trochidae Naticidae Rissoidae Muricidae Cerithidae Eatoniellidae Diaphanidae Cyclostrematidae Littorinidae Lamellaridae (Velutinidae) Philinidae Eulimidae Cancellaridae Nuculanidae Phylobryidae Pectinidae

Bivalvia

Approximate number of species 84 45 41 29 24 23 22 15 15 15 14 12 11 10 10 13 13 11

recognised from three genera in two families: Golﬁngiidae (Golﬁngia and Nephasoma) and Phascolidae (Phascolion).

Phylum Echiura This small phylum has largely been ignored in the Southern Ocean until the recent analysis of the extensive collections made by American research vessels (notably RV ELTANIN) between 1962 and 1986 and processed by the Smithsonian Oceanographic Sorting Centre (Saiz-Salinas et al. 2000). This collection of 855 individuals yielded nine species representing two families. The Echiuridae prefer sublittoral to shallow bathyal sediments, whereas the Bonellidae are more common at bathyal and abyssal depths.

Phylum Annelida Although this phylum contains several groups known to occur in the Southern Ocean, including the oligochaetes and leeches (grouped together in the Class Clitellata by Barnes 1998), only the polychaetes have received sufﬁcient attention for a preliminary species list to be assembled.

69

ANDREW CLARKE & NADINE M. JOHNSTON

Class Polychaeta The major taxonomic monographs for Antarctic polychaetes are those of Hartman (1964, 1966, 1967, 1978), although no single taxonomic revision for the whole group yet exists. More recent publications consulted in the compilation of the species lists were Benham (1921, 1927), Monro (1930, 1935, 1939), Levenstein (1964), Averintsev (1972), Sicinski (1986), Hopkins (1987), Hartmann-Schröder & Rosenfeldt (1988, 1989, 1990, 1992), Orensanz (1990), Vinogradova (1990), Ahn & Kang (1991), Gambi & Mazzella (1992), Cantone & Sanﬁllipo (1991), Castelli (1992), Gambi et al. (1994, 1997), Herman & Dahms (1992), Kudenov (1993), Sicinski & Janowska (1993), Cantone (1995), Stiller (1996), Kunzmann (1996), San Martin & Parapar (1997) and Cantone & Di Pietro (1998, 2001). For polychaete higher level systematics we have followed Fauchald (1977). Fauchald & Rouse (1997) emphasise that all polychaete taxonomies in current use are unsatisfactory for a variety of reasons, although the major source of uncertainty is the lack of consistent morphological information. Many polychaetes are difﬁcult to identify, and Knox & Lowry (1977) have suggested that the Southern Ocean polychaete fauna may exceed 800 species. Certainly the time-course of species description (Fig. 8a) gives us no indication of any slowing down and it is quite possible that the Southern Ocean may contain over 1000 polychaete taxa. Recent benthic sampling of soft sediments from the deeper waters of the continental shelf of the western Antarctic Peninsula has revealed many new undescribed taxa (Craig Smith, pers. comm.). A thorough taxonomic revision of the Southern Ocean polychate fauna is needed urgently, though this would be a major task. The frequency distribution of species to genus values is classical in shape (Fig. 8b), with over 100 genera containing only single species and a notable outlier for Harmathoe (21 species). As with elsewhere in the world, syllids and polynoids are the most speciose families (Table 7).

(a)

(b)

Figure 8 Polychaetes. (a) Time course of description of Southern Ocean polychaete taxa; data presented in percentage terms (total 645 species). (b) Distribution of species amongst genera for polychaetes, presented as a frequency distribution of species per genus values.

70

ANTARCTIC MARINE BENTHIC DIVERSITY

Table 7 The most speciose families of Southern Ocean polychaetes. The threshold for inclusion was 20 species, and species richness can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. Families are listed in order of richness. Family

Approximate number of species

Syllidae Polynoidae Terebellidae Phyllodocidae Sabellidae Onuphidae Ampharetidae Maldanidae Flabelligeridae Opheliidae Serpulidae

67 66 46 37 32 29 27 27 26 23 21

Phylum Pogonophora Although sometimes grouped with annelids (for example by Nielsen 1995), here we follow Barnes (1998) in assigning them to a separate phylum. To date only three species have been described from the Southern Ocean, all in the genus Spirobrachia (Smirnov 2000). It is possible that more taxa remain to be described from the extensive deep-sea collections made by scientists of the former Soviet Union. Although no vestimentiferans have yet been described from the Southern Ocean, it is now known that hydrothermal vents and related features are present (Chin et al. 1996, German et al. 2000, Klinkhammer et al. 2001) and so they may remain to be discovered. If vent faunas are located, their biogeographic afﬁnities will be of great interest because of their intermediate location between the very different vent faunas of the Atlantic and Paciﬁc basins.

Phylum Brachiopoda Brachiopods are a group of sessile suspension feeders which once dominated the seas. The rise of the bivalve molluscs has been associated with a decrease in the importance of brachiopods, which nowadays are conﬁned largely to polar, deep-sea and cave habitats. The most recent taxonomic revisions of Southern Ocean brachiopods are those of Foster (1965, 1974). Recent references consulted were Foster (1989, 1997) and Sieg & Wägele (1990). The systematics of brachiopods is in a state of ﬂux, particularly with the current revision of the Treatise of Invertebrate Palaeontology. Specialists are revising each group, and the assignment of genera to families is likely to change radically (and indeed has already done so in comparison with the 1965 edition of the Treatise). We have followed Foster (1965), accepting that this is likely to change in the future. The Southern Ocean brachiopod fauna is small, with 19 species described to date. Two are inarticulates (Pelagodiscus and Crania), with the most important articulate genus being Liothyrella. 71

ANDREW CLARKE & NADINE M. JOHNSTON

Phylum Bryozoa Bryozoans are now a relatively well-known group in the Southern Ocean, with a recent and very thorough taxonomic revision of the Cheilostomatida by Hayward (1995). The most recent synopses of the Cyclostomatida and Ctenostomatida were those of Androsova (1968). Recent references consulted are Androsova (1972), Moyano (1984), Rosso (1990), Winston (1994), Hayward (1996) and Ostrovskii & Taylor (1996). For cheilostome systematics we have followed Hayward (1995). There appears to be no generally accepted higher level classiﬁcation for ctenostome or cyclostome bryozoans. Hayward (pers. comm.) regards the current species list for Southern Ocean ctenostome bryozoans as a severe underestimate of the true species richness, and the cyclostome species richness may also be underestimated. The ctenostomes are the most ancient bryozoan lineage. They are rare in the Southern Ocean and almost absent from shallow waters (Barnes & De Grave 2000). Only six species have been reported from Antarctic waters so far, although the taxonomy of Alcyonidium is particularly complex and difﬁcult so this list may rise. The relatively low diversity of cyclostomes is not unusual for isolated areas (Barnes & De Grave 2000) and the bryozoan fauna overall is dominated by cheilostomes (Table 8). The time-course of species description (Fig. 9a) shows two important periods of taxonomic work, in the ﬁrst two decades of the last century, and in the past twenty years. The frequency distribution of species per genus values is normal (Fig. 9b), with the highest value being for Cellarinella. Over a dozen cheilostome families contain ten or more species, with the most diverse being Smittinidae (Table 8). Table 8 The most speciose families of Southern Ocean bryozoans. The threshold for inclusion was ten species, and species richness can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. The absence of an agreed intermediate level taxonomy for cyclostomes and ctenostomes means that species richness can only be presented at the order level. Ctenostomes are not very diverse in the Southern Ocean, but are included for completeness. Higher level taxonomy follows Barnes (1998) and Hayward (1995). Class

Order

Stenolaemata

Cyclostomatida Ctenostomatida Cheilostomatida

Gymnolaemata

Family

Approximate number of species

Smittinidae Bugulidae Cabereidae Cellariidae Sclerodomidae Exochellidae Celleporidae Microporellidae Flustridae Calloporidae Chaperidae Arachnopusidae Phidoloporidae

72

67 6 34 19 18 17 17 14 14 12 11 11 10 10 10

ANTARCTIC MARINE BENTHIC DIVERSITY

(a)

(b)

Figure 9 Bryozoa. (a) Time course of description of Southern Ocean bryozoan taxa; data presented in percentage terms (total 322 species). (b) Distribution of species amongst genera for Southern Ocean bryozoans, presented as a frequency distribution of species per genus values.

Phylum Echinodermata Echinoderms have been collected by almost every biological expedition to the Southern Ocean, and typically being large and conspicuous are generally well described. The major taxonomic synopses used for asteroids were Koehler (1920), Fisher (1940), Clark (1962) and McKnight (1976). For systematics we have followed Clark (1962). The key references and synopses for echinoids were Koehler (1926), Grieg (1929) and Mortensen (1936). Additional records came from Retamal et al. (1983), Larrain (1985), Voss (1988) and Mironov (1995). The key ophiuroid references were Koehler (1922), Madsen (1967), Alarcon-Castillo (1967) and Seno & Irimura (1968). Additional records came from Bernasconi & D’Agostino (1973a,b). The key references used in compiling the holothurian species list were Agatep (1967), Cherbonnier (1973), Pawson (1977) and Gutt (1988). For crinoids the key references used were Clark (1937), John (1938) and Speel & Dearborn (1983), with the latter references used for systematics. General echinoderm references consulted were Arnaud (1964), McKnight (1967), Cherbonnier & Guille (1974) and Piepenburg et al. (1997). The time-course of species description (Fig. 10a) shows that over 80% of the known fauna had been described by the middle of the last century. This reﬂects the generally large and conspicuous nature of echinoderms, which together with molluscs and crustaceans tend to be among the ﬁrst groups to be tackled by systematists. The frequency distribution of species to genus values (Fig. 10b) is normal, with over 80 genera containing only a single described species, and two outliers with high species numbers (the ophiuroid genera Amphiura and Ophiura). The most speciose families are spread across all ﬁve echinoderm classes for which we have species lists in the Southern Ocean (Table 9). 73

ANDREW CLARKE & NADINE M. JOHNSTON

(a)

(b)

Figure 10 Echinoderms. (a) Time course of description of Southern Ocean echinoderm taxa; data presented in percentage terms (total 410 species). (b) Distribution of species amongst genera for all echinoderms combined, presented as a frequency distribution of species per genus values.

Table 9 The most speciose families of Southern Ocean echinoderms. The threshold for inclusion was ten species, and species richness can only be approximate because of unresolved taxonomic difﬁculties and undiscovered species. The higher level taxonomy follows Barnes (1998). Families are listed within classes in order of species richness. Class

Family

Crinoidea

Antedonidae Asteriidae Odontasteridae Pterasteridae Ophiolepidae Ophiacanthidae Amphiuridae Schizasteridae Cidaridae Psolidae Elpidiidae Cucumariidae Paracucumidae Synallactidae

Asteroidea Ophiuroidea

Echinoidea Holothuroidea

Approximate number of species 22 37 11 10 69 20 20 22 15 19 17 15 15 13

74

ANTARCTIC MARINE BENTHIC DIVERSITY

Phylum Chordata This phylum comprises three major groups which are sometimes elevated to phyla in themselves, the urochordates (ascidians, larvaceans and salps), cephalochordates (lancelets) and vertebrates. Here we follow Barnes (1998) in regarding them as subphyla. There are no cephalochordates known from the Southern Ocean, and amongst the urochordates only the ascidians are benthic.

Subphylum Urochordata, Class Ascidiacea Ascidians are a conspicuous and ecologically important component of the continental shelf fauna. The most recent taxonomic revisions are those of Kott (1969) and Monniot & Monniot (1983, 1994). Recent references consulted were Monniot (1990) and Sreepada et al. (1995). For systematics we have followed Monniot & Monniot (1983), itself a modiﬁcation of Kott (1969). The time-course of species description (Fig. 11a) shows that over half the known Southern Ocean ascidians had been described by 1920. Since then new species have been described only slowly, with a small pulse in the latter part of the last century. The frequency distribution of species per genus values (Fig. 11b) is fairly typical apart from a single striking outlier caused by the genus Aplidium with 25 species. The four most speciose families are the Polyclinidae (37 species in 6 genera), Stylidae (19 species in 6 genera), Molgulidae (16 species in 4 genera) and Pyruidae (15 species in 6 genera). (a)

(b)

Figure 11 Ascidians. (a) Time course of description of Southern Ocean ascidian taxa; data presented in percentage terms (total 118 species). (b) Distribution of species amongst genera for Southern Ocean ascidians, presented as a frequency distribution of species per genus values.

75

ANDREW CLARKE & NADINE M. JOHNSTON

Subphylum Vertebrata The vertebrates are traditionally classiﬁed into seven major groups, here regarded as classes, of which three are referred to colloquially as ﬁshes. These are the Agnatha (hagﬁshes and lampreys), Chondrichthyes (cartilaginous ﬁshes, comprising sharks, rays, skates and their allies) and Osteichthyes (bony ﬁshes, a group which includes the dominant ﬁshes living today, the teleosts). All three ﬁsh classes are represented in the Southern Ocean though their mobility means that they are difﬁcult to classify as truly benthic organisms. Of the other four classes, there are no amphibians or reptiles known from the Southern Ocean but both birds and mammals are important components of the oceanic food-web. Although some seabirds and marine mammals feed on or close to the sea bed, neither group can be regarded as benthic organisms. Table 10 Dominant taxa (families) of benthic ﬁshes inhabiting the continental shelf and upper continental slope of the Southern Ocean. The phylogenetic sequence is from Gon & Heemstra (1990). Table based on Eastman (1993) and Eastman & Clarke (1998). The liparids and six notothenioid families (*) together comprise 85% of all Southern Ocean species. Class

Order

Family

Number of species

Agnatha Myxiniformes Myxinidae (hagﬁshes) Petromyzontiformes Petromyzontidae (lampreys) Chondrichthyes Rajiformes Rajidae (skates) Osteichthyes Ophidiiformes Carapidae (pearlﬁshes) Gadiformes Moridae (deepsea cods) Muraenolepididae (eel cods) Gadidae (true cods) Scorpaeniformes Congiopodidae (horseﬁshes) Liparidae (snailﬁshes) Zoarcidae (eelpouts) Perciformes Bovichtidae (thornﬁshes)* Nototheniidae (Antarctic cods or notothens)* Artedidraconidae (plunderﬁshes)* Harpagiferidae (spiny plunderﬁshes)* Bathydracondiae (dragonﬁshes)* Channichthyidae (iceﬁshes)* Tripterygiidae (tripleﬁns) Pleuronectiformes Achiropsettidae (southern ﬂounders)

1 1

8

1 4 4 1 1 64 23 1 34 24 6 15 15 1 4

76

ANTARCTIC MARINE BENTHIC DIVERSITY

The Southern Ocean ﬁsh fauna is well known. The most recent taxonomic monographs are by Gon & Heemstra (1990) and Miller (1993). Since the publication of these volumes, Russian ichthyologists have described several new species of nototheniids and many new liparids. For systematics we have followed Eastman (1993), as updated by Eastman & Clarke (1998); this is itself based on Gon & Heemstra (1990). The Southern Ocean ﬁsh fauna is relatively low in diversity (Table 10). It is also unusual in containing by two striking radiations, the nototheniids on the continental shelf and the liparids on the continental slope (Eastman & Clarke 1998). The continental shelf ﬁsh fauna is dominated by nototheniids, not only in terms of diversity but more particularly in biomass: in a typical single haul over 95% of individuals can be nototheniids. This level of dominance by a single clade of ﬁshes is unique in the sea, and the radiation of nototheniids exhibits many parallels with the species ﬂocks of freshwater ﬁshes in some lakes (Eastman & Clarke 1998). The radiation of liparid ﬁshes on the continental slope is less well documented, partly because this is a very difﬁcult group taxonomically but also because samples are fewer. Much of what we do know is the result of extensive work by Russian scientists.

Discussion How many benthic species are there in Antarctica? A long history of collection and taxonomic work in the Southern Ocean has meant that the benthic fauna is better known than might be thought. Some groups inevitably are in urgent need of taxonomic attention, but we are in a position to make some useful generalisations. A summary diversity inventory for Southern Ocean benthos is shown in Table 11. Although most macrobenthic taxa are listed, this cannot be regarded as an all-taxon biodiversity inventory (ATBI) of the type widely promulgated following the publication of the Convention on Biological Diversity (Yoon 1993). This is because many important classes of fauna are excluded (notably the meiofauna and hyperbenthos), and no plants, protistans or viruses are listed. Because the benthos is linked intimately to the overlying water column, whereas the nekton are able to move freely into deeper water, it becomes very difﬁcult to decide on a meaningful circumscription of an area or habitat within which to attempt to construct an ATBI. Recently a brave attempt has been made for Hawaii (Eldredge & Miller 1995, updated by Paulay 1997), including all plant and animal taxa, and encompassing marine, freshwater and terrestrial habitats. The total species list for the Southern Ocean benthos currently exceeds 4100 (Table 11, see also Addendum, p. 114). The most species-rich group is the polychaete worms. This group is important in soft substrata, but most of the remaining taxa are important components of the epifauna. Of particular signiﬁcance is the strong representation of suspension feeding taxa (bryozoans, sponges, hydrozoans, ascidians and anthozoans). Ranking by species richness thus emphasises the comments of many ecologists as to the predominance of suspension feeding communities in Antarctica. To some extent, however, this may represent under-sampling of soft substrata. The overall diversity of the Southern Ocean will also be inﬂuenced by the affect of ice scour on rocky intertidal and subtidal habitats, and the almost complete absence of rivers, estuaries and intertidal mud-ﬂats; all of which habitats typically contain rich and diverse benthic communities elsewhere in the world. 77

ANDREW CLARKE & NADINE M. JOHNSTON

Table 11 The best available estimates of higher taxon species richness for free-living Southern Ocean benthos. The tabulated data are from this study and from a review by Arntz et al. (1997). The reliability of the estimated richness for each taxon is classiﬁed on a subjective three point scale ranging from A (good) to C (poor) as deﬁned in the text (see Methods section); also provided is the percentage

Phylum

Subphylum or Class

Order

Porifera Symplasma Cnidaria Medusozoa Anthozoa Ctenophora Platyhelminthes Gastrotricha Priapula Kinorhyncha Nematoda Tardigrada Arthropoda

Turbellaria

Crustacea Malacostraca

Amphipoda Isopoda Tanaidacea Decapoda

Cirripedia Ostracoda

Estimated species richness

Reliability

Arntz

This study

This study

Winston

⬃300 –

250 29

B A

50

⬃200 85 nd nd

186 86 nd nd

C C C C

90 50–95

3 nd nd nd

3 nd nd nd

A C C C

520 346 50 (19)

496 257 80 13

B A B A

37 nd

50 nd

C C

nd 150 nd

45 175 31

C A B

20

C A

40–50

⬃15 nd

10 530 110 6 ⬃34 15 9

650

645

C

nd nd nd 16 310

nd 3 nd 19 322

C C C B A

nd

nd

C

22 nd nd 44 88

28 108 119 49 106

B B B B B

130 nd nd nd

118 2 8 198

B A A A

Chelicerata Arachnida Pycnogonida

Acarina

Nemertea Mollusca Polyplacophora Gastropoda Bivalvia Scaphopoda Cephalopoda

nd nd

Sipuncula Echiura Annelida Polychaeta Clitellata Pogonophora Entoprocta Brachiopoda Bryozoa Hemichordata Echinodermata Crinoidea Asteroidea Ophiuroidea Echinoidea Holothuroidea Chordata Urochordata Vertebrata

Agnatha Chondrichthyes Osteichthyes

78

C B A A

40–50

70 80 80 95

ANTARCTIC MARINE BENTHIC DIVERSITY

of the shallow-water Southern Ocean fauna described, as estimated by Winston (1992). Higher level taxonomy follows Barnes (1998). nd no data. Groups in bold italic are known to occur in the Southern Ocean, but for which no species lists appear to be available.

Comments Arntz et al. include hexactinellids. Demosponges well reviewed; calcareous forms in need of revision Benthic hydrozoan taxa only Dunn (1983) estimates 90 actinarians alone for the Southern Ocean One platyctenean ctenophore (Lyrocteis ﬂavopallidus) reported from shallow benthos of McMurdo Sound Free-living turbellarians known from intertidal sublittoral and sea-ice habitats

Larger forms well studied; many smaller species probably yet to be described (De Broyer, pers. comm.) Recently well revised Arntz et al. included some demersal and mid-water/pelagic taxa; total includes lithodid crab taxa reported recently from deeper water at South Georgia Known from Southern Ocean; no recent taxonomic summary known to authors Species list for marine mites only Recently reviewed thoroughly by Munilla León (2001a)

Winston reliability estimate for opisthobranchs only

A small but relatively well known group A small but relatively well known group Knox & Lowry (1977) comment that the total Southern Ocean polychaete fauna probably exceeds 800 species; a thorough review is badly needed Both oligochaetes and leeches known from Southern Ocean, but no species lists appear to be available Vent faunas likely to increase known species; many deep-sea forms probably await description Recent work has established new species; major recent review of cheilostomes; data for ctenostomes and cyclostomes probably underestimates Pterobranchs known from the Southern Ocean

Data for ascidians only Arntz et al. estimated 139 for all benthic ﬁshes; deﬁnition of benthic/demersal taxa difﬁcult

79

ANDREW CLARKE & NADINE M. JOHNSTON

Before undertaking any further analyses, however, we need to ask how reliable these estimates of diversity are.

How good are our estimates of Southern Ocean marine benthic diversity? For any meaningful interpretation of a biological diversity inventory, it is essential to understand the strengths and weaknesses of the information on which it is built. Thus only a knowledge of sampling coverage and intensity will allow a distinction to be made between a genuinely depauperate fauna (there are very few species at this location) and a poorly sampled fauna (there may be many species here, but we have only collected a small fraction of them). In this respect, Antarctica poses severe challenges for the benthic ecologist. Most of the Southern Ocean overlies the abyssal plain, the continental shelves are unusually deep, and access is impeded by ﬂoating ice shelves or vast areas of seasonal pack-ice. Despite these difﬁculties there is a long and proud history of benthic exploration in the Antarctic, and we now know a good deal about the fauna, at least for the major taxa of the continental shelves. The early work on Southern Ocean benthos took place largely during expeditions concerned primarily with geographic exploration or important physical observations such as geomagnetism or meteorology (Headland 1989, Fogg 1992, Clarke et al. 2000). Nevertheless these early collectors were remarkably thorough and Dayton (1990) laments that these important pioneers will never gain the recognition they deserve. The importance of these early contributions can be seen in the history of the published descriptions of major benthic taxa: the time-course of species description for many taxa show the importance of the systematic work based on the extensive collections made by early exploratory expeditions. In the past decade there has been a signiﬁcant increase in systematic work on Southern Ocean benthos, prompted largely by the SCAR programme on the ecology of the Antarctic sea ice zone (EASIZ). An important factor in assessing completeness of sampling of the fauna is the geographic spread of collections. The strengths and weaknesses of Southern Ocean benthic sampling to date are well shown by molluscs. Data for gastropod molluscs (Fig. 12) suggest that samples from the continental shelf have a good geographical spread, if somewhat thin away from the Ross Sea, Weddell Sea and Antarctic Peninsula. There are, however, relatively few samples from deeper water (Table 12) and the deep-sea fauna of Antarctica appears to be known only very poorly (Clarke in press). Extensive collections were made in many areas of Antarctica by biologists of the former Soviet Union, but relatively few data are currently available to western scientists. A broad scale analysis of the gastropod and bivalve data available for the Southern Ocean emphasises the patchy nature of the sampling. The number of samples taken in some areas is considerable (for example, Antarctic Peninsula, eastern Weddell Sea, Ross Sea and the islands of the Scotia arc): in other areas sampling is very thin (Table 12). To a large extent this pattern reﬂects ease of access; the Bellingshausen Sea is a very difﬁcult area for a ship to visit because of ice. There is also an effect of proximity to research stations undertaking marine biological work. Both of these effects are, of course, well known from studies of biological diversity elsewhere, both on land and in the sea. The number of samples taken in a given area is a key factor in the diversity recorded 80

ANTARCTIC MARINE BENTHIC DIVERSITY

Figure 12 Distribution of gastropod samples in the Southern Ocean. Isobaths are shown to mark the edge of the continental shelf (1000 m) and an arbitrary transition from continental slope to deep sea (3000 m). Also shown is the Polar Front, which is the northern boundary to the Southern Ocean. The database contains 7000 records, so not all data can be distinguished on the plot because many samples overlap. Data kindly provided by Katrin Linse and Alistair Crame, and the map plotted by Huw Grifﬁths.

(Table 12). This is an almost universal feature of studies of diversity in the marine environment where few areas have been studied in sufﬁcient detail for a high proportion of taxa to have been recorded. The exceptions are probably areas such as the northwest Atlantic, Mediterranean or Indo–West Paciﬁc with a long history of sampling, and taxa which are well known such as ﬁshes, echinoderms or molluscs. Even for a common and reasonably well known group such as bryozoans, patterns of recorded diversity in the north Atlantic are dominated by sampling effects (Clarke & Lidgard 2000). 81

82 148 203 157 19 98 147 2 132

209 351 89 9 191 408 2 231

South Georgia, South Sandwich Islands and South Orkney Islands Weddell Sea, east to 10ºW Dronning Maud Land, 10ºW to 65ºE Prydz Bay, 65ºE to 80ºE Wilkes Land, 80ºE to 170ºE Ross Sea, 170ºE to 150ºW Bellingshausen and Amundsen Seas, 80ºW to 150ºW Western Antarctic Peninsula, 80ºW to 50ºW, including South Shetland Islands but excluding Weddell Sea 42

44 36 14 0 10 40 0

Continental Shelf N SRgas SRbiv

Area

12

24 31 4 1 8 18 0

16

17 15 14 1 15 18 0

12

8 2 0 0 2 9 0

Continental Slope N SRgas SRbiv

5

16 2 1 0 1 5 6

3

9 10 1 0 2 4 2

4

4 8 0 0 0 3 4

Deep Sea N SRgas SRbiv

Table 12 Spatial variability in sampling intensity and observed species richness for Southern Ocean gastropods and bivalves. Data from Southern Ocean Molluscan Database, courtesy of Alistair Crame, Katrin Linse and Huw Grifﬁths (British Antarctic Survey), updating previous compilation by authors. Data broken down by depth strata and broad geographical areas. The depth strata were continental shelf (0–1000 m), continental slope (1000–3000 m) and deep sea (3000 m). Data shown are number of sites sampled (N), together with the number of species observed, in each depth/area bin: SRgas number of gastropod species in that bin, SRbiv number of bivalve species.

ANDREW CLARKE & NADINE M. JOHNSTON

ANTARCTIC MARINE BENTHIC DIVERSITY

Which taxa are important in the Southern Ocean? The Southern Ocean is geographically and environmentally extreme, in the sense that it has the highest southerly marine latitude and comprises water masses that are colder than anywhere else on earth. It is therefore of interest to ask whether the marine fauna shows any features which differ from those elsewhere. A simple preliminary approach would be to determine which taxa are most speciose, and which are least, and compare this ranking with other marine areas. Those taxa for which 100 species have been described for the Southern Ocean are listed in Table 13. Of these perhaps the most noteworthy are the pycnogonids and ascidians, the remainder being taxa which are frequently speciose in seas elsewhere. Where taxa are low in species richness this could simply be an evolutionary characteristic of the taxa overall; examples for the Southern Ocean would be sipunculans, echiurans and priapulans. In other cases, however, a low species richness in the Southern Ocean might not be typical of that group in warmer waters; examples would include gastropods, decapod crustaceans and teleost ﬁshes, all of which are highly speciose in many oceans but poorly represented in the Southern Ocean. A simple comparison of the species richness of major groups in the Southern Ocean with recent taxonomic summaries for Hawaii (Eldredge & Miller 1995) and northwestern European seas around the UK (Hayward & Ryland 1995, Howson & Picton 1997) highlights the importance of ascidians, echinoderms, polychaetes, pycnogonids, bryozoans, amphipods, isopods and hydroids, all of which have more species in the Southern Ocean than in either Hawaii or northwest Europe (Fig. 13). Of these, the most striking contrast by far is for pycnogonids where the Southern Ocean fauna is an order of magnitude more diverse than in either Hawaii or northwest Europe; pycnogonids are genuinely diverse in the Southern Ocean. Echinoderms are generally a well-described group and we must conclude that these are also particularly diverse in the Southern Ocean (although not all lineages are equally Table 13 Benthic invertebrate taxa showing the highest species richness in the Southern Ocean. Colloquial names are used as the groups represent different taxonomic levels (phylum, class or order). Taxa are ranked according to total species richness, with a threshold for inclusion in the table of 100. Also included are estimates of the percentage of the total world fauna for that taxonomic group, based on the range of data in Table 2.

Polychaetes Gastropods Amphipods Bryozoans Isopods True sponges Hydrozoans (benthic forms only) Pycnogonids Ophiuroids Ascidians Bivalves Asteroids (sea-stars) Holothurians (sea cucumbers)

Species richness

Percentage of world species

645 530 496 322 257 250 186 175 119 118 110 108 106

12.2 1 8.3 6.4–8.1 2.6 2.6–5.6 nd 17.5 6.0 5.9–9.4 5.5 7.2 9.2

83

ANDREW CLARKE & NADINE M. JOHNSTON

Figure 13 Comparison of species richness for selected benthic marine invertebrate groups in the Southern Ocean (this study), Hawaii (Eldredge & Miller 1995), and northwestern European waters around UK (Hayward & Ryland 1995, Howson & Picton 1997).

represented: Poulin & Féral 1996, Poulin et al. 2002). The data also suggest that bryozoans are notably diverse in the Southern Ocean. For other groups (polychaetes, sponges, hydroids, amphipods, isopods and ascidians) species richness in the Southern Ocean signiﬁcantly exceeds that in Hawaii but is broadly comparable with northwest Europe (Fig. 13). This may reﬂect any number of factors including isolation from continental shelves elsewhere, history of taxonomic work, and habitat (notably the balance between hard substrata and soft sediments). At present we can conclude that amphipods and isopods appear to be relatively well represented in the Southern Ocean, though not particularly diverse. Some amphipod and isopod lineages are, however, known to have radiated in the Southern Ocean (Watling & Thurston 1989, Brandt 1991a, 2000, Held 2000). The same may also be true of ascidians. The comparison in Figure 13 also emphasises the genuinely low diversity of bivalves and decapods in the Southern Ocean. Gastropods are comparable in richness with both Hawaii and UK, but a global comparison emphasises the tendency for gastropods to be lower in diversity at high latitudes compared with tropical faunas (Crame 1996, 2001). This contrast is driven by the enormous richness of the gastropod fauna of the Indo–West Paciﬁc region. To take a single example, intensive sampling of a single site in New Caledonia recovered 16 species of chiton, 16 scaphopods, 519 bivalves and 2187 gastropods (Bouchet et al. 2002). The data in Figure 13 also emphasise the diversity of anthozoans in tropical regions. Comparisons such as these are inevitably very crude, being beset with difﬁculties caused by differences in sampling intensity, taxonomic completeness, and area. Nevertheless when used with care, they can produce useful ﬁrst-order conclusions and help to formulate more focused biogeographic or evolutionary questions. An alternative approach to assessing which taxa are most important in the Southern 84

ANTARCTIC MARINE BENTHIC DIVERSITY

Ocean would be to calculate the fraction of the world’s continental shelf fauna found in Antarctica for each group. Assuming that Antarctic contains 11% of the world’s continental shelves (Table 3), then this might form a benchmark for a group of organisms whose representation in the Southern Ocean fauna is broadly proportionate to area. A large body of ecological work on terrestrial fauna, particularly of islands, has shown that the relationships between species richness and area is complex (and certainly not linear) (Rosenzweig 1995, Hubbell 2001). Nevertheless a benchmark of 11% might be taken as a ﬁrst-order indication of pro rata representation in the fauna. There are, however, numerous problems in the calculation of such data which must therefore be treated with great caution. The ﬁrst is the wide range of estimates for the world species richness of many taxa (Table 2), the second is that the data for world species richness mixes species from a range of habitats in addition to continental shelves, and a third is that global data are not always easily available for lower taxonomic levels such as classes or orders. Even allowing for these difﬁculties, none of the Southern Ocean taxa achieves levels of representation greater than about 15%, and most are well below this (Table 13). The data do, however, reinforce the conclusion that polychaetes, bryozoans, sponges, pycnogonids, amphipods and ascidians are well represented in the Southern Ocean benthic fauna, whereas gastropod molluscs are not. Despite the inevitable difﬁculties inherent in such crude calculations, these do show that generalisations about the Southern Ocean marine fauna (and perhaps even generalisations about the marine fauna in general) based on data for molluscs alone, or any other single taxon, are unlikely to be valid. This contrasts with the terrestrial environment where it has long been assumed that broad-scale patterns exhibited by plants, birds or butterﬂies reﬂect general patterns for other terrestrial taxa (MacArthur 1972, but see Platnick 1992). Both of the above comparisons are made problematic by difﬁculties of sampling intensity. One approach to correcting for sampling errors, which has been much used in diversity studies, has been to use rarefaction to estimate the number of species to be expected in a sample of a given number of individuals (Hurlbert 1971). Brey et al. (1994) used this technique for samples from Agassiz trawls in the Weddell Sea, and showed that data for bivalves, gastropods and isopods were distinctly higher than for northern polar waters (70°N), and in the upper range of tropical values. This is a somewhat counter-intuitive result for the molluscan taxa, as the Southern Ocean contains less than 1% of world gastropod species (Table 13). Although rarefaction brings its own suite of statistical problems (discussed by Gotelli & Graves 1996) the differences between the results from these two approaches is probably related to differences in spatial turnover in species identity (beta diversity) between polar and tropical habitats. Thus Hubold (1992) has shown that whereas a given sample of demersal ﬁshes from the Weddell Sea and the North Sea may contain similar number of species, the larger spatial variability in species identity in the North Sea leads to a larger overall species list (regional diversity). Abele (1974) documented a similar pattern for decapods, where within-habitat diversity was broadly similar across latitudes whereas regional diversity was much greater in the tropics because of larger differences in species composition between different habitats (turnover or beta diversity). Overall, current data suggest that pycnogonids are the one group which is especially diverse in the Southern Ocean, with bivalves, decapods and teleost ﬁshes notably underrepresented. The most diverse group overall is the polychaetes, and a number of groups are well represented, including ascidians, echinoderms, bryozoans, sponges, amphipods and isopods. 85

ANDREW CLARKE & NADINE M. JOHNSTON

Endemism Many authors have commented on the high degree of endemism in the Southern Ocean fauna. The most recent compilation is that of Arntz et al. (1997) who showed a range of values from about 35% (scleractinian corals) to about 90% (pycnogonids). The recent compilation of amphipod species richness by De Broyer & Jazdzewski (1996) indicated a level of endemism of 80% at the species level, and 17% at the genus level for their Antarctic region. These data indicate the length of time for which the Southern Ocean fauna has been isolated (Clarke & Crame 1997). The degree of endemism is also scale-dependent in that it is a strong function of the area over which it is calculated: endemism is 100% at the level of the whole globe, and 0% at the level of a small bay. The high level of endemism shown by the Antarctic benthic fauna is thus partly a function of the size of the Southern Ocean. Endemism remains high, however, on smaller spatial scales: De Broyer & Jazdzewski (1996) have shown that endemism at the species level in gammaridean amphipods is 38% in East Antarctica and 54% in West Antarctica. Overall we must conclude that a high level of endemism is a real feature of the Southern Ocean benthic fauna.

Spatial variation in species richness Although the overall levels of species richness in the Southern Ocean are moderately high in some taxa, the Southern Ocean is a large place. Many Southern Ocean taxa appear to have a circum-Antarctic distribution, but not all species will be found in all places, even where apparently suitable habitat exists. There are still relatively few places in the Southern Ocean for which comprehensive faunal lists exist. Two relatively well studied areas are the Weddell Sea and Admiralty Bay. Although a considerable amount of biological work has been undertaken in the Ross Sea sector, until recently relatively little of this work was taxonomic. A comparison, albeit crude, can, however, be made for the whole Southern Ocean with the fauna known for the Weddell Sea, the Ross Sea, and very preliminary data for the much smaller areas of Admiralty Bay, Arthur Harbour and Signy Island (Table 14). These data show that the general patterns of dominance by certain taxa are reﬂected at all scales. Species/area relationships have been described for many areas of the globe, over a variety of spatial scales, and for many taxa (for a thorough review of this topic see Rosenzweig 1995). Few such relationships have been established for marine systems. The data collated in Table 14 indicate that there is a positive species/area relationship within the Southern Ocean, but with the spatial extent of sampling being known only poorly and with data not available for some taxa, we cannot even start to derive any meaningful quantitative relationships. The picture will also be confounded by latitudinal or other spatial variations in benthic species richness.

Latitudinal variations in Southern Ocean marine diversity Although the Southern Ocean extends over a wide latitudinal band, from 54°S at South Georgia to 78°S at McMurdo Sound in the Ross Sea, there have been remarkably few 86

ANTARCTIC MARINE BENTHIC DIVERSITY

Table 14 A comparison of species richness in two regions and three smaller areas of the Southern Ocean with the total regional fauna. Data from Richardson (1976), Arnaud et al. (1986), De Broyer & Jazdzewski (1993), Siciñski & Janowska (1993), Jazdzewski et al. (1986), Arntz et al. (1997), Arnaud et al. (1998, 2001), Gutt et al. (2000) and unpublished British Antarctic Survey records. Further sampling will undoubtedly increase many of these richness ﬁgures. nd no data. Southern Ocean Porifera Symplasma Cnidaria Hydrozoa Anthozoa Brachiopoda Bryozoa Cheilostomatida Mollusca Gastropoda Bivalvia Echinodermata Asteroidea Crinoidea Echinoidea Ophiuroidea Holothuroidea Crustacea Decapoda Amphipoda Isopoda Pycnogonida Annelida Polychaeta Urochordata

Regions Weddell Ross Sea Sea

Admiralty Bay

Areas Signy Island

Arthur Harbour

279

nd

57

nd

nd

186 86 19

36 33 nd

25 19 7

nd nd nd

nd nd nd

nd nd nd nd nd

249

180

163

nd

87

nd

530 110

145 43

nd nd

35 27

82 nd

8 20

108 28 49 119 106

50 6 nd 43 35

28 9 nd nd nd

15 1 4 15 3

nd nd nd nd nd

nd nd 2 6 2

8 496 257 175

4 174 68 69

nd nd 33 64

2 99 nd nd

nd nd nd nd

nd 48 38 6

645 118

225 24

97 40

⬃100 nd

78 nd

142 4

studies of latitudinal variation in marine species richness within Antarctica. It is generally accepted that, as on land, benthic species richness will be greatest around the sub-Antarctic islands such as South Georgia, and decline southwards through the maritime Antarctic to the high Antarctic regions of the Ross Sea and Weddell Sea. However, there have been almost no tests of this. The clearest latitudinal cline reported so far in the marine realm of the Southern Ocean is for macroalgae along the Antarctic Peninsula (Moe & DeLaca 1976). Although the sampling was not comprehensive being limited to shallow depths, it was broadly comparable between sites and revealed a strong decrease in macroalgal species richness from north to south along the western Antarctic Peninsula (Fig. 14). Although this cline is likely to result in a similar cline of those epifauna reliant on macroalgae as a substratum, it is important to establish whether a latitudinal cline is a general feature of benthic diversity in the Southern Ocean. Latitudinal diversity clines elsewhere in the marine system tend to be strongest where there are strong meridional clines in environmental and oceanographic variables (for 87

ANDREW CLARKE & NADINE M. JOHNSTON

Figure 14 Latitudinal cline in macroalgal diversity along the Antarctic Peninsula. Data from Moe & DeLaca (1976).

example, along the Atlantic and Paciﬁc coasts of North America), or where the imprint of glacial history is strong. Antarctica has relatively few places where the continental shelf has a strongly meridional orientation and the predominant features of the oceanography are circumpolar. This would suggest that latitudinal patterns will not be a major feature of the biogeography of Southern Ocean benthos, with the possible exception of the Antarctic Peninsula. Apart from the sub-Antarctic islands such as South Georgia, benthic water temperatures are relatively homogeneous across the continental shelves of Antarctica. Winter minima are usually around 1.9ºC, and annual variability typically less than 2°C; this annual range is small compared with annual ranges greater than 10 to 15°C typical of many lower latitude seas. The major differences within the Southern Ocean are in the timing and duration of the summer warming, and the peak temperatures reached (all of which are subject to signiﬁcant variability between years). Coupled with variability in ice dynamics and latitudinal variations in photoperiod, these lead to strong spatial variation in the timing of the summer water column phytoplankton bloom. These factors are likely to inﬂuence much of the ecology of the benthos and together with the relatively small annual variation in temperature make the Antarctic benthic environment an ideal natural laboratory for distinguishing the ecological effects of temperature and food (Clarke 1991, 1998). It is not at all clear what effects, if any, these variations will have on the overall diversity of the fauna.

Biogeographic patterns in Southern Ocean marine benthos There has long been an interest in the biogeography of Antarctica (Darlington 1965). In the terrestrial realm most attention has been directed at historical inﬂuences such as the break up of Gondwana and glaciation. In the marine realm the two major concerns have traditionally 88

ANTARCTIC MARINE BENTHIC DIVERSITY

Figure 15 The basic biogeographical division of the Southern Ocean benthos as proposed by Hedgpeth (1970) and Dell (1972). This classiﬁcation has stood the test of time, remaining valid in the context of considerable new sampling. The precise location of the boundaries between the Antarctic Peninsulas and continental Antarctic regions are, however, not well deﬁned as they fall in areas where sampling intensity is especially low.

been the routes of faunal exchange, and the inﬂuence of the predominantly circumpolar oceanography. By the early 1970s, knowledge of the Southern Ocean benthic fauna was sufﬁcient for major biogeographic and taxonomic reviews to be written (Hedgpeth 1969, 1970, 1971, Dell 1972). These resulted in a biogeographic scheme for Southern Ocean benthos that has remained the working paradigm ever since (Fig. 15). More recent analyses for amphipods (De Broyer & Jazdzewski 1996) and bryozoans (Barnes & De Grave 2000) have largely conﬁrmed the previously established patterns, though adding the extra dimension of differences between the faunas of East and West Antarctica. The basic biogeographic subdivisions of the Southern Ocean benthic fauna are thus South Georgia, the Antarctic Peninsula (including the South Orkney Islands) and high Antarctica, comprising the fauna of the continental shelf at highest latitudes although with signiﬁcant differences between East and 89

ANDREW CLARKE & NADINE M. JOHNSTON

West Antarctica. The intermediate latitudes of the Antarctic Peninsula and the South Orkney Islands are sometimes referred to as the maritime Antarctic. These patterns are usually interpreted in terms of history (Clarke & Crame 1997, in press, Crame & Clarke 1997, Crame & Rosen 2002). The fauna of the Antarctic Peninsula shows many features of exchange with South America along the Scotia arc, and the differences between East and West Antarctica will reﬂect the different evolutionary histories of these areas during the break up of Gondwana (especially in relation to the inﬂuence of Tethys). The continental shelves around Antarctica have been both more and less extensive in the past as the size and extent of the continental ice sheet has waxed and waned, and at times the Ross and Weddell Sea embayments would have represented extensive shallow seas (Clarke & Crame 1989). Brey et al. (1996) have shown that many Antarctic benthic taxa have a wider bathymetric range than relatives elsewhere, and this would appear to reﬂect previous extensions of the continental ice sheet reducing the area of available shelf habitat and forcing many benthic taxa into deeper water refugia. It is possible that frequent ﬂuctuations in the extent to which the continental ice sheet covered and exposed the continental shelf would have driven speciation in the Southern Ocean benthos (Clarke & Crame 1997, in press). Such variations in habitat, together with the effects of the long-term cooling of sea water, may have been a key factor in the very low species richness of some taxa (such as bivalves or decapod crustaceans), or in the complete absence of others (for example, brachyuran crabs and many families of teleost ﬁshes). The difﬁculty faced by ecologists concerned with the evolutionary history of the Antarctic marine fauna is moving from post hoc explanations to testable ideas. The fossil record would be the obvious source of data, but unfortunately the fossil record of Antarctic marine invertebrates is patchy with gaps in many key periods (Crame 1992, Clarke & Crame 1989).

Some preliminary macroecological analyses The main aim of this review has been to derive the best inventory for Southern Ocean benthic diversity from the available data. Although these data suffer from severe sampling problems, they do permit a number of preliminary analyses which throw light on current problems of macroecological interest.

The latitudinal cline in marine diversity The latitudinal cline in diversity is a distinctive, well described feature of both ﬂora and fauna on land. Although it had long been assumed that a similar cline would be found in the sea, evidence for this has proved to be more equivocal (Clarke 1992, Clarke & Crame 1997). A clear cline has been described for continental shelf molluscs on the Atlantic and Paciﬁc coasts of North America (Roy et al. 1994, 1998), bryozoans in the North Atlantic (Clarke & Lidgard 2000) and some taxa in the deep sea (Rex et al. 1993, but see Gray 1994). In other areas, notably in the southern hemisphere, there is little evidence for a strong latitudinal cline in diversity (Poore & Wilson 1993, Brey et al. 1994, Clarke & Crame 1997). The Southern Ocean fauna is critical to the existence of any cline in marine diversity in the southern hemisphere. Given the high diversity of many tropical marine habitats, a low 90

ANTARCTIC MARINE BENTHIC DIVERSITY

Southern Ocean diversity would essentially create a latitudinal gradient by itself, whereas high polar diversity would signiﬁcantly decrease the intensity of any such gradient. The problem in determining the existence of a diversity cline in southern hemisphere shallowwater marine benthos is the current disposition of continental land masses. In strong contrast to the northern hemisphere there are relatively few continental shelves with extensive meridional (north–south) alignment. Comparison of marine diversity between Antarctica and either southern Africa or Australasia would be interesting, but the only sites where the existence of latitudinal gradients in shallow water benthos is likely to be demonstrated unequivocally are the Paciﬁc and Atlantic coasts of South America, the Scotia arc and the Antarctic Peninsula. The low diversity of gastropod and bivalve molluscs in the Southern Ocean is striking in comparison with the richness of both groups in some tropical regions. This suggests that in these two groups, at least, diversity broadly declines from a tropical high towards both poles (Crame 2000a,b, 2001). These diversity clines are steepest for the youngest clades, suggesting a pattern dominated by tropical diversiﬁcation and subsequent migration polewards (Crame 2001). It is, however, important to recognise that some molluscan taxa have undoubtedly radiated at high southern latitudes. An example here would be the predatory buccinid and muricid gastropods (Crame 1996), and molecular phylogenetic work will undoubtedly uncover other high latitude molluscan radiations. Two other groups with notably low diversity in the Southern Ocean are decapod crustaceans and teleost ﬁshes. The reasons for this are unclear, as both groups are well represented in the fossil record and the demise of the decapod fauna appears to have been relatively recent (Feldmann & Tshudy 1989). Among ﬁshes the notothenioid radiation is now fairly well understood, and it has been argued that this shows many of the features of a species ﬂock analogous to the cichlid ﬁshes of African rift lakes or the cottids of Lake Baikal (Eastman & Clarke 1998). In the case of both ﬁshes and decapod crustaceans the earlier Southern Ocean fauna appears to have been almost eradicated by a historical event. It is not clear what this event might have been, although the best estimate for the date of the basal radiation of the notothenioids suggests that a critical event was the initial cooling of Antarctica and the onset of glaciation (Clarke & Johnston 1996). The reason for the extinction of many decapods is still obscure. In contrast to the decapods and ﬁshes, some groups appear to be strikingly diverse in the Southern Ocean. These include sponges, ascidians, amphipods, isopods and pycnogonids (Fig. 13). These data are not easy to interpret in terms of latitudinal clines, however, because there are few comparative data available. One exception to this are bryozoans, for which there has been recent extensive work in South America (Moyano 1984, López Gappa 2000). These studies have shown that whereas there is a cline in diversity along the South American coasts, the Southern Ocean fauna is both rich and quite distinct. Moreover, for cheilostomes there is a strong relationship between the age of a bryozoan family and the likelihood of having representatives in the Southern Ocean (Barnes & De Grave 2000). These examples show clearly that the diversity and composition of the Southern Ocean benthic marine fauna can be understood only in terms of historical processes (Lipps & Hickman 1982). The origins of the present fauna are with the coastal fauna of Gondwana prior to break up. The fragmentation of Gondwana inﬂuenced the fauna particularly by the introduction of a Tethyan element through the Weddell province (Clarke & Crame 1989). Subsequent development of the fauna was inﬂuenced particularly by exchange (in both directions) along the Scotia arc and with the deep sea, but the dominant feature was 91

ANDREW CLARKE & NADINE M. JOHNSTON

evolution in situ. The relative representation of the different major taxa would appear to reﬂect accidents of history rather than the nature of the Antarctic marine environment. Climatic change and glaciation resulted in the extinction of some groups, and provided evolutionary opportunities for others. This is not to say that ecological factors have had no role to play; there is, for example, strong evidence that reproductive ecology has had an important role to play in determining which echinoderm taxa are present in the modern Antarctic benthic fauna (Poulin & Féral 1996, Poulin et al. 2002). Overall, however, we should look to historical processes as the ﬁrst order explanation of why the fauna has the diversity and composition it does.

Distribution of species amongst genera It has been known since the pioneering work of Willis & Yule (1922) that the distribution of species amongst genera tends to follow a characteristic shape. Monotypic genera are typically the most frequent, genera with two species somewhat less common, and so on through a tail of fewer genera containing increasingly large numbers of species. Willis & Yule (1922) named this the hollow curve distribution, and noted that the relationship was often linear after logarithmic transformation of both variables. This frequency distribution appeared to apply both within taxonomic groups (for example many plant families) and also for local ﬂoras (where all taxa are pooled). Willis (1922) provided a conceptual basis for these observations through his age and area hypothesis, with younger taxa being less speciose and also less widely distributed geographically than older ones. This simple and intuitively appealing explanation has not stood the test of time (see, for example, Cronk 1989), but there remains a feeling that this more or less universal distribution must be telling us something about evolutionary processes. More recent analyses have been built around a series of theoretical frequency distributions or evolutionary null models. Dial & Marzluff (1989) explored a series of null models of evolutionary diversiﬁcation and Minelli et al. (1991) examined the fractal nature of diversity within the taxonomic hierarchy. Such analyses depend on particular models of evolutionary diversiﬁcation, and also assume that taxonomy reﬂects evolutionary history in some meaningful way. Walters (1986) takes the extreme view that the patterns are entirely artefactual being driven exclusively by taxonomic biases. The long period over which the Southern Ocean benthic marine fauna has evolved in isolation (Clarke & Crame 1989) suggests that the distribution of species per genus would be worth examining in that it might reveal the patterns to be expected for a radiation in the relative absence of dilution by immigrant taxa. Many of those taxa which are well represented in the Southern Ocean exhibit hollow curve distributions (Figs 2–11). Few of these are, however, convincingly linear on a double logarithmic plot (Fig. 16), suggesting that null models predicting a power law relationship do not reﬂect the underlying evolutionary processes accurately.

Size and abundance in Southern Ocean benthos In general, studies of Antarctic benthos have relied on remote sampling techniques that are essentially destructive (for example, grabs, dredges and bottom trawls). Trawls and dredges 92

ANTARCTIC MARINE BENTHIC DIVERSITY

(a)

(b)

Figure 16 The distribution of species amongst genera for Southern Ocean benthos. Note logarithmic axes. For clarity of presentation, taxa have been divided into those with high species richness in the Southern Ocean (a, left) and low species richness there (b, right).

in particular suffer from the disadvantages of damaging many specimens, destroying any structure in the community (such as tiering or ecological associations) and mixing samples from many different habitats, thereby masking any heterogeneity in the distribution of different assemblages. In shallow waters these difﬁculties can be overcome by the use of SCUBA techniques, but safety considerations restrict both the depth and the amount of time available underwater and thereby limit the types of study that can be undertaken. Remote photographic techniques were used in deeper waters for some of the earliest studies of Antarctic benthos (Bullivant & Dearborn 1967) but were then largely ignored. Recently, however, German biologists have made extensive use of modern photographic techniques in the Weddell Sea, and these new studies have demonstrated clearly the value of such techniques to marine ecology (Barthel et al. 1991, Ekau & Gutt 1991, Gutt et al. 1991, 1994, Gutt & Piepenburg 1991, Barthel & Gutt 1992, Gutt & Schickan 1998, Gutt & Starmans 1998). One of the primary considerations for any ecologist faced with making the choice of sampling gear is that of size selectivity. Photographic techniques are by their very nature limited almost exclusively to epifaunal macroinvertebrates above a certain size. Sampling techniques based on nets will also miss all organisms below a size set by the mesh of net being used, and the mesh size of sieves used for sorting will also inﬂuence the size distribution or organisms ﬁnally collected. These considerations are of particular relevance to Antarctic benthic communities, for although the small number of Antarctic taxa which are unusually large have long attracted attention (Arnaud 1974) it is now recognised that many Southern Ocean benthic organisms are very small. Brey & Clarke (1993) collated all the data on the population dynamics of Antarctic marine benthos available up to 1992. Comparison of the polar species with species from temperate and tropical waters indicated no signiﬁcant difference in mean adult size. This 93

ANDREW CLARKE & NADINE M. JOHNSTON

result, however, reﬂects the tendency of ecologists to choose larger organisms for study rather than the actual size distribution of polar marine invertebrates. Antarctic bivalves and gastropods, for example, have long been known to be typically small (Nicol 1964, 1966a,b, 1978, Arnaud 1974). In the case of gastropods many Antarctic species are so small as to be missed by some conventional types of sampling or sorting gear. The most thorough investigation of size in Antarctic benthos has been for amphipods. Chapelle & Peck (1999) examined size spectra for amphipod assemblages from a range of marine regions and showed clearly that while minimum size was effectively invariant, maximum size was inversely related to water temperature. The explanation for the existence of larger taxa in colder water was a combination of absolute oxygen content and the architecture of the vascular system. This explanation was conﬁrmed by examining amphipod assemblages from freshwater lakes (where oxygen solubility is different from sea water) and at high altitude (Chapelle & Peck 1999). Whereas maximum size in amphipods appears to be determined by a combination of an environmental factor and internal bauplan, the constraint on minimum size is less clear. It may be related to constraints on egg size but this is by no means certain (Chapelle & Peck 1999). The minimum size for Antarctic amphipods, however, does suggest that there may be many undescribed small taxa waiting to be discovered (De Broyer & Jazdzewski 1996). The importance of bauplan (speciﬁcally vascular system architecture) in determining maximum body size in amphipods suggests that the environmental factors setting size constraints will vary between groups of organisms. It has been suggested that an important factor in determining the size and shell morphology of molluscs is the cost of calciﬁcation (Graus 1974), and this might explain the generally small size of polar molluscs (Nicol 1964, Clarke 1983). Some Antarctic molluscs do achieve a large size, however, and although in some cases this is achieved through a poorly calciﬁed or very thin shell, this is not universal. It is quite possible that the absence of crushing predators (notably crabs, lobsters and ﬁshes specialising in molluscs) may also be an important factor. If water temperature does play a role in determining body size and/or skeletal construction through its effect on calciﬁcation costs, then this inﬂuence should be apparent in echinoderms, brachiopods and solitary corals. To date these groups have not been examined in this context. A second feature of importance in describing patterns in animal communities is the distribution of individuals amongst species. Typically species/abundance plots will exhibit a lognormal distribution, and the conventional graphical representation is that proposed by Preston (1962). Construction of Preston plots requires careful attention to sampling techniques: the sampling must be quantitative to ensure all individuals are taken, cover the entire size range, and be sufﬁciently thorough to ensure that most of the rare species are sampled. Two Antarctic samples where these criteria have been fulﬁlled, at least approximately, are shown in Figure 17. These two plots derive from very different types of sample. The gastropod data were obtained by pooling a series of repeated monthly samples obtained by suction sampling of standard sized quadrants (0.25 m2) in an area of shallow water (2–12 m-depth) with a mixed substratum (Picken 1980). These data therefore represent a fully quantitative sample for all animals above about 2 mm in size. The amphipod data are from a semi-quantitative study which covered a range of different habitats using a variety of sampling techniques. The full collection was then analysed by Thurston (1972). The Preston plot for the gastropod data (Fig. 17a) shows a roughly normal shape, whereas that for amphipods (Fig. 17b) suggests the inﬂuence of sampling error (no data in octaves 3 to 5, and three species represented by two or fewer individuals). 94

ANTARCTIC MARINE BENTHIC DIVERSITY

(a)

(b)

Figure 17 Preston plots for two groups of Southern Ocean benthos. Note that the abscissa is in the conventional units of log2(n) where n is the number of individuals per species. (a) Gastropod molluscs from Signy Island (Picken, 1980). (b) Amphipod crustaceans from Signy Island (Thurston 1972).

Ecologists have long debated the signiﬁcance of the shape of Preston species/abundance plots. A typical plot of logarithmically transformed data where a large number of individuals have been sampled often looks as though it would be well described by a normal distribution truncated to the left. This truncation (the so-called veil line) is inevitable because those species which are so rare that the expected number of individuals in the sample being plotted is less than one, would be unlikely to be recorded (Pielou 1969). The position of the veil line would be expected to move to the left as sample size increases (Magurran 1988). The underlying process (or processes) leading to the frequently (but by no means universally) observed log-normal distribution of species’ abundances has remained elusive. Although some theoretical models have proved successful under some circumstances, no single model has been found to provide a general description of species/abundance data. In fact the log-normal distribution would be expected as a result of a large number of interacting inﬂuences, and hence its general applicability as a descriptor may simply be telling us that the abundance of species in the wild is dictated by many factors and not a single process. Figure 17 shows the only species/abundance data for Antarctic marine benthos known to the authors where a quantitative collection has been accompanied by a thorough taxonomic study. These data indicate that although many species are small, species/abundance distributions are not discernibly different from those found in marine habitats elsewhere (Magurran 1988, Gaston 1994, Gray 1997).

95

ANDREW CLARKE & NADINE M. JOHNSTON

Concluding remarks The benthic fauna of the Southern Ocean continental shelves is now reasonably well known. Although detailed knowledge probably lags behind that of the fauna of the North Atlantic, the Caribbean or parts of the Indo–West Paciﬁc, it is by no means as poor as might be envisaged from the isolation of the Antarctic continent. The long period of evolutionary history in situ around Gondwana has resulted in a generally diverse fauna, though some taxa are absent or poorly represented. Notable amongst the latter are decapod crustaceans and teleost ﬁshes apart from the notothenioids. Neither bivalve nor gastropod molluscs are diverse in the Southern Ocean when compared with many tropical regions, whereas, in contrast, amphipod and isopod crustaceans, pycnogonids and many suspension feeding taxa, especially bryozoans, sponges and ascidians, are well represented. In all cases it is only some lineages which have diversiﬁed, indicating that evolutionary questions concerning the origin, diversiﬁcation or extinction of the Southern Ocean marine fauna will have no single answer. The evolutionary history of each group appears to reﬂect a different response to the tectonic, climatic and oceanographic changes to which they have been subject through history. Endemism is high, mainly because of the large area of the Antarctic continental shelf, and the oceanographic isolation of the fauna. Although most of the Southern Ocean is deep water, the fauna of the continental slope and the abyssal plain of Antarctica are known only poorly. It is only in recent years that the spectacular radiation of liparid ﬁshes on the continental slope of Antarctica has become recognised, largely through the work of biologists of the former Soviet Union. There is much we have still to learn of the marine fauna of this unique and important area.

Acknowledgements We wish to acknowledge the important contribution to this work by Jenny Twelves, who undertook an initial compilation of species richness for many taxa, and Rebecca Mitchell who prepared the original versions of many of the ﬁgures. We would also particularly like to thank Wolf Arntz for providing us with details of his own compilations used in preparing his 1997 review written with colleagues. Angelika Brandt, Peter Hayward, David Barnes, Katrin Linse, Alistair Crame and Claude De Broyer have also provided extensive advice relating to particular taxonomic groups. Alex Rogers and Lloyd Peck have been a valuable source of ideas and criticism. In preparing the GIS maps we have received considerable help from Sharon Grant, Paul Cooper and Steve Evans who undertook the enormous task of editing the GEBCO bathymetry for the Southern Ocean into a GIS-compatible format. Huw Grifﬁths and Katrin Linse prepared the map of gastropod distribution and provided the data for Table 13 from their extensive database. Christine Phillips provided invaluable help in checking citations and locating some of the older or more obscure literature. This work was funded by the United Kingdom Foreign and Commonwealth Ofﬁce, and was undertaken at the British Antarctic Survey. It forms a contribution to the SCAR programme Ecology of the Antarctic Sea-Ice Zone (EASIZ).

96

ANTARCTIC MARINE BENTHIC DIVERSITY

References Abele, L. G. 1974. Species diversity of decapod crustaceans in marine habitats. Ecology 55, 156–161. ADD Consortium 2000. Antarctic digital database, version 3.0. Database manual and bibliography. Cambridge: Scientiﬁc Committee on Antarctic Research. Agatep, C. P. 1967. Some elasipodid holothurians of Antarctic and Subantarctic seas. Antarctic Research Series 11, 49–71. Ahn, I.-Y. & Kang, Y.-C. 1991. Preliminary study on the macrobenthic community of Maxwell Bay, South Shetland Islands, Antarctica. Korean Journal of Polar Research 2, 61–71. Alarcon-Castillo, J. C. 1967. Ophiuroides colectados por la XIX Expedition Antarctica Chileno. Publicación Instituto Antárctico Chileno 13, 1–35. Allcock, A. L. 1997. The genetics and taxonomy of Southern Ocean Octopodidae, with special reference to the genus Pareledone. PhD thesis, University of Liverpool. Allcock, A. L., Piatkowski, U., Rodhouse, P. G. K. & Thorpe, J. P. 2001. A study on octopodids from the eastern Weddell Sea, Antarctica. Polar Biology 24, 832–838. Androsova, E. I. 1968. Antarctic and subantarctic bryozoans of Cyclostomata and Ctenostomata orders. Issledovaniya Fauny Morei 6, 35–84. Androsova, E. I. 1972. Marine invertebrates collected from Adelie Coast by the 12th and 15th French Antarctic Expeditions. 6. Bryozoa. Expéditions Polaires Françaises. Missions Paul-Emile Victor 333, 87–102. Arnaud, P. M. 1964. Echinodermes littoraux de Terre Adélie (Holothuries exceptées) et Pélécypodes commensaux d’Echinoides Antarctiques. Expéditions Polaires Françaises. Missions Paul-Emile Victor 258, 1–69. Arnaud, P. M. 1972a. Marine invertebrates collected from Adelie Coast by the 12th and 15th French Antarctic Expeditions. 9. Pycnogonida. Expéditions Polaires Françaises. Missions Paul-Emile Victor 333, 135–155. Arnaud, P. M. 1972b. Invertébrés marins des XIIéme et XVéme expéditions antarctiques françaises en Terre Adélie. 8. Gastéropodes Prosobranches. Téthys Supplement 4, 105–134. Arnaud, P. M. 1974. Contribution à la bionomie marine benthique des régions antarcticques et subantarctiques. Téthys 6, 465–656. Arnaud, P. M., Jazdzewski, K., Presler, P. & Sicinski, J. 1986. Preliminary survey of benthic invertebrates collected by Polish Antarctic expeditions in Admiralty Bay (King George Island, South Shetland Islands, Antarctica). Polish Polar Research 7, 7–24. Arnaud, P. M., López, C. M., Olaso, I., Ramil, F., Ramos-Esplá, A. A. & Ramos, A. 1998. Semiquantitative study of macrobenthic fauna in the region of the South Shetland Islands and the Antarctic Peninsula. Polar Biology 19, 160–166. Arnaud, P. M., Troncoso, J. S. & Ramos, A. 2001. Species diversity and assemblages of macrobenthic Mollusca from the South Shetland Islands and Bransﬁeld Strait (Antarctica). Polar Biology 24, 105–112. Arnaud, P. M. & van Mol, J. J. 1979. Anatomy, ecology and distribution of the Volutidae and Volumitridae (Gastropoda: Prosobranchia) of the southern Indian Ocean. Veliger 22, 19–31. Arntz, W. E., Brey, T. & Gallardo, V. A. 1994. Antarctic zoobenthos. Advances in Marine Biology 32, 241–304. Arntz, W. E., Gutt, J. & Klages, M. 1997. Antarctic marine biodiversity: an overview. In Antarctic communities: species, structure and survival, B. Battaglia et al. (eds). Cambridge: Cambridge University Press, 3–14. Averintsev, V. G. 1972. Benthic polychaetes errantia from the Antarctic and sub-Antarctic collected by the Soviet Antarctic Expeditions. Issledovaniya Fauny Morei 19, 88–293. (In Russian.) Bamber, R. N. 1995. A new species of Pycnogonum Brunnich, 1764 (Arthropoda, Pycnogonida) with other pycnogonid records from the Falkland and South Shetland Islands. Ophelia 40, 199–205.

97

ANDREW CLARKE & NADINE M. JOHNSTON

Barnard, J. L. 1969. The families and genera of marine gammaridean amphipods. Bulletin of the United States National Museum 271, 1–535. Barnard, J. L. & Karaman, G. S. 1991. The families and genera of marine gammaridean Amphipoda (except marine gammaroids). Parts 1 and 2. Records of the Australian Museum (Suppl. 13), 1–866. Barnes, D. K. A. & De Grave, S. 2000. Biogeography of Southern Ocean bryozoans. Vie et Milieu 50, 261–273. Barnes, R. S. K. (ed.) 1998. The diversity of living organisms. Oxford: Blackwell Science. Barthel, D. 1997. Fish eggs and pentacrinoids in Weddell Sea hexactinellids: further examples for the structuring role of sponges in Antarctic benthic ecosystems. Polar Biology 17, 91–94. Barthel, D. & Gutt, J. 1992. Sponge associations in the eastern Weddell Sea. Antarctic Science 4, 137–150. Barthel, D., Gutt, J. & Tendal, O. S. 1991. New information on the biology of Antarctic deep-water sponges derived from underwater photography. Marine Ecology Progress Series 69, 303–307. Barthel, D. & Tendal, O. 1994. Antarctic Hexactinellida. Koenigstein: Koeltz Scientiﬁc. Battershill, C. N. 1989. Distribution and abundance of benthic marine species at Cape Armitage, Ross Island, Antarctica: initial results. New Zealand Antarctic Record 9, 35–52. Bayer, F. 1980. Armadillogorgia cythella and Ophidiogorgia paradoxa, two new genera and species of primnoid octocorals (Coelenterata: Anthozoa) from South Georgia and South Orkney Islands. Proceedings of the Biological Society of Washington 93, 216–228. Bayer, F. M. 1996a. The Antarctic genus Callozotron and its relationship to Primnoella (Octocorallia, Gorgonacea, Primnoidae). Proceedings of the Biological Society of Washington 109, 150–203. Bayer, F. M. 1996b. New primnoid gorgonians (Coelenterata, Octocorallia) from Antarctic waters. Bulletin of Marine Science 58, 511–530. Bayer, F. M. & Stefani, J. 1987. New and previously known taxa of isidid octocorals (Coelenterata: Gorgonacea), partly from Antarctic waters. Proceedings of the Biological Society of Washington 100, 937–991. Benham, W. B. 1921. Polychaeta. Scientiﬁc Reports of the Australasian Antarctic Expedition 1911–1914. Series C 6–7, 5–124. Benham, W. B. 1927. Polychaeta. British Antarctic Terra Nova Expedition 1910. Zoology 7, 47–182. Bergquist, P. R. 1978. Sponges. Berkeley: California University Press. Bernasconi, I. & D’Agostino, M. M. 1973a. El genero Ophiomastus Lyman, 1878 y descricion e una especie abisal, O. molinae Castillo, 1968. Physis, Seccion A 32, 197–202. Bernasconi, I. & D’Agostino, M. M. 1973b. Nueva specie de oﬁuroideo Antarctico, Ophiurolepis granulifera. Physis, Seccion A 32, 263–266. Blanco, O. M. 1977. New Antarctic hydrozoans. Contribuciónes del Instituto Antártico Argentino 205, 1–15. Blanco, O. M. 1984. Data on Antarctic and Subantarctic Hydrozoa. Contribuciónes del Instituto Antártico Argentino 294, 1–53. Blanco, O. M. & Belluscí de Miralles, D. A. 1972. Hydrozoa of Peter I Island. Contribuciónes del Instituto Antártico Argentino 145, 1–43. Blazewicz, M. & Jazdzewski, K. 1996. A contribution to the knowledge of Tanaidacea (Crustacea, Malacostraca) of Admiralty Bay, King George Island, Antarctic. Polish Polar Research 17, 213–220. Borradaile, L. A. 1916. Crustacea. Part 3. Cirripedia. British Antarctic Terra Nova Expedition, 1910. Natural History Report. Zoology 3, 127–136. Boucher, G. & Lambshead, P. J. D. 1995. Ecological diversity of marine nematodes in samples from temperate, tropical, and deep sea regions. Conservation Biology 9, 1594–1604. Bouchet, P., Lozouet, P., Maestrati, P. & Heros, V. 2002. Assessing the magnitude of species richness in tropical marine environments: exceptionally high numbers of molluscs at a New Caledonia site. Biological Journal of the Linnaean Society 75, 421–436. Brandt, A. 1988. Antarctic Serolidae and Cirolanidae (Crustacea: Isopoda). Koenigstein: Koeltz Scientiﬁc Books.

98

ANTARCTIC MARINE BENTHIC DIVERSITY

Brandt, A. 1991a. Colonization of the Antarctic shelf by the Isopoda (Crustacea, Malacostraca). Berichte zur Polarforschung 91, 1–240. Brandt, A. 1991b. Redescription of the deep-sea asellote Isopod Mesosignum antarcticum Schultz, 1979 (Isopoda, Asellota, Mesosignidae). Zoologischer Anzeiger 227, 229–239. Brandt, A. 1992a. New Asellota from the Antarctic deep-sea (Crustacea, Isopoda, Asellota), with descriptions of 2 new genera. Zoologica Scripta 21, 57–78. Brandt, A. 1992b. New deep-sea Isopod from the Weddell Sea, Antarctica: Coperonus pinguis n-sp (Crustacea, Isopoda, Munnopsidae). Bijdragen tot de Dierkunde 62, 55–61. Brandt, A. 1999. Dolichiscus spinosetosus n. sp., a new arcturid isopod from the Weddell Sea, Antarctica. Antarctic Science 11, 399–407. Brandt, A. 2000. Hypotheses on Southern Ocean peracarid evolution and radiation (Crustacea, Malacostraca). Antarctic Science 12, 269–275. Brandt, A. & Janssen, H. H. 1994. Redescription of Zonophyrxus quinquedens Barnard, 1913 (Crustacea, Isopoda, Dajidae) from the Weddell Sea, Antarctica, with notes on its biology and zoogeography. Polar Biology 14, 343–350. Brey, T. & Clarke, A. 1993. Population dynamics of benthic marine invertebrates in Antarctic and Subantarctic environments: are there unique adaptations? Antarctic Science 5, 253–266. Brey, T., Dahm, C., Gorny, M., Klages, M., Stiller, M. & Arntz, W. E. 1996. Do Antarctic benthic invertebrates show an extended level of eurybathy? Antarctic Science 8, 3–6. Brey, T., Klages, M., Dahm, C., Gorny, M., Gutt, J., Hain, S., Stiller, M., Arntz, W. E., Wägele, J.-W. & Zimmerman, A. 1994. Antarctic benthic diversity. Nature 368, 297 only. Broch, H. 1959. Octocorals. Part I. Pennatularians. Discovery Reports 29, 245–280. Broch, H. J. 1965. Some octocorals from Antarctic waters. Scientiﬁc Results of the “Brategg” Expedition, 1947–1948 5, 18–39. Bronsted, H. V. 1931. Die Kalkschwamme der Deutschen Südpolar-Expedition 1901–1903. Deutsche Südpolar Expedition. 1901–1903. Zoology 20, 1–47. Browne, E. T. 1902. Hydrozoa. A preliminary account. British Museum (Natural History), 310–316. Bullivant, J. S. & Dearborn, J. H. 1967. The fauna of the Ross Sea. New Zealand Department of Science and Industry Research Bulletin 176, 1–76. Burton, M. 1929. Porifera II. Antarctic sponges. British Antarctic Terra Nova Expedition, 1910. Natural History Report. Zoology 6, 393–458. Cairns, S. D. 1983. Antarctic and subantarctic Stylasterina (Coelenterata: Hydrozoa). Antarctic Research Series 38, 61–164. Cairns, S. D. 1990. Antarctic Scleractinia. Koenigstein: Koeltz Scientiﬁc Books. Cantera, J. R. & Arnaud, P. M. 1984. Les gastéropodes prosobranches des iles Kerguelen et Crôzet (sud de l’Océan indien). Comparison écologique et peculiarités biologiques. Comité National Français des Recherches Antarctiques 56, 1–169. Cantone, G. 1995. Polychaeta sedentaria of Terra Nova Bay (Ross Sea, Antarctica). Capitellidae to Serpulidae. Polar Biology 15, 295–302. Cantone, G. & Di Pietro, N. 1998. A new species of Myriochele (Polychaeta, Oweniidae) from Antarctica, with considerations on the Antarctic oweniids. Polar Biology 19, 421–423. Cantone, G. & Di Pietro, N. 2001. Benthic littoral Polychaeta ‘Sedentaria’ of Terra Nova Bay (Ross Sea, Antarctica). Antarctic Science 13, 3–8. Cantone, G. & Sanﬁlippo, R. 1991. Polychaeta from Terra Nova Bay (Ross Sea, Antarctica). In ATTI Seminario Internazionale, V. A. Gallardo et al. (eds). Concepción, Chile: ENEA-Progetto Antartide-Italia, Centro EULA-Universidad de Concepción-Chile, 371–381. Castelli, A. 1992. Primi dati sulla distribuzione batimetrica dei Policheti erranti della baia di Terra Nova (Mare di Ross, Antartide). Oebalia 17, 331–333. Cattaneo-Vietti, R., Chiantore, M., Schiaparelli, S. & Albertelli, G. 2000. Shallow- and deep-water mollusc distribution at Terra Nova Bay (Ross Sea, Antarctica). Polar Biology 23, 173–182. Chapelle, G. & Peck, L. S. 1999. Polar gigantism dictated by oxygen availability. Nature 399, 114–115.

99

ANDREW CLARKE & NADINE M. JOHNSTON

Cherbonnier, G. 1973. Invertébrés marins des XXIIème et XVème Expéditions Antarctiques Français en Terre Adélie. Holothuroides. Téthys 5, 601–610. Cherbonnier, G. & Guille, A. 1974. Echinodermata. I. Holothuroidea and Echinoidea. II. Asteroidea and Ophiuroidea. Comité National Français des Recherches Antarctiques 35, 25–44. Child, C. A. 1987. New and little known Pycnogonida from Antarctic and Sub-Antarctic waters. Proceedings of the Biological Society of Washington 100, 902–916. Child, C. A. 1994a. Antarctic and Subantarctic Pycnogonida: 1. The family Ammotheidae. Antarctic Research Series 63, 1–49. Child, C. A. 1994b. Antarctic and Subantarctic Pycnogonida: 2. The family Austrodecidae. Antarctic Research Series 63, 49–99. Child, C. A. 1995a. Antarctic and Subantarctic Pycnogonida: III. The family Nymphonidae. Antarctic Research Series 69, 1–68. Child, C. A. 1995b. Antarctic and Subantarctic Pycnogonida: IV. The families Colossendeidae and Rhychothoraxidae. Antarctic Research Series 69, 69–112. Child, C. A. 1995c. Antarctic and Subantarctic Pycnogonida: V. The families Pycnogonidae, Phoxichilidiidae, Endeididae, and Callipallenidae, including the genus Pallenopsis. Antarctic Research Series 69, 113–160. Chin, C. S., Klinkhammer, G. P., Wilson, C., Lawver, L. & Lupton, J. E. 1996. Hydrothermal activity in a nascent backarc basin. EOS Transactions of the American Geophysical Union 77, F413. Claridge, M. F., Dawah, H. A. & Wilson, M. R. eds. 1997. Species: the units of biodiversity. London: Chapman & Hall. Clark, A. H. 1937. Echinodermata. Crinoidea collected during the Australian Antarctic Expedition 1911–1914. Scientiﬁc Reports of the Australasian Antarctic Expedition 1911–1914. Series C 8, 5–18. Clark, A. M. 1962. Asteroidea. Report. B.A.N.Z. Antarctic Research Expedition 1929–1931. Series B 9, 1–104. Clarke, A. 1983. Life in cold water: the physiological ecology of polar marine ectotherms. Oceanography and Marine Biology: an Annual Review 21, 341–453. Clarke, A. 1991. What is cold adaptation and how should we measure it? American Zoologist 31, 81–92. Clarke, A. 1992. Is there a latitudinal diversity cline in the sea? Trends in Ecology & Evolution 7, 286–287. Clarke, A. 1996. Benthic marine habitats in Antarctica. In Foundations for ecological research west of the Antarctic Peninsula, E. E. Hofmann et al. (eds). Washington, DC: American Geophysical Union, 123–133. Clarke, A. 1998. Temperature and energetics: a review of cold ocean physiology. In Cold ocean physiology, H. O. Pörtner & R. C. Playle (eds). Cambridge: Cambridge University Press, 3–30. Clarke, A. In press. The polar deep seas. In Ecosystems of the world: the deep sea, P. A. Tyler (ed.). Amsterdam: Elsevier. Clarke, A. & Crame, J. A. 1989. The origin of the Southern Ocean marine fauna. In Origins and evolution of the Antarctic biota, J. A. Crame (ed.). London: The Geological Society, 253–268. Clarke, A. & Crame, J. A. 1997. Diversity, latitude and time: patterns in the shallow sea. In Marine biodiversity: causes and consequences, R. F. G. Ormond et al. (eds). Cambridge: Cambridge University Press, 122–147. Clarke, A. & Crame, J. A. In press. The importance of historical processes in global patterns of diversity. In Macroecology: reconciling divergent views, T. M. Blackburn & K. J. Gaston (eds). Oxford: Blackwell Scientiﬁc. Clarke, A. & Johnston, I. A. 1996. Evolution and adaptive radiation of Antarctic ﬁshes. Trends in Ecology & Evolution 11, 212–218. Clarke, A., Johnston, N. M. & Mitchell, R. E. 2000. Antarctic marine diversity: a status report. Cambridge: British Antarctic Survey.

100

ANTARCTIC MARINE BENTHIC DIVERSITY

Clarke, A. & Lidgard, S. 2000. Spatial patterns of diversity in the sea: bryozoan species richness in the North Atlantic. Journal of Animal Ecology 69, 799–814. Clarke, K. R. & Warwick, R. M. 1998. A taxonomic distinctness index and its statistical properties. Journal of Applied Ecology 35, 523–531. Clarke, K. R. & Warwick, R. M. 1999. The taxonomic distinctness measure of biodiversity: weighting of step lengths between heirarchical levels. Marine Ecology Progress Series 184, 21–29. Clarke, K. R. & Warwick, R. M. 2001. A further biodiversity index applicable to species lists: variation in taxonomic distinctness. Marine Ecology Progress Series 216, 265–278. Clayton, M. N. 1994. Evolution of the Antarctic marine benthic algal ﬂora. Journal of Phycology 30, 897–904. Clayton, M. N., Wiencke, C. & Klöser, H. 1997. New records of temperate and sub-Antarctic marine benthic macroalgae. Polar Biology 17, 141–149. Coleman, C. O. 1994. New Epimeria species (Crustacea: Amphipoda: Epimeriidae) and redescriptions of three other species in the genus from the Antarctic Ocean. Journal of Natural History 28, 555–576. Coleman, C. O., Adorf, L., Eikmeier, N., Hoffmann-Kobert, B., Leistikow, A., Meier, D., Petzold, D., Remmert, K., Schierl, F., Schulte, M. C., Siebensohn, E., Stavridis, T. & Wagner, K. 1994. Taxonomy of two iphimediid amphipods (Crustacea) from the Southern Ocean. Journal of Natural History 28, 1059–1075. Coleman, C. O. & Barnard, J. L. 1991. Review of the genus Pariphimedia (Crustacea: Amphipoda: Iphimediidae) with redescription of the two species from the Southern Ocean. Invertebrate Taxonomy 5, 527–539. Costello, M. J., Emblow, C. S. & Picton, B. E. 1996. Long term trends in the discovery of marine species new to science which occur in Britain and Ireland. Journal of the Marine Biological Association of the United Kingdom 76, 255–257. Crame, J. A. 1992. The evolutionary history of the polar regions. Historical Biology 6, 37–60. Crame, J. A. 1996. Evolution of high-latitude molluscan faunas. In Origin and evolutionary radiation of the Mollusca, J. D. Taylor (ed.). Oxford: Oxford University Press, 119–131. Crame, J. A. 2000a. Evolution of taxonomic diversity gradients in the marine realm: evidence from the composition of Recent bivalve faunas. Paleobiology 26, 188–214. Crame, J. A. 2000b. The nature and origin of taxonomic diversity gradients in marine bivalves. In The evolutionary biology of the Bivalvia, E. M. Harper et al. (eds). London: The Geological Society, 347–360. Crame, J. A. 2001. Taxonomic diversity gradients through geological time. Diversity and Distributions 7, 175–189. Crame, J. A. & Clarke, A. 1997. The historical component of taxonomic diversity gradients. In Marine biodiversity: causes and consequences, R. F. G. Ormond et al. (eds). Cambridge: Cambridge University Press, 258–273. Crame, J. A. & Rosen, B. R. 2002. Cenozoic palaeogeography and the rise of modern biodiversity patterns. In Palaeobiogeography and biodiversity change: the Ordovician and Mesozoic-Cenozoic radiations, J. A. Crame & A. W. Owen (eds). London: Geological Society of London. 153–168. Cronk, Q. C. B. 1989. Measurement of biological and historical inﬂuences in plant classiﬁcations. Taxon 18, 177–187. Darlington, P. J. 1965. Biogeography of the southern end of the world. Cambridge, Mass.: Harvard University Press. Dawson, E. W. 1957. Checklist of marine nemertines from Antarctica. Special Reports of the Royal Society of New Zealand Antarctic Research Committee 22, 1–10. Dayton, P. K. 1990. Polar benthos. In Polar oceanography, Part B: chemistry, biology and geology, W. O. Smith (ed.). San Diego: Academic Press, 631–685. De Broyer, C. 1984. Evolution du complexe Orchomene Boeck (Amphipoda, Lysianassoidea). Annales de la Société Royale Zoologique de Belgique 114, 197–198.

101

ANDREW CLARKE & NADINE M. JOHNSTON

De Broyer, C. 1985a. Necrophagous lysianassoid amphipods (Crustacea) from Kerguelen Is., Orchomenella guillei, n. sp. Paris. Muséum National d’Histoire Naturelle. Bulletin. Série 4 7A, 205–217. De Broyer, C. 1985b. Description of Falklandia gen. n. from the Southern Ocean and deﬁnition of the uristid group of Lysianassoidea (Crustacea, Amphipoda). Zoologica Scripta 14, 303–312. De Broyer, C. & Jazdzewski, K. 1993. Contribution to the marine biodiversity inventory. A checklist of the Amphipoda (Crustacea) of the Southern Ocean. Bruxelles. De Broyer, C. & Jazdzewski, K. 1996. Biodiversity of the Southern Ocean: towards a new synthesis for the Amphipoda (Crustacea). Bolletino del Museo Civico di Storia Naturale 20, 547–568. De Pina, G. M. A. 1995. New species of Oradarea and notes on some other eusirids from Antarctica (Crustacea, Amphipoda). Cahiers de Biologie Marine 36, 251–258. Dell, R. K. 1964a. Marine Mollusca from Macquarie and Heard Islands. Records of the Dominion Museum 4, 267–301. Dell, R. K. 1964b. Antarctic and subantarctic Mollusca: Amphineura, Scaphopoda and Bivalvia. Discovery Reports 33, 93–250. Dell, R. K. 1972. Antarctic benthos. Advances in Marine Biology 10, 1–216. Dell, R. K. 1990. Antarctic Mollusca, with special reference to the fauna of the Ross Sea. Bulletin of the Royal Society of New Zealand 27, 1–311. Dial, K. P. & Marzluff, J. M. 1989. Nonrandom diversiﬁcation within taxonomic assemblages. Systematic Zoology 38, 26–37. Dunn, D. F. 1983. Some Antarctic and sub-Antarctic sea anemones (Coelenterata: Ptychodactiaria and Actiniaria). Antarctic Research Series 39, 1–67. Dunn, D. F. 1984. More Antarctic and sub-Antarctic sea anemones (Coelenterata: Corallimorpharia and Actiniaria). Antarctic Research Series 41, 1–42. Eastman, J. T. 1993. Antarctic ﬁsh biology: evolution in a unique environment. San Diego, California: Academic Press. Eastman, J. T. & Clarke, A. 1998. A comparison of adaptive radiations of Antarctic ﬁsh with those of non-Antarctic ﬁsh. In Fishes of Antarctica: a biological overview, G. di Prisco et al. (eds). Berlin: Springer-Verlag, 3–26. Egorova, E. N. 1982. Molluscs of the Davis Sea, East Antarctica. Biological Reports on the Soviet Antarctic Expedition 1955–1958. 7, 1–143. [In Russian]. Ekau, W. & Gutt, J. 1991. Notothenioid ﬁshes from the Weddell Sea and their habitat, observed by underwater photography and television. Proceedings of the NIPR Symposium on Polar Biology 4, 36–49. El Beshbeeshy, M. 1991. Systematische, Morphologische und Zoogeographische Untersuchungen an den Thekaten Hydroiden des Patagonischen Schelfs. PhD thesis, University of Hamburg. Eldredge, L. G. & Miller, S. E. 1995. How many species are there in Hawaii? Bishop Museum Occasional Papers 41, 1–18. Fauchald, K. 1977. The polychaete worms. Deﬁnitions and keys to the orders, families and genera. Science Series. Natural History Museum of Los Angeles County 28, 1–88. Fauchald, K. & Rouse, G. 1997. Polychaete systematics: past and present. Zoologica Scripta 26, 71–138. Feldmann, R. M. & Crame, J. A. 1998. The signiﬁcance of a new nephropid lobster from the Miocene of Antarctica. Palaeontology 41, 807–814. Feldmann, R. M. & Quilty, P. G. 1997. First Pliocene decapod crustacean (Malacostraca: Palinuridae) from the Antarctic. Antarctic Science 9, 56–60. Feldmann, R. M. & Tshudy, D. M. 1989. Evolutionary patterns in macrurous decapod crustaceans from Cretaceous to early Cenozoic rocks of the James Ross Island region, Antarctica. In Origins and evolution of the Antarctic biota, J. A. Crame (ed.). London: The Geological Society, 183–195. Feldmann, R. M. & Wilson, M. T. 1988. Eocene decapod crustacean from Antarctica. In Geology and paleontology of Seymour Island, Antarctic Peninsula, R. M. Feldmann & M. O. Woodburne (eds). Boulder, Colorado: Geological Society of America, 465–488.

102

ANTARCTIC MARINE BENTHIC DIVERSITY

Fisher, W. K. 1940. Asteroidea. Discovery Reports 20, 69–306. Fogg, G. E. 1992. A history of Antarctic science. Cambridge: Cambridge University Press. Foster, M. W. 1965. Brachiopoda. New York: Geological Society of America; Lawrence: University of Kansas Press. Foster, M. W. 1974. Recent Antarctic and Subantarctic brachiopods. Antarctic Research Series 21, 1–189. Foster, M. W. 1989. Brachiopods from the extreme South Paciﬁc and adjacent waters. Journal of Paleontology 63, 268–391. Fry, W. G. & Hedgpeth, J. W. 1969. The fauna of the Ross Sea. Part 7. Pycnogonida, 1. Colossendeidae, Pycnogonidae, Endeidae, Ammotheidae. Bulletin of the New Zealand Department of Scientiﬁc and Industrial Research 198, 1–139. Gambi, M. C., Castelli, A. & Guizzardi, M. 1997. Polychaete populations of the shallow soft bottoms off Terra Nova Bay (Ross Sea, Antarctica): distribution, diversity and biomass. Polar Biology 17, 199–210. Gambi, M. C., Lorenti, M., Russo, G. F. & Scipione, M. B. 1994. Benthic associations of the shallow hard bottoms off Terra Nova Bay, Ross Sea: zonation, biomass and population structure. Antarctic Science 6, 449–462. Gambi, M. C. & Mazzella, L. 1992. Quantitative and functional studies on coastal benthic communities of Terra Nova Bay (Ross Sea, Antarctica): hard bottoms. In ATTI Seminario Internazionale, V. A. Gallardo et al. (eds). Concepción, Chile, ENEA-Progetto Antartide-Italia, Centro EULA-Universidad de Concepción-Chile, 401–415. García-Alvarez, O., Urgorri, V. & von Salvini-Plawen, L. 2001. Dorymenia troncosoi sp. nov. (Mollusca Solenogastres: Proneomeniidae), a new species from the South Shetland Islands (Antarctica). Polar Biology 20, 382–387. García-Alvarez, O., Urgorri, V. & von Salvini-Plawen, L. 2000. Two new species of Dorymenia (Mollusca: Solenogastres: Proneomeniidae) from the South Shetland Islands (Antarctica). Journal of the Marine Biological Association of the United Kingdom 80, 835–842. Gaston, K. J. 1994. Rarity. London: Chapman & Hall. Gaston, K.J. (ed.) 1996. Biodiversity: a biology of numbers and difference. Oxford: Blackwell Science. German, C. R., Livermore, R. A., Baker, E. T., Bruguier, N. I., Connelly, D. P., Cunningham, A. P., Morris, P., Rouse, I. P., Statham, P. J. & Tyler, P. A. 2000. Hydrothermal plumes above the East Scotia Ridge: an isolated high-latitude back-arc spreading centre. Earth and Planetary Science Letters 184, 241–250. Gibson, R. 1983. Antarctic nemerteans: the anatomy, distribution and biology of Parborlasia corrugatus (McIntosh, 1876) (Heteronemertea, Lineidae). Antarctic Research Series 39, 289–316. Gibson, R. 1985. Antarctic nemerteans: Heteronemertea – descriptions of new taxa, reappraisals of the systematic status of existing species and a key to the heteronemerteans recorded south of latitude 50°S. Zoological Journal of the Linnean Society 83, 95–227. Gloersen, P., Campbell, W. J., Cavalieri, D. J., Comiso, J. C., Parkinson, C. L. & Zwally, H. J. 1992. Arctic and Antarctic sea ice 1979–1987: satellite passive-microwave observations and analysis. NASA SP-511. Washington, DC: National Aeronautics and Space Administration. Gon, O. & Heemstra, P. C. 1990. Fishes of the Southern Ocean. Grahamstown, South Africa: J. L. B. Smith Institute of Ichthyology. Gotelli, N. J. & Graves, G. R. 1996. Null models in ecology. Washington and London: Smithsonian Institution Press. Grassle, J. F. 2001. Marine ecosystems. In Encyclopedia of biodiversity, S. A. Levin (ed.). San Diego: Academic Press, 13–25. Grassle, J. F. & Maciolek, N. J. 1992. Deep sea species richness: regional and local diversity estimates from quantitative bottom samples. American Naturalist 139, 313–341.

103

ANDREW CLARKE & NADINE M. JOHNSTON

Graus, R. R. 1974. Latitudinal trends in the shell characteristics of marine gastropods. Lethaia 7, 303–314. Gray, J. S. 1994. Is deep-sea species diversity really so high? Species diversity of the Norwegian coastal shelf. Marine Ecology Progress Series 112, 205–209. Gray, J. S. 1997. Marine biodiversity: patterns, threats and conservation needs. Biodiversity and Conservation 6, 153–175. Greig, J. A. 1929. Echinodermata from the Palmer Archipelago, South Shetlands, South Georgia and the Bouvet Islands. Scientiﬁc Results of the Norwegian Antarctic Expeditions (1927–1928 and 1928–1929) 2, 1–16. Gruvel, A. 1907. Cirripedes. In Natural History. National Antarctic Expedition 1901–1904. London: British Museum, 1–4. Gruvel, A. 1910. Cirripedia of the German South Polar Expedition 1901–1903. Deutsche SüdpolarExpedition. 1901–1903. Zoology 3, 193–230. Grygier, M. 1981. A representative of the genus Dendrogaster (Cirripeda: Ascothoracica) parasitic in an Antarctic starﬁsh. Antarctic Research Series 32, 1–15. Grygier, M. 1984. Ascothoracid crustacean Ascothorax gigas: redescription, larval development, and notes on its infestation of the Antarctic ophiuroid Ophionotus victoriae. Antarctic Research Series 16, 43–58. Grygier, M. J. 1987. Antarctic records of asteroid-infesting Ascothoracica (Crustacea), including a new genus of Ctenosculidae. Proceedings of the Biological Society of Washington 100, 711–712. Gutt, J. 1988. On the distribution and ecology of sea cucumbers (Holothuroidea, Echinodermata) in the Weddell Sea (Antarctica). Berichte zur Polarforschung 41, 1–87. Gutt, J., Ekau, W. & Gorny, M. 1994. New results on the ﬁsh and shrimp fauna of the Weddell Sea and Lazarev Sea (Antarctic). Proceedings of the NIPR Symposium on Polar Biology 7, 91–102. Gutt, J., Gorny, M. E. & Arntz, W. E. 1991. Spatial distribution of Antarctic shrimps (Crustacea: Decapoda) by underwater photography. Antarctic Science 3, 363–369. Gutt, J. & Koltun, V. M. 1995. Sponges of the Lazarev and Weddell Sea, Antarctica: explanations for their patchy occurrence. Antarctic Science 7, 227–234. Gutt, J. & Piepenburg, D. 1991. Dense aggregations of three deep-sea holothurians in the southern Weddell Sea, Antarctica. Marine Ecology Progress Series 68, 277–285. Gutt, J. & Schickan, T. 1998. Epibiotic relationships in the Antarctic benthos. Antarctic Science 10, 398–405. Gutt, J., Sirenko, B. I., Arntz, W. E., Smirnov, I. S. & De Broyer, C. 2000. Biodiversity of the Weddell Sea: macrozoobenthic species (demersal ﬁsh included) sampled during the expedition ANT XIII/3 (EASIZ I) with RV ‘Polarstern’. Berichte zur Polarforschung 372, 1–103. Gutt, J. & Starmans, A. 1998. Structure and biodiversity of megabenthos in the Weddell and Lazarev Seas (Antarctica): ecological role of physical parameters and biological interactions. Polar Biology 20, 229–247. Hain, S. 1990. Die beschalten Mollusken (Gastropoda und Bivalvia) des Weddellmeeres, Antarktis. Berichte zur Polarforschung 70, 1–184. Hartman, O. 1964. Polychaeta errantia of Antarctica. Antarctic Research Series 3, 1–131. Hartman, O. 1966. Polychaeta Myzostomidae and Sedentaria of Antarctica. Antarctic Research Series 7, 1–158. Hartman, O. 1967. Polychaetous annelids collected by the USNS Eltanon and Staten Island cruises, chieﬂy from Antarctic seas. Allan Hancock Monographs in Marine Biology 2, 1–387. Hartman, O. 1978. Polychaeta from the Weddell Sea quadrant, Antarctica. Antarctic Research Series 26, 125–223. Hartman, W. D. 1982. Porifera. In Synopsis and classiﬁcation of living organisms, S. P. Parker (ed.). New York: McGraw-Hill, 640–666. Hartmann-Schröder, G. & Rosenfeldt, P. 1988. Die Polychaeten der “Polarstern” – Reise ANT III/2 in die Antarktis 1984. Tiel 1: Euphrosinidae bis Chaetopteridae. Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 85, 25–72.

104

ANTARCTIC MARINE BENTHIC DIVERSITY

Hartmann-Schröder, G. & Rosenfeldt, P. 1989. Die Polychaeten der “Polarstern” – Reise ANT III/2 in die Antarktis 1984. Tiel 2: Cirratulidae bis Serpulidae. Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 86, 65–106. Hartmann-Schröder, G. & Rosenfeldt, P. 1990. Die Polychaeten der Walther Herwig-Reise 68/1 nach Elephant Island (Antarktis) 1985. Teil 1: Aphroditidae bis Cirratulidae. Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 87, 89–122. Hartmann-Schröder, G. & Rosenfeldt, P. 1992. Die Polychaeten der “Polarstern”- Reise ANTV/1 in die Antarktis 1986. Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 89, 85–124. Hawksworth, D. L. ed. 1995. Biodiversity: measurement and estimation. London: Chapman & Hall. Hayward, P. J. 1995. Antarctic cheilostomatous Bryozoa. Oxford: Oxford University Press. Hayward, P. J. 1996. New species in the Antarctic genus Toretocheilum (Bryozoa: Cheilostomatida). Journal of Natural History 30, 299–302. Hayward, P. J. & Ryland, J. S. eds. 1995. Handbook of the marine fauna of North-West Europe. Oxford: Oxford University Press. Headland, R. K. 1989. Chronological list of Antarctic expeditions and related historical events. Cambridge: Cambridge University Press. Hedgpeth, J. W. 1969. Introduction to Antarctic zoogeography. Distribution of selected groups of marine invertebrates in waters south of 35°S latitude. In Antarctic map folio series, Folio 11, V. C. Bushnell & J. W. Hedgpeth (eds). New York: American Geographical Society, 1–29. Hedgpeth, J. W. 1970. Marine biogeography of the Antarctic regions. In Antarctic Ecology, M. W. Holdgate (ed.). New York: Academic Press, 97–104. Hedgpeth, J. W. 1971. Perspectives of benthic ecology in Antarctica. In Research in the Antarctic, L. O. Quam (ed.). Washington, DC: American Association for the Advancement of Science, 93–136. Held, C. 2000. Phylogeny and biogeography of serolid isopods (Crustacea, Isopoda, Serolidae) and the use of ribosomal expansion segments in molecular systematics. Molecular Phylogenetics and Evolution 15, 165–178. Herman, R. L. & Dahms, H. U. 1992. Meiofaunal communities along a depth transect off Halley Bay (Weddell Sea – Antarctica). Polar Biology 12, 313–320. Heywood, V. H. (ed.) 1995. Global biodiversity assessment. Cambridge: Cambridge University Press. Hoek, P. P. C. 1907. Zoology: cirripedes. Résultats du voyage du S. Y. Belgica en 1897–1899 sous le commandement de A. de Gerlache de Gomery. Rapports Scientiﬁque 9, 1–9. Hopkins, T. L. 1987. Midwater food web in McMurdo Sound, Ross Sea, Antarctica. Marine Biology 96, 93–106. Howson, C. M. & Picton, B. E. (eds) 1997. The species directory of the marine fauna and ﬂora of the British Isles and surrounding seas. Belfast and Ross on Wye: Ulster Museum and The Marine Conservation Society. Hubbell, S. P. 2001. The uniﬁed neutral theory of biodiversity and biogeography. Princeton, NJ: Princeton University Press. Hubold, G. 1992. Zur Ökologie der Fische im Weddellmeer. Berichte zur Polarforschung 103, 1–157. Hurlbert, S. H. 1971. The non-concept of species diversity: a critique and alternative parameters. Ecology 52, 577–586. Huston, M. A. 1994. Biological diversity: the coexistence of species on changing landscapes. Cambridge: Cambridge University Press. Ijima, I. 1927. The Hexactinellidae of the Siboga Expedition. Siboga Expedition Monographs 6, 1–383. Jarms, G. & Tiemann, H. 1996. On a new hydropolyp without tentacles, Microhydrula limopsicola n. sp., epibiotic on bivalve shells from the Antarctic. Scientia Marina 60, 109–115. Jaya Sree, V., Sreepada, R. A. & Parulekar, A. H. 1993. An unusual giant pycnogonid (Pycnogonida, Colossendeidae) Decolopoda qasimi sp nov from Antarctic waters. Current Science 65, 179–181.

105

ANDREW CLARKE & NADINE M. JOHNSTON

Jazdzewski, K. & De Broyer, C. 1990. Morphology and systematic position of the Antarctic and Subantarctic synopiid Cardenio paurodactylus Stebbing, 1888 (Crustacea: Amphipoda). Beaufortia 41, 129–133. Jazdzewski, K., Jurasz, W., Kittel, W., Presler, E., Presler, P. & Sicinski, J. 1986. Abundance and biomass estimates of the benthic fauna in Admiralty Bay, King George Island, South Shetland Islands. Polar Biology 6, 5–16. Jazdzewski, K., Weslawski, J. M. & De Broyer, C. 1995. Comparison of the amphipod faunal diversity in two polar fjords: Admiralty Bay, King George Island (Antarctic) and Hornsund, Spitsbergen (Arctic). Polskie Archiwum Hydrobiologii 42, 382–384. John, D. D. 1938. Crinoidea. Discovery Reports 18, 121–222. Keller, N. B. 1990. On the distribution of the Madreporarian in the Antarctic and Subantarctic basins. Trudy Instituta Okeanologii 126, 74–79. Kirkwood, J. M. 1984. A guide to the Decapoda of the Southern Ocean. ANARE Research Notes 11, 1–47. Klages, M., Gutt, J., Starmans, A. & Bruns, T. 1995. Stone crabs close to the Antarctic continent: Lithodes murrayi Henderson 1888 (Crustacea, Decapoda, Anomura) off Peter I Island (68° 51 S, 90° 51 W). Polar Biology 15, 73–75. Klinkhammer, G. P., Chin, C. S., Keller, R. A., Daehlmann, A., Sahling, H., Sarthou, G., Petersen, S., Smith, F. & Wilson, C. 2001. Discovery of hydrothermal vent-sites in Bransﬁeld Strait, Antarctica. Earth and Planetary Science Letters 193, 395–407. Knowlton, N. 1993. Sibling species in the sea. Annual Review of Ecology and Systematics 24, 189–216. Knox, G. A. & Lowry, J. K. 1977. A comparison between the benthos of the Southern Ocean and the North Polar Ocean with special reference to the Amphipoda and the Polychaeta. In Polar oceans, M. J. Dunbar (ed.). Calgary: Arctic Institute of North America, 423–462. Koehler, R. 1920. Echindermata: Asteroidea. Scientiﬁc Reports of the Australasian Antarctic Expedition 1911–1914. Series C 8, 1–308. Koehler, R. 1922. Echindermata: Ophiuroidea. Scientiﬁc Reports of the Australasian Antarctic Expedition 1911–1914. Series C 8, 1–98. Koehler, R. 1926. Echindermata: Echinoidea. Scientiﬁc Reports of the Australasian Antarctic Expedition 1911–1914. Series C 8, 1–134. Koltun, V. M. 1964. Sponges of the Antarctic. Part 1. Tetraxonida and Cornacuspongida. Results of a biological investigation by the Soviet Antarctic expedition (1955–1958). Issledovaniya Fauny Morei 2, 6–133. Koltun, V. M. 1976. Porifera – Part 1: Antarctic sponges. Report. B.A.N.Z. Antarctic Research Expedition. Series B (Zoology and Botany) 9, 153–198. Kott, P. 1969. Antarctic Ascidiacea. Antarctic Research Series 13, 1–239. Kramp, P. L. 1957. Medusae. Report. B.A.N.Z. Antarctic research Expedition 1929–1931. Series B (Zoology and Botany) 6, 151–164. Kudenov, J. D. 1993. Amphinomidae and Euphrosinidae (Annelida: Polychaeta) principally from Antarctica, the Southern Oceans and Subantarctic Regions. Antarctic Research Series 58, 93–150. Kunzmann, K. 1996. Associated fauna of selected sponges (Hexactinellida and Demospongiae) from the Weddell Sea, Antarctica. Berichte zur Polarforschung 210, 1–93. Kuznetsova, I. A. & Neurova, A. M. 1986. The sea duck Lepas australis in the fouling of sub-Antarctic waters. Okeanologiya 26, 1010 only. Lambshead, P.J.D., Tietjen, J., Ferrero, T.J. & Jensen, P. 2000. Latitudinal diversity gradients in the deep sea with special reference to North Atlantic nematodes. Marine Ecology Progress Series 194, 159–167. Larrain, A. P. 1985. Brachysteraster, new genus, and Brachyaster chesheri, new species of Antarctic echinoid (Spatangoidea, Schizasteroidae). Polar Biology 4, 121–124. Laubitz, D. R. 1993. Caprellidae (Crustacea: Amphipoda): towards a new synthesis. Journal of Natural History 27, 965–976.

106

ANTARCTIC MARINE BENTHIC DIVERSITY

Lee, M. S. Y. 1997. Documenting present and past biodiversity: conservation biology meets palaeontology. Trends in Ecology & Evolution 12, 132–133. Levenstein, R. Y. 1964. Polychaeta. In Biological Reports on the Soviet Antarctic Expedition 1955–1958, A. P. Andriashev & P. V. Ushakov (eds). Jerusalem: Israel Program for Scientiﬁc Translation, 173–256. Lévi, C. 1964a. Éponges de la baie du Morbihan (Iles de Kerguelen). CNFRA Biologie 1, 149–155. Lévi, C. 1964b. Spongiaires des zones bathyale, abyssale et hadale. Galathea Report 7, 63–112. Lipps, J.H. & Hickman, C.S. 1982. Origin, age and evolution of Antarctic and deep-sea faunas. In Environment of the deep sea, W.G. Ernst & J.G. Morris (eds). Englewood Cliffs: Prentice Hall, 324–356. Lipps, J. H., Krebs, W. N. & Temnikow, N. K. 1977. Microbiota under Antarctic ice shelves. Nature 265, 232–235. Lipps, J. H., Ronan, T. E. & DeLaca, T. E. 1979. Life below the Ross Ice Shelf. Science 203, 447–449. Longhurst, A. 1998. Ecological geography of the sea. San Diego: Academic Press. López Gappa, J. 2000. Species richness of marine Bryozoa in the continental shelf and slope off Argentina (south-west Atlantic). Diversity and Distributions 6, 15–27. MacArthur, R. H. 1965. Patterns of species diversity. Biological Reviews 40, 510–533. MacArthur, R. H. 1972. Geographical ecology: patterns in the distribution of species. New York: Harper & Row. MacArthur, R. H. & Wilson, E. O. 1963. An equilibrium theory of insular zoogeography. Evolution 17, 373–387. MacArthur, R. H. & Wilson, E. O. 1967. The theory of island biogeography. Princeton, NJ: Princeton University Press. Madsen, F. J. 1967. Ophiuroidea. Report. B.A.N.Z. Antarctic Research Expedition 1929–1931. Series B 9, 123–145. Magurran, A. E. 1988. Ecological diversity and its measurement. Princeton, NJ: Princeton University Press. Margulis, L. & Schwartz, K. V. 1982. Five kingdoms: an illustrated guide to the phyla of life on earth. New York: W. H. Freeman. May, R. M. 1988. How many species are there on earth? Science 241, 1441–1449. May, R. M. 1993. Marine species richness. Nature 361, 398. May, R. M. 1994. Biological diversity: differences between land and sea. Philosophical Transactions of the Royal Society of London, Series B 343, 105–111. McKnight, D. G. 1967. Echinoderms from Cape Hallett, Ross Sea. New Zealand Journal of Marine and Freshwater Research 1, 314–323. McKnight, D. G. 1976. Asteroids from the Ross Sea and the Balleny Islands. Records of the New Zealand Oceanographic Institute, Wellington 3, 21–31. Miller, R. G. 1993. History and atlas of the ﬁshes of the Antarctic Ocean. Carson City, Nevada: Foresta Institute for Ocean and Mountain Studies. Minelli, A. 1993. Biological systematics: the state of the art. London: Chapman & Hall. Minelli, A., Fusco, G. & Sartori, S. 1991. Self-similarity in biological classiﬁcations. BioSystems 26, 89–97. Mironov, A. N. 1995. Holasteroid echinoids 2. The genus Pourtalesia. Zooologichesky ii Zhurnal 74, 59–75. (In Russian.) Moe, R. L. & DeLaca, T. E. 1976. Occurrence of macroscopic algae along the Antarctic Peninsula. Antarctic Journal of the United States 11, 20–24. Molander, A. R. 1929. Die Octactinarien. Further Results of the Swedish Antarctic Expedition 1901–1903 2, 1–86. Monniot, C. 1990. Ascides of MD 42. Terres Australes et Antarctiques Française. Mission de Recherche. Rapports des campagnes de la mer 85, 197–200. Monniot, C. & Monniot, F. 1983. Ascides antarctiques et subantarctiques: morphologie et biogeographie. Mémoires du muséum national d’histoire naturelle, Paris 125, 1–168.

107

ANDREW CLARKE & NADINE M. JOHNSTON

Monniot, C. & Monniot, F. 1994. Ascidians collected in the Weddell Sea by the RV “Polarstern” (EPOS cruise leg 3). Bulletin de muséum national d’histoire naturelle, Paris 16, 13–37. Monro, C. C. A. 1930. Polychaete worms. Discovery Reports 2, 3–218. Monro, C. C. A. 1935. Polychaete worms. 2. Discovery Reports 12, 61–198. Monro, C. C. A. 1939. Polychaeta. Report. B.A.N.Z. Antarctic Research Expedition. Series B (Zoology and Botany) 4, 87–156. Mortensen, T. H. 1936. Echinodermata and Ophiuroidea. Discovery Reports 12, 199–348. Moyano, G. H. I. 1984. On a small bryozoan collection from near Ross Island, Antarctica. Serie Cientíﬁca. Instituto Antártico Chileno 31, 75–83. Mühlenhardt-Siegel, U. 1989. Antarktische Bivalvia der reisen des FS ‘Polarstern’ und des FFS ‘Walter Herwig’ aus den jahren 1985 bis 1986. Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 86, 153–178. Munilla León, T. 1989. Two new species of Antarctic Pycnogonida. In Actas del tercer Symposium Español de Estudios Antárticos, J. Castellvi (ed.). Gredos: Comision Interministerial de Ciencia y Tecnologia, Madrid, 74–79. Munilla León, T. 2001a. Synopsis of the pycnogonids from Antarctic and Subantarctic waters. Polar Biology 24, 941–945. Munilla León, T. 2001b. A new species of Ammothea (Pycnogonia) and other pycnogonids from Livingston Island and surrounding waters (South Shetland Islands, Antarctica). Antarctic Science 13, 144–149. Murina, V. V. 1975. Taxonomic rank and geographic distribution of abyssal priapulids. Trudy Instituta Okeanologii 103, 154–161. (In Russian.) Newman, W. A. & Ross, A. 1971. Antarctic Cirripedia; monographic account based on specimens collected chieﬂy under the United States Antarctic research program, 1962–1965. Antarctic Research Series 14, 1–257. Nicol, D. 1964. An essay on the size of marine pelecypods. Journal of Paleontology 38, 968–974. Nicol, D. 1966a. Descriptions, ecology, and geographic distribution of some Antarctic pelecypods. Bulletin of American Paleontology 51, 1–102. Nicol, D. 1966b. Size of pelecypods in recent marine faunas. Nautilus 79, 109–113. Nicol, D. 1971. Species, class and phylum diversity of animals. Quarterly Journal of the Florida Academy of Sciences 34, 191–194. Nicol, D. 1978. Size trends in living pelecypods and gastropods with calcareous shells. Nautilus 92, 70–79. Nielsen, C. 1995. Animal evolution: interrelationships of the living phyla. Oxford: Oxford University Press. Nilsson-Cantell, C. A. 1930a. Thoracic cirripedes collected in 1925–1936. Discovery Reports 2, 223–260. Nilsson-Cantell, C. A. 1930b. Zoologische Ergebnisse der Reisen von Rr. L. Kohl-Larsen nach den subantarktischen Inseln bei Neuseeland und nach Sudgeorgien. No. 2. Cirripedien von der StewartInsel und von Sudgeorgien. Senckenbergiana 12, 210–213. Nilsson-Cantell, C. A. 1939. Thoracic cirripedes collected in 1925–1936. Discovery Reports 18, 223–238. Numanami, H. 1996. Taxonomic study on Antarctic gastropods collected by Japanese Antarctic Research Expedition. Memoirs of the National Institute of Polar Research, Series E (Biological and Medical Sciences) 39, 1–244. Numanami, H., Okutani, T., Iwami, T., Takeuchi, I., Igarashi, A., Tsuchiya, Y. & Fukuchi, M. 1996. A record of gastropods and bivalves collected from East Antarctica by the Japanese Antarctic Research Expeditions in 1992–1994. Proceedings of the NIPR Symposium on Polar Biology 9, 207–223. Oliver, P. G. & Picken, G. B. 1984. Prosobranch gastropods from Signy Island: Buccinacea and Muricacea. British Antarctic Survey Bulletin 62, 95–115.

108

ANTARCTIC MARINE BENTHIC DIVERSITY

Orensanz, J. M. 1990. The eunicemorph polychaete annelids from Antarctic and Subantarctic Seas with addenda to eunicemorphs of Argentina, New Zealand, Chile, Australia and the South Indian Ocean. Antarctic Research Series 52, 1–183. Orsi, A. H., Whitworth, T. & Nowlin, W. D. 1995. On the meridonal extent and fronts of the Antarctic Circumpolar Current. Deep Sea Research 42, 641–673. Ostrovskii, A. N. & Taylor, P. D. 1996. Systematics of some Antarctic Idmidronea and Exidmonea (Bryozoa: Cyclostomata). Journal of Natural History 30, 1549–1575. O’Sullivan, D. 1982. A guide to the Hydromedusae of the Southern Ocean and adjacent waters. ANARE Research Notes 5, 1–136. Pansini, M., Calcinai, B., Cattaneo-Vietti, R. & Sarà, M. 1994. Demosponges from Terra Nova Bay (Ross Sea, Antarctica). In Italy, National Scientiﬁc Commission for Antarctica. Ross Sea expeditions 1987–1988 and 1989–1990 and Magellan expedition 1991: Data Report Part III: physical, chemical and biological oceanography, F. Faranda & L. Guglielmo (eds). Genova: National Scientiﬁc Commission for Antarctica, 67–84. Pasternak, F. A. 1961. Pennatularians (Octocorallia) and antipatharians (Hexacorallia) collected by the Soviet Antarctic expedition of 1955–1958. Trudy Instituta Okeanologii 46, 217–230. (In Russian.) Pasternak, F. A. 1975. New data on the speciﬁc composition and distribution of the deep-sea pennatularians (Octocorallia, Pennatularia) of the Peru-Chile region and South Atlantic. Trudy Instituta Okeanologii 103, 101–118. (In Russian.) Pasternak, F. A. 1993. Coelenterata collected in the South Atlantic Ocean. Trudy Instituta Okeanologii 127, 82–88. (In Russian.) Paulay, G. 1997. Diversity and distribution of reef organisms. In Life and death of coral reefs, C. Birkeland (ed.). New York: Chapman & Hall, 298–353. Pawson, D. L. 1977. Molpadiid sea cucumbers (Echinodermata: Holothuroidea) of the South Atlantic, Paciﬁc and Indian Oceans. Antarctic Research Series 26, 97–123. Peña Cantero, A. L. 1997a. Species of Staurotheca Allmann, 1888 (Cnidaria: Hydrozoa) from recent Antarctic expeditions with R. V. Polarstern, with the description of six new species. Journal of Natural History 31, 31–32. Peña Cantero, A. L. 1997b. On Antarctoscyphus (Cnidaria, Hydrozoa), a new genus of antarctic hydroids and description of two new species. Polar Biology 81, 23–32. Peña Cantero, A. L. 1998a. Two new Antarctic species of the genus Schizotricha Allman, 1883 (Cnidaria, Hydrozoa). Polar Biology 19, 77–84. Peña Cantero, A. L. 1998b. On two new species of Oswaldella Stechow, 1919 (Cnidaria, Hydrozoa) from Bransﬁeld Strait (Antarctica). Polar Biology 20, 33–40. Peña Cantero, A. L. 1998c. Two new species of the genus Schizotricha Allman, 1883 (Cnidaria, Hydrozoa). Polar Biology 19, 77–84. Peña Cantero, A. L. & Garcia Carrascosa, A. M. 1991. Hidrozoos de la campana “Antartida 8611”. In 4th Spanish Symposium on Antarctic Studies, J. Castellvi (ed.). Madrid: Comisión Interministerial de Ciencia y Technologia, 117–140. Peña Cantero, A. L., Garcia Carrascosa, A. M. & Vervoort, W. 1995. On two species of Oswaldella Stechow, 1919: O. shetlandica Stepan’yants, 1979 and O. elongata spec. nov. (Cnidaria Hydrozoa) (Notes on Antarctic Hydroids, II). Zoologische Mededeelingen 60, 341–351. Peña Cantero, A. L., Garcia Carrascosa, A. M. & Vervoort, W. 1996. A full description of Selaginopsis vanhoeffeni Peña Cantero and Garcia Carrascosa, 1994 (Sertulariidae: Hydrozoa, Leptomedusae) with notes on related species. Journal of Natural History 30, 1–10. Peña Cantero, A. L., Garcia Carrascosa, A. M. & Vervoort, W. 1997a. On Antarctoschyphus (Cnidaria, Hydrozoa), a new genus of Antarctic hydroids and the description of two new species. Polar Biology 18, 23–32. Peña Cantero, A. L., Svoboda, A. & Vervoort, W. 1997b. Species of Staurotheca Allman, 1888 (Cnidaria: Hydrozoa) from recent Antarctic expeditions with R. V. Polarstern, with the description of six new species. Journal of Natural History 31, 329–381.

109

ANDREW CLARKE & NADINE M. JOHNSTON

Peña Cantero, A. L., Svoboda, A. & Vervoort, W. 1997c. Species of Oswaldella Stechow, 1919 (Cnidaria, Hydrozoa) from recent Antarctic expeditions with R. V. ‘Polarstern’, with the description of eight new species. Zoological Journal of the Linnean Society 119, 339–388. Peña Cantero, A. L. & Vervoort, W. 1995. Redescription of Oswaldella antarctica (Jaderholm, 1904) (Cnidaria Hydrozoa) with notes on related species (Notes on Antarctic hydroids, I). Zoologische Mededeelingen 69, 329–340. Peña Cantero, A. L. & Vervoort, W. 1998. On two new species of Oswaldella Stechow, 1919 (Cnidaria, Hydrozoa) from Bransﬁeld Strait (Antarctica). Polar Biology 20, 33–40. Picken, G. B. 1980. The nearshore prosobranch gastropod epifauna of Signy Island, South Orkney Islands. PhD thesis, University of Aberdeen. Pielou, E. C. 1969. An introduction to mathematical ecology. New York: John Wiley. Piepenburg, D., Voss, J. & Gutt, J. 1997. Assemblages of sea stars (Echinodermata: Asteroidea) and brittle stars (Echinodermata: Ophiuroidea) in the Weddell Sea (Antarctica) and off Northeast Greenland (Arctic): a comparison of diversity and abundance. Polar Biology 17, 305–322. Pirez, A. M. S. & Sumida, P. Y. G. 1997. The valviferan isopods (Crustacea Peracarida) from Bransﬁeld Strait and adjacent waters, Antarctica. Ophelia 46, 11–34. Platnick, N. I. 1992. Patterns of biodiversity. In Systematics, ecology and the biodiversity crisis, N. Eldredge (ed.). New York: Columbia University Press, 15–24. Ponder, W. F. 1971. Some New Zealand and Subantarctic bivalves of the Cyamaicea and Leptonacea with descriptions of new taxa. Records of the Dominion Museum 7, 119–141. Ponder, W. F. 1983. Rissoaform gastropods from the Antarctic and Sub-antarctic. Scientiﬁc Reports. British Antarctic Survey 108, 1–96. Ponder, W. F. & Lindberg, D. R. 1996. Gastropod phylogeny – challenges for the 90s. In Origin and evolutionary radiation of the Mollusca, J. D. Taylor (ed.). Oxford: Oxford University Press, 135–154. Ponder, W. F. & Lindberg, D. R. 1997. Towards a phylogeny of gastropod molluscs: an analysis using morphological characters. Zoological Journal of the Linnaean Society 119, 83–265. Poore, G. C. B. & Wilson, G. D. F. 1993. Marine species richness. Nature 361, 597–598. Poulin, E. & Féral, J.-P. 1996. Why are there so many species of brooding Antarctic echinoids? Evolution 50, 820–830. Poulin, E., Palma, A. T. & Féral, J.-P. 2002. Evolutionary versus ecological success in Antarctic benthic invertebrates. Trends in Ecology & Evolution 17, 218–222. Powell, A. W. B. 1960. Antarctic and subantarctic Mollusca. Records of the Auckland Institute and Museum 5, 117–193. Powell, A. W. B. 1973. The patellid limpets of the world (Patellidae). Indo-Paciﬁc Mollusca 3, 75–206. Preston, F. W. 1962. The canonical distribution of commonness and rarity. Ecology 43, 185–212. Pugh, P. J. A. 1993. A synonymic catalogue of the Acari from Antarctica, the sub-Antarctic islands and the Southern Ocean. Journal of Natural History 27, 323–421. Pushkin, A. F. 1976. Revision of the Pycnogonida (Pantopoda) of the Pallenopsis patagonica group in Antarctica and neighbouring waters. Sovetskaya Antarkticheskaya Ekspeditsiya. Information Bulletin 8, 682–689. Rauschert, M. & Andres, H. G. 1993. Scaphodactylus, a new stenothoid genus from the sublittoral of South Shetland Islands (Crustacea: Amphipoda: Gammaridea). Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 69, 347–358. Reaka-Kudla, M. L. 1997. The global diversity of coral reefs: a comparison with rain forests. In Biodiversity II: understanding and protecting our biological resources, M. L. Reaka-Kudla et al. (eds). Washington, DC: National Academy of Sciences, 83–108. Ren, X. Q. & Huang, L. Q. 1989. Short report of the Cirripeda (Crustacea) in the waters northwest of the Antarctic Peninsula. In First China Symposium on Southern Ocean Expedition. Hangzhou: Shanghai Science and Technology Publisher, 190–196. (In Chinese.) Retamal, M. A., Quintara, R. & Neira, F. 1983. Qualitative and quantitative analysis of benthic

110

ANTARCTIC MARINE BENTHIC DIVERSITY

communities in Foster Bay, Deception Island. Serie Cientíﬁca. Instituto Antártico Chileno 29, 5–15. Rex, M. A., Stuart, C. T., Hessler, R. R., Allen, J. A., Sanders, H. L. & Wilson, G. D. F. 1993. Global scale latitudinal patterns of species diversity in the deep-sea benthos. Nature 365, 636–639. Richardson, M. D. 1976. The classiﬁcation and ordination of marine macrobenthic assemblages at Arthur Harbor, Anvers Island, Antarctica. PhD thesis, Oregon State University. Robilliard, G. A. & Dayton, P. K. 1972. A new species of platyctenean ctenophore, Lyrocteis ﬂavopallidus sp. nov., from McMurdo Sound, Antarctica. Canadian Journal of Zoology 50, 47–52. Rosenzweig, M. L. 1995. Species diversity in space and time. Cambridge: Cambridge University Press. Rosso, A. 1990. Bryozoan community of IB3 Station (Ross Sea, Antarctica). In Italy, National Scientiﬁc Commission for Antarctica. Ross Sea expeditions 1987–1988: Data Report Part I: physical, chemical and biological oceanography, F. Faranda & L. Guglielmo (eds). Genova: National Scientiﬁc Commission for Antarctica, 428–430. Roy, K., Jablonski, D. & Valentine, J. W. 1994. Eastern Paciﬁc molluscan provinces and latitudinal diversity gradient: no evidence for “Rapoport’s rule”. Ecology 91, 8871–8874. Roy, K., Jablonski, D. & Valentine, J. W. 1996. Higher taxa in biodiversity studies: patterns from eastern Paciﬁc marine molluscs. Philosophical Transactions of the Royal Society of London, Series B 351, 1605–1613. Roy, K., Jablonski, D. & Valentine, J. W. 1998. Marine latitudinal diversity gradients: tests of causal hypotheses. Proceedings of the National Academy of Sciences of the United States of America 95, 3699–3702. Saiz-Salinas, J.I. 1995. Sipuncula of the southeastern Weddell Sea (Antarctica). Polar Biology 15, 307–317. Saiz-Salinas, J. I., Dean, H. K. & Cutler, E. B. 2000. Echiura from Antarctic and adjacent waters. Polar Biology 23, 661–670. San Martin, G. & Parapar, J. 1997. “Errant” polychaetes of the Livingston Island shelf (South Shetlands, Antarctica), with the description of a new species. Polar Biology 17, 285–295. Sarà, M., Balduzzi, A., Bartbieri, M., Bavestrello, G. & Burlado, B. 1992. Biogeographic traits and checklist of Antarctic demosponges. Polar Biology 12, 559–585. Seno, J. & Irimura, S. 1968. Ophiuroidea collected from around the Ross Sea in 1964 with a description of a new sp. Reports of the 3rd Antarctic Expedition 1964–1965 of the T. S. Umitaka-Maru. Journal of the Tokyo University of Fisheries. Special Edition 9, 147–154. Sicinski, J. 1986. Benthic assemblages of Polychaeta in chosen regions of the Admiralty Bay (King George Island, South Shetland Islands). Polish Polar Research 7, 63–78. Sicinski, J. & Janowska, E. 1993. Polychaetes of the shallow sublittoral of Admiralty Bay, King George Island, South Shetland Islands. Antarctic Science 5, 161–167. Sieg, J. 1983. Tanaidomorpha (Crustacea, Tanaidacea) from the Ross Sea, Antarctica. Journal of the Royal Society of New Zealand 13, 395–418. Sieg, J. 1984a. Neue Erkenntnisse zum naturlichen System der Tanaidacea. Eine phylogenetische Studie. Zoologica Stuttgart 136, 1–132. Sieg, J. 1984b. Tanaidacea of the United States Navy’s 1947–1948 Antarctic Expedition (Crustacea). Journal of Crustacean Biology 4, 298–306. Sieg, J. 1986a. Crustacea Tanaidacea of the Antarctic and the Subantarctic. 1. On material collected at Tierra del Fuego, Isla de los Estados, and the west coast of the Antarctic Peninsula. Antarctic Research Series 45, 1–180. Sieg, J. 1986b. Distribution of the Tanaidacea: synopsis of the known data and suggestions on possible distribution patterns. Crustacean Issues 4, 165–194. Sieg, J. 1986c. Tanaidacea (Crustacea) von der Antarctic und Subantarctic. Mitteilungen aus dem Museum University Kiel 2, 1–80. Sieg, J. 1988. Das phylogenetische System der Tanaidacea und die Frage nach Alter und Herkunft der Crustaceanfauna des antarktischen Festlandsockels. Zeitschrift für Zoologische Systematik und Evolutionsforschung 26, 363–379.

111

ANDREW CLARKE & NADINE M. JOHNSTON

Sieg, J. & Wägele, J. W. (eds) 1990. Fauna der Antarktis. Berlin: Paul Parey. Smirnov, R. V. 2000. A new species of Spirobrachia (Pogonophora) from the Orkney Trench (Antarctica). Polar Biology 23, 567–570. Snelgrove, P. V. R. 2001. Marine sediments. In Encyclopedia of biodiversity, S. A. Levin (ed.). San Diego: Academic Press, 71–84. Snelgrove, P. V. R. & Smith, C. R. 2002. A riot of species in an environmental calm: the paradox of the species-rich deep-sea ﬂoor. Oceanography and Marine Biology: an Annual Review 40, 311–342. Speel, J. A. & Dearborn, J. H. 1983. Comatulid crinoids from R. V. Eltanin cruises in the Southern Ocean. Antarctic Research Series 38, 1–60. Sreepada, R. A., Jaya Sree, V. & Parulekar, A. H. 1995. Benthic faunal composition along Princess Astrid Coast, East Antarctica. Scientiﬁc Report of the Tenth Indian Expedition to Antarctica. Technical Report 3, 184–185. Steiner, G. & Linse, K. 2000. Systematics and distribution of the Scaphopoda (Mollusca) in the Beagle Channel (Chile). Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut 97, 13–30. Stiboy-Risch, C. 1992. New species of Ammotheidae from Antarctica. Bonner Zoologische Beitrage 43, 333–338. (In German.) Stiller, M. 1996. Distribution and biology of the aphroditides and polynoids (Polychaeta) in the eastern Weddell Sea and the Lazarev Sea (Antarctica). Berichte zur Polarforschung 185, 1–193. (In German.) Stork, N. E. 1988. Insect diversity: facts, ﬁction and speculation. Biological Journal of the Linnaean Society 35, 321–337. Thomas, P. A. & Mathew, K. J. 1986a. Sponges collected during the third Indian Antarctic Research expedition. Scientiﬁc Report of the Third Indian Expedition to Antarctica. Technical Report 3, 109–116. Thomas, P. A. & Mathew, K. J. 1986b. Primnoisis spicata (Hickson) (Order Gorgonacea Lmx. Family Isididae Lmx.) from the Antarctic Sea. Scientiﬁc Report of the Third Indian Expedition to Antarctica. Technical Report 3, 129–132. Thomson, J. A. & Rennet, N. 1931. Alcyonaria, Madreporia and Antipatharia. Scientiﬁc Reports of the Australasian Antarctic Expedition, 1911–1914 9, 1–46. Thurston, M. H. 1972. The Crustacea Amphipoda of Signy Island, South Orkney Islands. Scientiﬁc Reports. British Antarctic Survey 71, 1–133. Thurston, M. H. 1989. A new species of Valettia (Crustacea: Amphipoda) and the relationship of the Valettidae to the Lysianassoidea. Journal of Natural History 23, 1093–1107. Tiefenbacher, L. 1990. Eualus kinzeri, a new hippolytid shrimp from the Weddell Sea (Antarctica). Spixiana 13, 117–120. Trathan, P. N., Brandon, M. A. & Murphy, E. J. 1997. Characterisation of the Antarctic Polar Frontal Zone to the north of South Georgia in summer 1994. Journal of Geophysical Research 102, 10483–10497. Trathan, P. N., Brandon, M. A., Murphy, E. J. & Thorpe, S. E. 2000. Transport and structure within the Antarctic Circumpolar Current to the north of South Georgia. Geophysical Research Letters 27, 1727–1730. Tréguer, P. & Jacques, G. 1992. Dynamics of nutrients and phytoplankton, and ﬂuxes of carbon, nitrogen and silicon in the Antarctic Ocean. Polar Biology 12, 149–162. Turpaeva, E. P. 1974. Pycnogonids of the Scotia Sea and surrounding waters. Trudy Instituta Okeanologii 98, 277–305. (In Russian.) Turpaeva, Y. P. 1990. Pycnogonida from South Atlantic. Trudy Instituta Okeanologii 126, 108–126. Utinomi, H. 1965. Giant Antarctic barnacle Hexelasma antarcticum Borradaile (Cirripedia, Thoracica). Japanese Antarctic Research Expedition Scientiﬁc Reports. Series E. Biology 25, 1–15. van der Land, J. 1970. Systematics, zoogeography and ecology of the Priapulida. Zoologische Verhandelingen 112, 1–118. Vinogradova, N. G. 1990. Data on quantitative characteristics of sea-bed fauna in the region of Elephant Island (South Shetland Islands). Trudy Instituta Okeanologii 126, 40–44.

112

ANTARCTIC MARINE BENTHIC DIVERSITY

von Salvini-Plawen, L. 1979. Antarktische und subantarktische Solenogastres. Stuttgart: E. Schweizerbart’sche Verlagsbuchhandlung. Voss, J. 1988. Zoogeographie und Gemeinschaftsanalyse des Makrozoobenthos des Weddellmeeres (Antarktis). Berichte zur Polarforschung 45, 1–145. Wägele, J. W. 1989. Evolution und phylogenetisches System der Isopoda. Sand der Forschung und neue Erkenntnisse. Zoologica, Stuttgart 140, 1–262. Wakabara, Y., Tararam, A. & Miyagi, V. K. 1995. The amphipod fauna of the west Antarctic region (South Shetland Islands and Bransﬁeld Strait). Polskie Archiwum Hydrobiologii 42, 347–365. Walsh, J. J. 1988. On the nature of continental shelves. San Diego: Academic Press. Walters, S. M. 1986. The name of the rose: a review of ideas on the European bias in angiosperm classiﬁcation. New Phytologist 104, 527–546. Warén, A., Arnaud, P. M. & Cantera, J. R. 1986. Descriptions of two new gastropods of the Trichotropidae from Kerguelen and Crozet Islands (South Indian Ocean). Veliger 29, 157–165. Warwick, R. M. & Clarke, K. R. 2001. Practical measures of marine biodiversity based on relatedness of species. Oceanography and Marine Biology: an Annual Review 39, 207–231. Watling, L. & Thurston, M. H. 1989. Antarctica as an evolutionary incubator: evidence from the cladistic biogeography of the amphipod family Iphimediidae. In Origins and evolution of the Antarctica biota, J. A. Crame (ed.). London: The Geological Society, 297–309. Webb, C. O. 2000. Exploring the phylogenetic structure of ecological communities: an example for rain forest trees. American Naturalist 156, 145–155. Whittaker, R. H. 1972. Evolution of measurements of species diversity. Taxon 21, 213–251. Whittaker, R. H. 1977. Evolution of species diversity in land communities. In Evolutionary biology, M. K. Hecht et al. (eds). New York: Plenum, 1–67. Wiencke, C. & Clayton, M. N. 2002. Antarctic seaweeds. Rugell, Lichtenstein: A. R. G. Gantner Verlag KG. Williams, R. B. 1981. Sea anemone, Edwardsia meridionalis sp. nov., from Antarctica and a preliminary revision of the genus Edwardsia de Quatrefages, 1841 (Coelenterata: Actiniaria). Records of the Australian Museum 33, 325–360. Willis, J. C. 1922. Age and area: a study in geographical distribution and origin of species. Cambridge: Cambridge University Press. Willis, J. C. & Yule, G. U. 1922. Some statistics of evolution and geographical distribution in plants and animals, and their signiﬁcance. Nature 109, 177–179. Wilson, E. O. 1988. The current state of biological diversity. In Biodiversity, E. O. Wilson & F. M. Peter (eds). Washington, DC: National Academy Press, 3–18. Wilson, E. O. 1992. The diversity of life. Cambridge, Mass.: The Belknap Press of Harvard University Press. Wilson, E. O. & Peter, F. M. (eds) 1988. Biodiversity. Washington, DC: National Academy Press. Winston, J. E. 1992. Systematics and marine conservation. In Systematics, ecology and the biodiversity crisis, N. Eldredge (ed.). New York: Columbia University Press, 144–168. Winston, J. E. 1994. Bryozoa of the US Antarctic Research Program: Preliminary report. In Biology and palaeobiology of bryozoans, P. J. Hayward et al. (eds). Fredensborg: Olsen & Olsen, 205–210. World Conservation Monitoring Centre 1992. Global biodiversity: status of the earth’s living resources. London: Chapman & Hall. Wu, B. & Zhao, J. 1986. On the occurrence of priapulids in the Antarctic Great Wall Bay. Journal of Oceanography of Huanghai and Bohai Seas 4, 38–41. Yoon, C. K. 1993. Counting creatures great and small. Science 260, 620–622. Young, P. S. & Leta, A. C. 1996. Antarctic Cirripeda (Crustacea) collected by the Brazilian Antarctic Program (1983–1986). Iheringia Serie Zoologia 80, 121–126. Zeidler, W. 1994. New information and locality records for the Antarctic amphipod Clarencia chelata K. H. Barnard, 1931, and a reappraisal of the family Clarenciidae J. L. Barnard & Karaman, 1987 (Amphipoda, Gammaridea). Crustaceana 66, 219–226.

113

ANDREW CLARKE & NADINE M. JOHNSTON

Zevina, G. B. 1964. Cirripedia of the genus Scalpellum Leach, collected by the Soviet Antarctic Expedition of the diesel electric ‘Ob’ in Antarctic and sub-Antarctic waters. Issledovaniya Fauny Morei 2, 252–253. Zevina, G. B. 1968. New and rare species of Cirripedia Thoracica from the Antarctic. Issledovaniya Fauny Morei 6, 85–96. (In Russian.) Zevina, G. B. 1990. Deep-sea barnacles (Cirripeda, Thoracica) in the South Atlantic. Trudy Instituta Okeanologii 126, 80–89. (In Russian.)

Addendum Recently, Julian Gutt and colleagues (pers. comm., manuscript under review) have used data from quantitative bottom samples taken in the Weddell Sea to estimate how many species might exist on the entire continental shelf of Antarctica. A variety of species-accumulation, jack knife and incidence-based average techniques led to estimates in the range 11 000 to 17 000. Gutt and colleagues propose that the entire Southern Ocean macrobenthic fauna of the continental shelves probably exceeds 17 000 taxa.

114

Oceanography and Marine Biology: an Annual Review 2003, 41, 115–159 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

INFLUENCE OF MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES I. COLOMBINI 1 & L. CHELAZZI 2 Dipartimento di Biologia Animale e Genetica “Leo Pardi”, Università di Firenze, Via Romana 17, 50125 Florence, Italy 2 Istituto per lo Studio degli Ecosistemi del CNR Sezione di Firenze, Via Romana 17, 50125 Florence, Italy e-mail:[email protected]

1

Abstract This review provides an overview of the importance of beach accumulations of macrophytes and other organic beach-cast material on the ecology of sandy beach ecosystems. It describes the composition of these allochthonous subsidies, their abundance on beaches in relation to seasonal, lunar, tidal and spatial trends, their decomposition and utilisation by bacterial, meio- and macrofaunal communities. The paper then analyses the community structure and the species succession in both macrophyte wrack and carrion and reports the most important ﬁndings on individual wrack-inhabiting species (amphipods, isopods, dipterans). Other aspects, such as feeding and microclimatic preferences of certain species and their interactions in wracks, are also discussed. Links to vertebrate species and other secondary consumers that exploit beach-cast macrophytes and carrion as trophic resource are considered, and the importance of wrack in recycling nutrients to nearshore coastal ecosystems is stressed. The beneﬁcial and detrimental effects of organic beach-cast material on both plants and animals of beach and nearshore communities and on the geomorphology of coastal beach-dune systems are pointed out. Another section is dedicated to human use of beach-cast macrophytes through harvesting of economically important species and of other stranded material through its exploitation for traditional reasons. The effects of harvesting on local faunal communities and on the stability of the dunes is discussed. A ﬁnal section of the paper includes the positive and negative effects of man-made debris on sandy-beach ecosystems and brieﬂy reviews the major ﬁndings.

Introduction The ocean/land interface occupies about 8% of the earth’s surface (Ray & Hayden 1992) along 594 000 km of coastline (Hammond 1990). In this ecotone terrestrial and aquatic habitats interact intensively altering salinity, turbidity, nutrients and climatic regimens in coastal waters through freshwater runoffs on the one hand and affecting productivity of coastal areas through a large input of material from the adjacent aquatic systems on the other. The spatial boundaries are continuously crossed by the basic components of food webs. Ecologist are now more aware of how ecosystems are closely bound to one another and how factors outside a system may signiﬁcantly affect or even dominate local patterns and dynamics. Polis et al. (1997) deﬁned this exchange of organic matter between habitats “a spatial subsidy”, as 115

I. COLOMBINI & L. CHELAZZI

a donor-controlled resource (prey, detritus, nutrient) from one habitat to a recipient (plant or consumer) from a second habitat which increases population productivity of the recipient, potentially altering consumer resource dynamics in the recipient system. Physical (wind and water) and biotic (mobile consumers) vectors are mainly responsible for cross-boundary transportation of nutrients both towards sea and land. Compared with rocky and estuarine mud-ﬂats, sandy beaches have very little in situ primary production (Inglis 1989, Brown & McLachlan 1990). The resident primary producers are represented by epi-psammic diatoms. On ﬂats of ﬁne sand found on sheltered beaches these may contribute to some primary productivity but never with high values (from 0 g C m2 yr1 to 50 g C m2 yr1 according to beach exposure, Leach 1970, Hartwig 1978). Thus macrofaunal communities rely almost entirely on organic inputs that arrive from the surf zone (surf diatoms, ﬂagellates) or from the sea (stranded macrophytes, carrion, dissolved organics, particulates) through oceanic processes, such as upwelling, currents, waves and tides. Primary consumers, such as suspension feeders (crabs, bivalves) and herbivores (crustaceans, insects) consuming phytoplankton, particulate organic material, kelp, and seagrass, then become prey of secondary consumers (crabs, beetles). In turn, vertebrate predators (ﬁshes, lizards, shorebirds, foxes and other scavenger species) prey upon both primary and secondary consumers and are largely responsible for the consumption of drift carrion. Hence allochthonous resources, which in many cases are ephemeral and present seasonal and spatial ﬂuctuations, are likely to affect all levels of the food web of a sandy beach. Marine inputs on beaches all over the world, however, include both organic and inorganic material. The latter, mainly composed of man-made debris of different origin, recently has become a growing concern in many countries. Floating, submerged and stranded beach litter is more than an aesthetic issue, causing a signiﬁcant threat to marine mammals, seabirds, turtles and ﬁshes either through entanglement or ingestion (Laist 1987). From a recreational standpoint the economic value of tourist resorts may be reduced by these factors (Ryan & Moloney 1990) and, in some cases, the debris may become hazardous to human health (Philipp et al. 1993, 1994). Unfortunately, management strategies induced by stranded litter on beaches frequently have a negative impact on beach communities. The aim of this paper is to review the most recent and major ﬁndings upon organic and inorganic beach-cast material. Organic allochthonous input is analysed from several aspects, starting from its composition, abundance on beaches, processes of decomposition and exploitation by meio- and macrofaunal communities. Species succession is discussed for both macrophyte beach cast and carrion and the most important ﬁndings on individual wrack-inhabiting species are reported. Attention is given to intra- and interspeciﬁc interactions in wracks and to feeding preferences of arthropod species. The term “wrack” in this review refers not only to plants or stranded seaweed but is applied generally to any organic beach-cast material. The inﬂuence of organic beach-cast material on both terrestrial vertebrate species (shorebirds and other land species) and offshore consumers (ﬁlter-feeders, grazers, ﬁshes) is also considered. The paper then analyses the beneﬁcial and detrimental effects of wrack availability on animals and on the geomorphology of beaches, the exploitation of organic beach casts and their use by humans. The effects of removal of this material from beaches on invertebrate populations is also discussed. A ﬁnal section is dedicated to inorganic beach-cast material. Because of its impact on beaches it has attracted the attention of scientists worldwide. The major studies are 116

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

reviewed, and the positive and negative effects of man-made debris on sandy-beach ecosystems are brieﬂy discussed.

Organic components of beach-cast material Macrophytes Algae In many coastal areas the production of marine macrophytes in offshore beds is extremely high. This production consists mainly of large brown algae, commonly referred to as kelps, and seagrasses. The dominant orders of the algae belong to the Laminariales (technically kelps) and Fucales. While the intertidal zone is inhabited primarily by the fucoids, the subtidal is dominated by laminarians. Because of their high productivity and complicated biological structure kelps have received much attention (Mann 1972, 1973, 1982, Field et al. 1977, 1980a,b, Velimirov et al. 1977, Chapman & Craigie 1978, Newell et al. 1982, Kirkman 1984, Dayton 1985, Kirkman & Kendrick 1997). Species of Laminaria occur worldwide and dominate both sides of the Atlantic and the coasts of China and Japan. In New Zealand, Australia and South Africa many kelp forests are dominated by Ecklonia species whereas the giant kelp, Macrocystis pyrifera is found in the northeast Paciﬁc, southern Australia, New Zealand, South Africa, Peru, Chile, and Argentina where it forms dense forests (Dayton 1985, Kirkman & Kendrick 1997). Environmental factors inﬂuencing kelp communities, such as light, substrata and sediment, nutrients, water motion, salinity and temperature together with population dynamics, life-history studies, patch dynamics, dispersal and grazers have been reviewed by Dayton (1985). An estimate of 12% of kelp biomass or 6% of kelp production has been reported to break free during storms (Grifﬁths & StentonDozey 1981, Jarman & Carter 1981) much of which (2.5%) is deposited annually on beaches (Koop et al. 1982b). The signiﬁcance of kelp-derived organic carbon to the nearshore secondary production has also been assessed (Duggins et al. 1989). Kelp biomass enters the nearshore food web by releasing particulate and dissolved organic matter during growth and senescence and provides a signiﬁcant resource of carbon and nitrogen for the assemblages of both pelagic and benthic suspension feeders (Stuart et al. 1982, Seiderer & Newell 1985, Mann 1988, Fielding & Davis 1989, Bustamante et al. 1995, Bustamante & Branch 1996). Also, the considerable quantity of stranded kelp on beaches may re-enter the nearshore food web as particulate and dissolved organic matter after decomposition (Robertson & Hansen 1982, Walker et al. 1988, Duggins et al. 1989).

Seagrasses Another consistent part of the organic beach cast is represented by freshly-detached seagrasses. Seagrasses are the sole marine representatives of the Angiospermae that by the Eocene had widely disappeared in the Asian–Paciﬁc and the neo-tropics. Three families (Cymodoceaceae, Posidoniaceae, Zosteraceae), belonging to the Order Potamogetonales 117

I. COLOMBINI & L. CHELAZZI

contain several widespread genera (Amphibolis, Cymodocea, Halodule, Syringodium, Thalassodendron, Posidonia, Zostera, Heterozostera) (classiﬁcation according to den Hartog 1970, Tomlinson 1982, Kuo & McComb 1989). Posidonia grows on sandy or silty substrata from the intertidal zone to a depth of 40 m according to the amount of photosynthetic light that it receives. It is considered a poor coloniser as it colonises only a few square metres over 10 yr to 50 yr (Kirkman 1985, Kirkman & Kuo 1990). Nine species of Posidonia are currently recognised. P. oceanica occurs throughout the Mediterranean and Atlantic Ocean and eight other species are found in western Australia, two of which (P. sinuosa and P. angustifolia) form the largest meadows (Kirkman & Kendrick 1997). Meadows of Zosteraceae are not extensive where Posidonia is the main meadow-forming species. In sheltered estuaries and intertidal mudﬂats, however, considerable areas can be taken up by zosteraceans (Hemminga & Nieuwenhuize 1991, Kirkman & Kendrick 1997). Species belonging to the genera Thalassodendron and Syringodium, commonly found along the coasts of the Indian Ocean, made up an average of 88 18% of the beach cast along the Kenyan coast with Thalassodendron ciliatum as the dominant species (Hemminga et al. 1995, Ochieng & Erftemeijer 1999, Obura 2001). This species, which attaches to both hard and soft substrata, provides an important habitat for many coral reef species. An extensive description of the composition of the seagrass beds in the Seychelles reported Thalassia hemprichii as the dominant species (Taylor & Lewis 1970). Direct grazing on seagrasses is generally limited (Thayer et al. 1984). The major part of the biomass ages and dies and forms the basis for the detritial food chain. A part of the organic material is retained in situ and recycled within the seagrass beds. Several workers (Kikuchi 1974, Adams 1976, Lenanton et al. 1982, Robertson & Lenanton 1984) have demonstrated that the associated invertebrate community provides sustenance for ﬁshes that use seagrasses as important nursery grounds. Another part of the seagrass biomass is exported from the meadows as dissolved organic matter or partially fragmented leaves. In this form it is transferred to submerged depressions on bare sand throughout the intertidal and subtidal areas or further offshore to the deep ocean ﬂoor where it contributes to the carbon and nutrient cycles (Josselyn et al. 1983, Fry & Virnstein 1988). Furthermore, storms, associated wave action and heavy swells remove huge amounts of seagrass material and accumulate them along shores forming large wrack banks (Lenanton et al. 1982, Brown & McLachlan 1990). Thus the accumulation of beach-cast material is a result of the interactions between dense nearshore seagrass meadows and physical factors such as winds, currents, waves and tides that determine their exposure to water motion (Ochieng & Erftemeijer 1999).

Abundance of stranded macrophytes The abundance of stranded macrophytes on beaches is extremely variable and depends on the vicinity to rocky shores and reefs or to seagrass meadows. Marsden (1991a) working on a sandy beach in New Zealand calculated an average monthly wet weight of organic material of 11.25 kg 5 m1 (SE mean 3.78 kg) indicating a low organic input. On other occasions inputs can be very high. An estimate of 2179 kg m1 yr1 of kelp wrack was calculated for a beach on the western coast of South Africa (Stenton-Dozey & Grifﬁths 1983) a value very similar to that established for a nearby rocky shore, namely 1200–1800 kg m1 yr1 (Koop & Field 1980). On another South African beach near Port Elisabeth the total wrack input estimated was 2920 kg m1 yr1 (McLachlan & McGwynne 1986) whereas on beaches near Perth, western Australia 240 t dw km1 coastline yr1of detached plants were processed 118

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

through the sandy beach system (Hansen 1983). In southern California an input of 473 kg ww m1 yr1 of macrophyte wrack was estimated for a beach by Hayes (1974) and on 15 midriff islands of the Gulf of California values of algal wrack deposited ranged between 1000–2000 kg m1 yr1 (Polis et al. 1997). Recently, Hobday (2000) in the Southern California Bight showed that there was an instantaneous abundance of 39 000 to 348 000 drifting rafts of kelp (Macrocystis pyrifera), which potentially could be stranded on beaches between late winter and early spring. Seasonal, lunar, tidal and spatial ﬂuctuations in beach wrack accumulations have been recorded by many authors (Messana et al. 1977, Koop & Field 1980, Behbehami & Croker 1982, Stenton-Dozey & Grifﬁths 1983, Hansen 1984, Moggi 1987, Chelazzi et al. 1990, Marsden 1991a, Ochieng & Erftemeijer 1999, Colombini et al. 2000). Studies carried out on South African beaches showed that input of kelp debris had a clear cyclical pattern of deposition with a higher standing stock during winter but a lower standing stock during summer periods (Koop & Field 1980, Stenton-Dozey & Grifﬁths 1983). Along the beaches of the Mombasa Marine National Park and Reserve in Kenya, Ochieng & Erftemeijer (1999) showed that accumulation of beach cast was markedly seasonal with largest amounts observed during the SE monsoon (March to October) and minimal amounts during the NE monsoon (November to March). Moggi (1987), on a Somalian beach, showed that the composition of the beach cast varied according to the season with a dominance of Thalassodendron ciliatum during October–November and a more or less equal quantity of Syringodium isoetifolium, Thalassodendron ciliatum and Sargassum sp. during July–August. The different wrack compositions in the two periods of study were again explained by the opposite prevailing winds (NE and SW monsoons in the two periods, respectively) and by the presence of seagrass meadows of Thalassodendron ciliatum to the north of the study area (Colombini et al. 2000). Beach debris accumulation may also be regulated by lunar and tidal phases. Wrack mounds can be moved up and down the beach from neap to spring tides and in some cases can be entirely replaced in a semi-lunar period (Messana et al. 1977, StentonDozey & Grifﬁths 1983, Colombini et al. 2000). During this period the wrack beds are subjected to dehydration, ageing, sand covering, fragmentation and decomposition. On a beach on the west coast of Cape Peninsula, South Africa, Grifﬁths & Stenton-Dozey (1981) showed that rates of dehydration were higher for individual kelp fronds than those of artiﬁcially prepared banks. The moisture content of a single plant fell to 25% on day 6 (neap tide) while that of the bank was still 55% after the same period of time. However, both kinds of deposits lost water rapidly at ﬁrst and then more gradually as the outer surface hardened. Furthermore, there was a very rapid decline in dry mass with kelp banks declining to 50% of their initial mass by day 2 and to 20% by day 14 (spring tide), whereas individual plants lost 65% of their initial dry mass over a fortnight. A marked tidal inﬂuence was also shown on the quantity of beach-cast material with greater accumulations during spring tide periods compared with neap tide periods (Ochieng & Erftemeijer 1999). Studies on the longshore distribution of beach-cast material showed that beach wrack was not uniformly distributed but often accumulated in speciﬁc sectors of the beach (Koop & Field 1980, Hansen 1984) or adjacent to rocky protrusions (Ochieng & Erftemeijer 1999). Over the width of the beach, following periods of high wave exposure, wracks can be fairly evenly distributed or can be deposited along one or more drift lines, usually at high water springs and in bands or in a band down to the level of the most recent high tide (Marsden 1991a, Ochieng & Erftemeijer 1999). On other occasions drift material can accumulate into patches from the extreme high water of spring tide to mean tidal levels (Marsden 1991a, Colombini et al. 2000). 119

I. COLOMBINI & L. CHELAZZI

Stranding of other organic material Besides algae and seagrasses, the organic material of beach cast can be composed of a heterogeneous quantity of wood fragments, fruits, seeds and carrion. This kind of material is highly erratic and its stranding on beaches depends mainly on currents, winds and wave action.

Driftwood Stranding of large logs is generally common on beaches close to ﬂuvial systems or to areas where beach dune erosion is high. This phenomenon is quite common over the Mediterranean, where beach dune systems, with their associated Mediterranean maquis, and historically important reforested areas (pinewood plantations) have been entirely devastated by processes of erosion due to the construction of ports, jetties, breakwaters and other coastal structures. Most of the wood material is taken offshore and then cast ashore where it tends to concentrate in speciﬁc areas of the coast. Large stranded driftwood becomes an important habitat to many beach arthropods (amphipods, isopods, chilopods, coleopteran insects) and vertebrates (toads, lizards, snakes, mice) that use it as temporary or permanent shelter (Caussanel 1970, Colombini & Chelazzi 1991) or as food source as in certain xylophagous insects. In Europe the curculionid, Mesites aquitanus, for example, feeds and carries out its entire life cycle exclusively in pinewood which previously drifted in the sea, whereas the amphipod Macarorchestia remyi and larvae of the scarabaeid, Callicnemis latreillei, are directly conﬁned to the sand underlying large beached logs. Here they ﬁnd suitable microclimatic conditions to grow and develop into the adult phase (Caussanel 1965, 1970). In other cases, scavenger species, such as Phytosus nigriventris, can be indirectly tied to driftwood as it feeds on fungus colonies associated with decomposing wood debris (Caussanel 1965, Moore & Legner 1976, Colombini et al. 2002). In South Africa, Callan (1964) reported that driftwood communities were mainly composed of beetles (Carabidae and Staphylinidae). Occasionally, however, the oedemerid beetle Apterosessinia peringueyi dominated. The larvae tunnelled directly into the wood, whereas the adults were found associated with the sand underneath.

Fruits and seeds Stranding of fruits and seeds together with macrophytes in beach-cast material is a common phenomenon in the tropics. At times these can be found mixed within the wracks or scattered along the beach. On a Somali beach, Moggi (1987) reported consistent quantities of fruits belonging to the Sterculiaceae, Pandanaceae, Palmae families, seeds of Mimosaceae family and more rarely mangrove fruits, seeds and hypocotyls in the wrack beds of the beach. Most seeds belonged to species not found among the ﬂora of Somalia and probably originated along the coasts of Tanzania and Kenya arriving on the beach by means of ocean currents. Seeds provide the vital genetic link and primary dispersal agent between successive generations of plants (Gunn & Dennis 1973). An array of plant species have propagules that are adapted for dispersal by sea and, especially on many oceanic islands, sea-dispersed species account for a disproportionate number of the total number of plant species. These propag120

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

ules are very buoyant with thick protective shells that are impervious to salt water. Tropical drift seeds and fruits are remarkable because they can survive months or even years at sea. Strandings of disseminules on beaches, originating from distant sources, are highly discontinuous and the rate of deposition of ﬂoating seeds can depend on surface winds and currents (Green 1999). The best known plant drifter is the coconut (Cocos nucifera). According to Dennis & Gunn (1971) this species is indigenous to the Indo–Malaysian region and has been spread to many Paciﬁc Island groups by sea currents. It was calculated that 3000 miles is the average maximum distance that a coconut can remain aﬂoat and still be viable. Coconuts establish on well-drained coral sand beaches of tropical islands and atolls with adequate rainfall and temperatures more than on shores of continents. Rosengarten (1986) assessed that naturally dispersed coconuts can withstand occasional brief saltwater ﬂooding when developing and that coconut palms obtain freshwater and mineral nutrients from a lens of freshwater that literally ﬂoats above the denser salt water beneath the beach sand. Generally both fruits and seeds represent a highly nutritional resource for the macrofauna of beach communities. On the beach of Sar Uanle, Somalia, stranded coconuts were exploited by several tenebrionid species both as refuge and food source. When the internal seed was accessible to the beetles, this constituted a precious food source, even on a highly subsidised beach like Sar Uanle (I. Colombini & L. Chelaizzi, pers. obs.).

Carrion Carrion represents another important component of beach-cast material. Drift carrion is common on most beaches, where it is mainly represented by jellyﬁshes, siphonophores, chondrophores, bivalve mussels, tunicates, ﬁshes, turtles, seabirds, cetaceans and other animals (Brown & McLachlan 1990). Polis & Hurd (1996a) estimated that beached carrion converts to more secondary productivity than an equal mass of algal detritus. In fact, the production efﬁciency (mass gained by consumers/mass of food) of terrestrial poikilothermic consumers of ﬂesh is about twice that of herbivores and ﬁve times that of detritivores (Braﬁeld & Llewellyn 1982). Compared with macrophyte and land productivity, carrion represents a minor food source. However, this material becomes extremely important on beaches where allochthonous input is low or on beaches near colonies of pinnipeds and seabirds (Lord & Burger 1984a,b). Attempts to measure carrion input have been made in only a few studies where calculated values of dry mass of animal carcass were 110 g m1 yr1, 340 g m1 yr1 and 530 g m1 yr1 in three islands of the Gulf of California and 120 g C m1 yr1 on a beach in the eastern Cape (McGwynne 1980, Polis & Hurd 1996a). On islands, the enormous amount of allochthonous input far exceeds the autochthonous terrestrial productivity on a per-square-metre basis (i.e. many islands worldwide receive more energy from the sea than from land plants) (Lord & Burger 1984a, Burger 1985, Polis & Hurd 1996a, Polis et al. 1997). Furthermore, seabirds transport a tremendous quantity of nutrients to land via guano, eggs, feathers, bodies of dead chicks and food scraps. Marine input of carrion supports many land detritivore, scavenger and predator populations, and this subsidy allows secondary consumers to maintain relatively larger coastal population densities. However, consumer dynamics cannot be solely explained by local productivity and more than one conduit of energy ﬂow must be considered. Morritt (2001) reported greater populations of the amphipod Orchomene nanus, a crab carrion specialist, in the South Basin of Lough Hyne (Ireland) and correlated its distribution with that of the 121

I. COLOMBINI & L. CHELAZZI

different crab species likely to provide a predictable source of carrion. The hydrographical conditions of the site (greater water movements) may also have been partly responsible for the higher population densities of the amphipod, as they helped the transport of organic molecules released from carrion, which attracted more O. nanus. In Baja California del Norte (Mexico), carrion and marine prey made up about 50% of the trophic needs of coastal coyotes and the coastal population densities were higher than those of inland areas (Rose & Polis 1998). Furthermore, the study showed that the allochthonous input depressed in situ prey. In fact, the increase of these marine-subsidised coyotes (Canis latrans) depressed local rodent populations whereas in other areas (islands), lacking large predators, higher populations of rodents were supported. In the Bering Sea, scavenging on walrus (Odobenus rosmarus) carcasses allowed red fox (Vulpes vulpes) populations to reach densities 10 times greater on islands compared with mainland populations (Hersteinsson & MacDonald 1982, Zabel & Taggart 1989). More examples of marine subsidised consumers include Arctic foxes (Alopex lagopus), which exploited littoral animals such as crabs, molluscs and sea urchins (Andriashek et al. 1985) and scavenged the remains of polar bear (Ursus maritimus) kills (Sheldon 1991). The same was true for black-back jackals (Canis mesomeles), lions (Panthera leo) and brown hyenas (Hyaena brunnea) which foraged on marine-based carrion along the African coastline and reached higher population densities compared with inland areas (Bridgeford 1985, Avery et al. 1987, Hiscocks & Perrin 1987; see also Moore 2002). In other cases, the behavioural responses of subsidised vertebrates change in relation to carrion. For example, on Round Island (Alaska) Zabel & Taggart (1989) found a direct relationship between polygyny of red foxes and abundance of food resources. They showed that foxes depended on nesting adult birds and chicks, and these were both preyed upon or scavenged on the beaches. When the nesting failure of seabirds occurred in relation to the occurrence of El Niño in the Bering Sea, there was a shift from facultative polygyny to monogamy in the red fox population. In another study on a Mediterranean coastal population of the red fox, an analysis of scats showed that beach carrion was deliberately included in the diet but other items, such as arthropods (amphipods, isopods, tenebrionids, dipterans and dermapterans), might have been accidentally ingested with this diet and not actively searched for by the foxes (Ricci et al. 1998). In a study of a massive wreck of guillemots (Uria aalge) on two beaches in Resurrection Bay, Alaska, scavenger species such as bald eagles (Haliaeetus leucocephalus), northern ravens (Corvus corax), glaucous-winged gulls (Larus glaucescens), river otters (Lutra canadensis), wolverine (Gulo gulo) and American black bears (Ursus americanus), were the main agents responsible for carcass removal, whereas beach to beach transport appeared to be of minimal importance (van Pelt & Piatt 1995). An inverse relationship between the persistence of beach-cast guillemots and their degree of freshness was also shown with fresh carcasses being searched for and actively removed within a few days after deposition. For invertebrate species there are quite a few behavioural studies that analyse feeding behaviour and competition for carrion on sandy shores of Hong Kong (Morton 1990, Britton & Morton 1992, 1993, 1994, Cheung 1994, Morton & Yuen 2000). These works have mainly considered the gastropod, Nassarius festivus, which feeds on moribund bivalves, ﬁshes and decapods washed ashore, and report the opportunistic behaviour of this species, including long-distance detection, fast locomotion towards the food source and rapid ingestion of large amounts relative to its body weight. Intra- and interspeciﬁc competition for carrion also occurs in many species (Britton & Morton 1993, 1994), and recently, through ﬁeld and laboratory experiments (Morton & 122

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

Yuen 2000), these interactions have been demonstrated between two sympatric scavenger species, N. festivus and the hermit crab Diogenes edwardsii. In another study it was stressed that carrion could serve as a “bank” for species. Pugh & MacAlister (1994) found that whalebone debris, a semi-artiﬁcial permanent substratum that had a humid temperature-buffered microclimate, formed a reservoir from which several species of mites could colonise adjacent supralittoral, littoral and nearshore habitats. Another important allochthonous input, which is not truly carrion but needs to be mentioned, is the spawning of certain species in the intertidal zone of sandy beaches. In the Delaware Bay, shorebirds such as red knot (Calidris canutus), ruddy turnstone (Arenaria interpres), sanderling (Calidris alba) and semipalmated sandpiper (Calidris pusilla) stop over and feed on horseshoe crab (Limulus polyphemus) eggs, as well as infaunal benthic invertebrates, and store enough fuel in the form of fat and muscle protein to complete spring migration towards breeding grounds in the Arctic (Castro & Myers 1993, Botton et al. 1994, Tsipoura & Burger 1999). The study of the distribution and abundance of migrant shorebirds in Delaware Bay (Clark et al. 1993) indicated that annual variations could be related to nutrient ﬂuctuations and these ﬂuctuations could inﬂuence stopover patterns of migrants.

Breakdown of beach-cast material Leaching and microbial processes, meiofauna Once macrophytes are cast ashore they undergo physical processes of fragmentation, decomposition and remineralisation by bacteria, meiofauna and grazers. The fragments and mineralised components are then transported to the nearshore marine environment or to the atmosphere or stored in situ within the beach. Valiela et al. (1985), in describing the decomposition of saltmarsh grasses using the litter bag technique, identiﬁed three distinct phases. Initially, the organic material is lost at a fast rate and leaching of hydrolysed compounds from the plant material is the major mechanism. Microbial degradation of the organic matter is said to be the major source of weight loss in the second stage, while the third stage is characterised by a slower decay of refractory material. However, this pattern was not so obvious in Inglis’ (1989) work where the loss of algal dry weight was essentially linear and this difference was explained by the presence of a decay-resistant fraction in the algal detritus. The importance of microbial regeneration of nutrients from the decomposition of stranded macrophytes has been assessed by several authors (McLachlan et al. 1981, Koop et al. 1982a,b, Koop & Grifﬁths 1982, Newell et al. 1982, Stenton-Dozey & Grifﬁths 1983, McLachlan 1985). Bacteria assume paramount importance in the energetics of the sandy beach and from standing stock estimates and annual turnover of the different components of the biota it was calculated that bacteria account for as much as 87% of annual beach production (Koop & Grifﬁths 1982). Koop et al. (1982b) showed that 90% of leachates derived from the stranded kelp Ecklonia maxima was utilised by bacteria. Calculations on carbon ﬂow indicated that 23–27% of the carbon in kelp was converted to bacterial carbon and the residual, which was not incorporated into bacteria, was mineralised by the sandy beach microbes within 8 days. These authors stressed that the microbial community occupied a 123

I. COLOMBINI & L. CHELAZZI

central role in the rapid regeneration of inorganic material. Bacteria are also important as they colonise faeces produced by primary consumers and convert them to a suitable food source for ﬁlter feeders (Newell et al. 1982). Subsequently all this material is returned to the sea and supports detrital-based nearshore foodwebs (McLachlan et al. 1981, Duggins et al. 1989). However, on a South African beach, Koop et al. (1982a) estimated that only 0.4 g m1 d1 of carbon and 0.2 g m1 d1 of nitrogen is returned to the sea. They suggested that primary production in the adjacent subtidal communities was more dependent on in situ remineralisation of nutrients and on local upwelling than on nutrients originating from beach-cast materials. Also in question was whether sandy beaches accumulated nitrogen by incorporation into microbial or faunal biomass or in inorganic form in the groundwater. This controversial issue was addressed by McLachlan & McGwynne (1986) who concluded that beaches can act as nitrogen sinks or sources for nutrients depending on their state. Eroding and stable beaches are unlikely to be nitrogen sinks, while prograding beaches, where ground water ﬂow is not high enough to ﬂush nutrients from the system faster than the sand accumulates them, probably act as sinks. High levels of interstitial nutrients associated with the remineralisation of beach-cast materials have also been reported at other beaches of western Australia (McLachlan 1985). Comparing two modally reﬂective beaches, with and without large amounts of wrack deposits, higher values of the dry biomass of the macro- and interstitial fauna were found when beach cast was present. This difference was due to the less dynamic conditions of the beach and to the enrichment of the interstitial system by wrack leachates. Furthermore, there was an inverse correlation between both bacterial and protozoan numbers and meiofauna numbers suggesting that grazing by the latter kept down bacteria densities. Studies on the meiofauna (McIntyre 1968, McLachlan 1977, 1985, Koop & Grifﬁths 1982, Brown & McLachlan 1990) and of its zonation (McLachlan 1980) indicate nematodes and oligochaetes as dominant taxa, and that their distribution patterns are directly related to the distribution of the wrack, below which concentrations of dissolved organic material (DOM) were exceedingly high (Koop & Grifﬁths 1982). This distribution suggests that the meiofauna used the leachates as a food source directly, but the possibility that the DOM is used initially by the bacteria, which in turn are used as a food source by the meiofauna cannot be precluded (McLachlan 1985). However, beach-wrack accumulation can also affect infauna adversely by restricting oxygen exchange and according to McLachlan (1985), meiofauna did not occur in areas of dense wrack accumulations. McGwynne et al. (1988) concluded that interstitial meiofauna can be affected by the toxic effects of hydrogen sulphide, low pH and low oxygen concentrations in the porewater under the wracks. Alkemade & van Rijswijk (1993), studying the decomposition of stranded seaweed along the Antarctic coast, found an interesting correlation of the number of nematodes with the height of the location (relative to water), the carbon : nitrogen ratio and the salinity of the sample. Larger nematode populations occurred as the nitrogen content increased relative to carbon content, and for material stranded at higher locations on the beach. The negative correlation between salinity and nematode numbers was presumed from the limited population growth in situations with high tidal inﬂuences. Nematode numbers were also indirectly dependent on sediment composition and water content with higher numbers found in situations with melt water runoff. Although numerically the meiofauna exceed the macrofauna, with an average ratio of 105 : 1, the former never attains the high biomass of the latter due to their small size. Estimates of mean biomass ratios therefore approximate to 1 : 5 (McLachlan 1985). Based on 124

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

production estimates, the meiofauna and macrofauna were of approximately equal importance on four exposed sandy beaches in South Africa, where meiofauna were dominant on two beaches and macrofauna on the other two (McLachlan 1977). Although beaches supporting high concentrations of macrofauna are generally associated with a sparser meiofauna (McIntyre 1968), McLachlan (1977) concluded that the meiofauna and the macrofauna were quite distinct components of the beach fauna and suggested that the meiofauna, even if not part of the macrofauna food chain, was of great quantitative importance in the energy ﬂow. The interstitial system thus has the prime function of processing the organic materials having heterotrophic bacteria at the base and predatory meiofauna at the apex of the food chain. Within this process, nutrients are then returned to the sea (Brown & McLachlan 1990). Rates of litter decomposition are highly variable and are inﬂuenced by both site- and time-speciﬁc environmental conditions and in most cases depend on the composition of the wracks themselves. Smith & Foreman (1984) individually tested ten of the most important seaweeds within a seaweed community in the southern Strait of Georgia (British Columbia, Canada) using the litter bag technique. They showed that rates of decomposition of submerged samples varied among species with the time required for litter to disappear from bags ranging from 6 days to 70 days. This rate depended mainly on the contents of crude ﬁbre and detritus particle size. More rapid decomposition occurred in those species with lower crude ﬁbre content and decreasing particle size, probably due to a greater surface area of particulate material exposed to microbial attack. Similar ﬁndings were recorded for seagrass leaves where rates of decomposition increased in relation to an increase of the rates of leaching and microbial degradation as particle size of the detritus decreased (Harrison & Mann 1975, Robertson & Mann 1980). On an East African coast, leaves of the seagrass Thalassodendron ciliatum took 42 days to lose 50% of their initial ash-free dry weight in litter bag experiments (Ochienga & Erftemeijer 1999). A slight reduction in the decomposition rate of seagrass material was obtained when litter bags were buried and the reduction was attributed to a reduced exposure to wave action at high tide and to reduced drying by the sun, oxygen availability and grazing by the fauna. In a mudﬂat area in Mauritania, Hemminga & Nieuwenhuize (1991), with this same technique, showed that the time required for 50% weight loss of Zostera and Cymodocea leaf litter was 158 days and 50 days, respectively, in the intertidal zone, whereas these ﬁgures were reduced to 49 days and 37 days in the subtidal zone. In another study on the decomposition of stranded seaweed along the Antarctic coast (Alkemade & van Rijswijk 1993), it was demonstrated that abiotic locationrelated characteristics, such as the water content of the stranded seaweed and sediment composition, strongly inﬂuenced the decomposition rates. These rates were high when water content of the deposits was high. Furthermore, tidal inundation and melt-water ﬂows increased weight losses of seaweed by carrying away small fragments into the sea. With lower water content, the debris dried out and there was a decrease in bacterial respiration, leading to a lower rate of breakdown of the decaying material (Newell et al. 1985). In stranded algal deposits found on a sandy beach of South Africa, Koop et al. (1982b) estimated that kelp decomposed and was replaced over an 8-day cycle. Only a small amount of this organic material was consumed by the macrofauna, whereas 94.2% of nitrogen from the kelp debris was incorporated into the bacteria and this incorporation was accompanied by a mineralisation of approximately 70% of the carbon (Koop et al. 1982a). In a similar experiment on the giant kelp Macrocystis pyrifera, Inglis (1989) calculated that within 18 days the algae in the litter bags had lost 41–64% of their initial dry weight and that the involvement 125

I. COLOMBINI & L. CHELAZZI

of the macrofauna was insigniﬁcant. This result contrasts with the ﬁndings of Grifﬁths & Stenton-Dozey (1981) who estimated that grazers removed 60–80% of the organic input from their study site in 20 days. Even if the above ﬁndings are controversial, the importance of detrital grazers on decomposition rate of macrophytes is generally accepted. Macrofaunal consumers such as isopods, amphipods and dipteran larvae fragmentise the detritus, accelerate decomposition through the spread of bacteria and make the material more available to decomposition through their burrowing-activities (Robertson & Mann 1980, Stenton-Dozey & Grifﬁths 1980, Harrison 1982, Bedford & Moore 1984, Inglis 1989). Also, nematodes and other meiofauna can be of considerable importance in re-elaborating the debris and their activity can stimulate bacterial metabolism and lead to a rapid decay of beach-cast material (Heip et al. 1985). However, leaching of compounds as polyphenols and tannins from seagrass leaves, which have been implicated as grazing inhibitors (Harrison 1982), may be needed before detritivores will begin to fragment the deposits. Harrison (1982) has demonstrated how minor constituents of Zostera marina leaves control rates of fragmentation and microbial decay by limiting the growth of algae and bacteria and the activity of amphipod grazers. On the other hand, grazers such as the amphipod Gammarus locusta can slow down or even inhibit algal decomposition by selectively removing rotting weeds (Bedford & Moore 1984).

Macrofaunal beach-wrack communities and species succession Herbivores, detritivores and predators of stranded macrophytes Numerous studies report on the macrofaunal community of macrophyte wracks from different parts of the world (Bigot 1970, Moore & Legner 1973, Grifﬁths & Stenton-Dozey 1981, Behbehani & Croker 1982, Grifﬁths et al. 1983, Stenton-Dozey & Grifﬁths 1983, Lavoie 1985, McLachlan 1985, Inglis 1989, Chevin 1998, Ochieng & Erftemeijer 1999, Colombini et al. 2000, de Rougemont 2000, Dugan & Hubbard in press). The fauna of beach-cast wrack is generally diverse due to location, beach morphology, season, climate and vegetation cover. In Madagascar, Bigot (1970) found a richer and more abundant fauna in wrack deposits higher up on the beach with tenebrionids and spider species dominant. On a South African beach a total of 22 species of Coleoptera, three Diptera and two Amphipoda were identiﬁed in the kelp wrack (Grifﬁths & Stenton-Dozey 1981). In New Zealand, Inglis (1989) distinguished 22 macrofaunal species of which six species (an amphipod, a dipteran, a centipede and three beetles) made up 93% of the individuals of the wrack-bed community. In northern New England, the beach wrack community was dominated by an amphipod species (Orchestia platensis) and oligochaetes that made up 86% of the total numbers (Behbehani & Croker 1982). These were followed by Collembola species, while predatory species constituted only 2% of the total. De Rougemont (2000), studying the beetle community associated with seaweed jetsam in Hong Kong, found that the fauna was less diverse in tropical and subtropical areas compared with temperate regions (e.g. 11 kelp ﬂy predators and ﬁve marine staphylinids in Hong Kong compared with 22 and 11, respectively, in Japan, Shibata (1993)) and that other families (Carabidae, Scarabaeidae, Tenebrionidae), abundant in Europe, appeared to be absent in the tropics. On a reﬂective beach of 126

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

western Australia wracks were generally poor in species and the only important macrofaunal species living in the wracks was the amphipod Allorchestes compressa (McLachlan 1985), also found in abundance in the drifting wracks of the surf zone (Robertson & Lucas 1983). Colombini et al. (2000), analysing the community structure of wrack beds along a Somalian beach, found that species composition and abundance varied according to the season with a dominant (90%) amphipod species (Talorchestia martensii) in October–November and a more diverse fauna, composed of amphipods, isopods, gastropods and coleopteran staphylinid species in July–August. In contrast to other works, where dipteran species were the major components of the beach-cast material (Stenton-Dozey & Grifﬁths 1980, Inglis 1989, Phillips & Arthur 1994, Phillips et al. 1995a,b, Hodge & Arthur 1997), in the wracks of the Somalian beach dipteran larvae were completely absent. This absence was related to the short residence time of the beach wrack (10 day cycle) that would not permit successful dipteran breeding (Dobson 1974a,b, Stenton-Dozey & Grifﬁths 1980). In a survey of 15 beaches of Southern California, Dugan & Hubbard (in press) calculated that the biomass of talitrid amphipods, Megalorchestia spp. ranged from 1 g m1 to 378 g m1 and composed most of the wrack-associated macrofaunal biomass at all but two sites. On the latter beaches, the isopod Tylos punctatus was the major component. These authors found a positive correlation between the number of wrack-associated species (especially Coleoptera species) and the total cover and the % cover of macrophyte wrack on ungroomed beaches. Furthermore, species richness of wrack-associated macrofauna varied signiﬁcantly between groomed and ungroomed beaches with very low values on mechanically cleaned beaches. Also, mean abundance of wrack-associated macrofauna was nine times greater on ungroomed beaches with low wrack cover than on groomed beaches. The authors showed the importance of wrack subsidies because they increased the total macrofaunal abundance on beaches by supporting an increased abundance of wrack-associated macrofauna and major consumers, such as amphipods.

Species succession in macrophyte wracks Ephemeral resource units, such as wrack-bed communities and carrion, are generally characterised by a successional change of species with different groups associated with different qualitative stages of decomposition and ageing of the wrack. Lavoie (1985) suggested that the major invertebrate groups could be subdivided into “early”, “mid” and “late” in relation to their appearance in the wrack. A temporal separation of arrival at a speciﬁc food source may have a number of important consequences with regard to subsequent species interaction and community development (Hodge et al. 1996, Hodge & Arthur 1997). In some cases, early arrival can confer a competitive advantage, with late species encountering severe inhibition and early species being unaffected by the subsequent presence of the later species. Another process consists of an interspeciﬁc facilitation, with late species beneﬁting from the presence of the earlier arrival of primary colonists. These early arrivals modify the resource qualitatively and make it more suitable for the later species (Connell & Slatyer 1977, Schoenly & Reid 1987, Heard 1994). Finally, a temporal separation may simply result in a reciprocal avoidance, reducing levels of interspeciﬁc interference at the food source. Amphipods are considered to be the primary colonisers of newly deposited wracks because they have a tendency to forage on freshly stranded material where high densities are reached (Moore & Legner 1973, Grifﬁths & Stenton-Dozey 1981, Behbehani & Croker 127

I. COLOMBINI & L. CHELAZZI

1982, Inglis 1989, Marsden 1991b, Colombini et al. 2000). A comparative study of the mean zonations of different wrack-inhabiting species showed that amphipods were the group closest to the sea and gradually changed their zonation according to the semi-lunar cycle (Colombini et al. 2000). Other early wrack colonisers are isopod species, such as Littorophiloscia tropicalis (Chelazzi & Ferrara 1978, Colombini et al. 2000). This species uses wracks as a refuge and a food source and during the nocturnal low tides leaves the wrack mounds and moves around like other terrestrial intertidal isopods (Kensley 1974). In New Zealand, adult Diptera colonised wrack beds within the ﬁrst 24 h (Inglis 1989) whereas Lavoie (1985) suggested that all the dipterans found on Californian wrack beds occurred within the ﬁrst 4 days. Similar results were obtained for species of British seaweed ﬂy (Dobson 1974a, Hodge & Arthur 1997). Generally, adults of sandy-beach kelp ﬂies are insigniﬁcant consumers of kelp because their activity is mainly limited to eating exuded substances and laying eggs (Grifﬁths & Stenton-Dozey 1981). However, the larvae, generally found associated with decaying seaweed, contribute greatly to the breakdown of kelp tissue as a result of their feeding and tunnelling activity (Stenton-Dozey & Grifﬁths 1980, Inglis 1989, Chown 1996). Using the consumption rates and larval densities of two dipteran species (Paractora trichosterna and Antrops truncipennis) Chown (1996) estimated kelp consumption to be 714–870 g dry mass kelp m2 over his 7-wk period of study. For another dipteran species, Fucellia capensis, a value of 0.25 mg dry kelp mg1 wet larva day1 was calculated (Stenton-Dozey & Grifﬁths 1980). The bacterial populations of seaweed may be consumed by Coelopa larvae. It has been suggested that either some constituent of bacterial cells is required or that there is a need for a chemical of algal origin which is released only through bacterial action. Experimental work indicated that Coelopa larvae were able to grow on bacteria not associated with marine algae, supporting the hypothesis that the bacteria themselves were being used as a supply of nutritive energy (Cullen et al. 1987). Early invaders of wracks are followed by other insect species, mainly Coleoptera, that colonise the beds as these dry out (Moore & Legner 1973, Grifﬁths & Stenton-Dozey 1981, Colombini et al. 2000). However, as for isopod species, beetles contribute only about 10% of the total number of animals (Stenton-Dozey & Grifﬁths 1983). Herbivorous coleopterans, such as Tenebrionidae, Hydrophilidae, Curculionidae and Scarabaeidae, can be found feeding on individual kelp fronds but are considered to be of minor signiﬁcance as consumers of kelp. In fact, only 3.5% of the kelp deposited on the beach is consumed by herbivorous Coleoptera, whereas amphipods and kelp-ﬂy larvae consume 52.7% and 14.7%, respectively (Grifﬁths et al. 1983). Carnivorous beetles, belonging to the Staphylinidae, Histeridae and Carabidae families, form a large component of the wrack-bed fauna and tend to prey on larvae of dipterans and other insects. Chevin (1998) has listed a number of predatory species that occurred in wracks along the French Atlantic coast. Colombini et al. (2000) reported that the histerid, Halacritus algarum, is a “late” coloniser of wrack because of its tendency to prey on insects in decaying plant material. On the other hand the staphylinid Caﬁus ragazii, a fairly “early” wrack invader, strictly followed the changes in zonation of the wracks in relation to the semi-lunar phase, indicating a strict predator–prey relationship. Other species, such as the ptilid, Actinopteryx fucicola, remained in the wracks only until the moisture conditions became unsuitable for saprophytic fungi, the spores of which constitute its principal food. The authors concluded that the successional colonisation of wracks was strongly inﬂuenced by the species’ physiological and trophic needs and their appearance or 128

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

disappearance in the wracks was due to the progressive microclimatic changes of the deposits in relation to their varying position on the beach.

Species succession in carrion Another discrete and ephemeral system is represented by carrion stranded on the shore. This, like macrophyte wracks and other resources limited in time (e.g. dung, Valiela 1974, Koskela & Hanski 1977), is good material for the study of arthropod invasion, utilisation and succession. Carrion communities are also excellent for elucidating time-dependent patterns. Carcass microhabitat is small, has clear boundaries and frequently less than 30 necrophilous species are found on small carcasses. The carrion arthropod community develops primarily as a continuum of gradual change (Schoenly & Reid 1987). The major temporal features of carrion arthropod development include: a rapid invasion by feeding dipterans and ants, a period of maximum arthropod diversity achieved when carcass tissues are most attractive to consumers and a period of monotonic decline in arthropod richness as the carcass resource becomes depleted (Schoenly 1983, Schoenly & Reid 1983, 1987). These ﬁndings are based on studies of carcass of other environments (namely deserts) but many similarities can be found with beach-cast carrion. In xeric habitats carcasses are subjected to rapid microhabitat changes (desiccation) and sand covering. Thus the utilisation of the resource by each species is limited in time. Furthermore, persistence of carcasses on beaches may vary widely from beach to beach and from season to season depending on beach aspect, orientation, exposure and substratum, nearshore currents, wave intensity, tidal ranges, weather and on the density and activity of scavengers (Bodkin & Jameson 1991). A facilitation mechanism on behalf of the early invaders (ﬂies, ants) has been suggested by Connell & Slatyer (1977) and appears to ﬁt well to species succession in carrion. Dipterans and ants are viewed as the most efﬁcient colonists because of their good dispersal abilities and high densities (McKinnerney 1978). Through their combined tunnelling and feeding actions they facilitate the access of necrophagous taxa, such as trogid and dermestid beetles, to internal tissues. Other taxa, such as burying beetles (Silphidae: Nicrophorus spp.), arrive later on larger carcasses for feeding but use smaller and secure carcasses for interment and reproduction (Wilson & Knollenberg 1984). In a study of arthropod succession on a leatherback turtle (Dermochelys coriacea) stranded along a beach of French Guiana (Fretey & Babin 1998) a facilitation mechanism was achieved by vertebrate scavengers (black vultures, Coragyps atratus) that fed on the fresh carcass some hours after stranding. Torn ﬂesh attracted dipterans (calliphorids) that soon started oviposition. On the following day, coleopteran species (Scarabaeidae, Carabidae) were observed, while during the night Blattodea and other nocturnal coleopterans, mainly tenebrionids, exploited the carcass. By the third day dipteran larvae emerged from the holes and attracted staphylinids and histerids. Species richness peaked on the ﬁfth day, with all the above orders contemporaneously present on the turtle. By the seventh day the carcass was picked clean, although Diptera and Coleoptera were still the main groups represented on the carcass. Carabids, staphylinids and histerids were mainly predator species, hence their arrival occurred after development of their prey (ﬂy larvae). Scarabaeids were coprophagous or detritivorous and contributed to the decomposition of the carcass, while tenebrionids were sometimes saprophagous, mycetophagous or predators according to the species. Necrophagous arthropods were selectively attracted by the odour released 129

I. COLOMBINI & L. CHELAZZI

during the process of decay and, when the changing conditions in the substratum became unfavourable, the earlier colonists were replaced by other species. Each member of the insect family had a speciﬁc role in the decomposition of carrion, intervening according to its feeding and reproductive requirements and contributing to nutrient recycling. In a similar study on species succession on red-eared turtle (Pseudemys scripta elegans) carrion in a wooded area in Massachusetts (Abell et al. 1982), an accelerated process of breakdown and decay in the presence of saprophagous arthropods was observed when this was compared with arthropod-free carrion. Calliphorid ﬂies were again the dominant species of the carrion community whereas the lack of large numbers of Formicidae and Scarabaeidae was related to the presence of the shell, which induced higher water retention levels and temperature. There was also a clear succession between sarcophagid, calliphorid, syrphid, drosophilid and muscid larvae while the Coleoptera species showed a greater degree of coexistence and temporal overlap. The importance of insects as decomposers was clearly pointed out in other studies on beach carrion (Lord & Burger 1984a,b). Studying the decomposition of gull (Larus argentatus and L. marinus) and harbour seal carcasses (Phoca vitulina) the workers emphasised that marine invertebrates (e.g. crabs, amphipods) and scavenging birds had a negligible impact upon decay and that few arthropod species dominated the process. On harbour seal carcasses (Lord & Burger 1984b) blow-ﬂy larvae and dermestid beetles were the primary carrion consumers. The former consumed the internal viscera and the soft tissues within approximately 14 days, whereas the latter ate the remaining tissues within 20 days. The importance of dermestids in the decomposition process was also reported for other carcass types in different environments (McKinnerney 1978). Spiders, staphylinid and histerid beetles were the major predators of the necrophagous species. However, several beetle taxa, such as Silphidae, which often occur as carrion inhabitants, were not found on seal carcasses and their absence was related to the harsh marine conditions encountered in the rocky supratidal zone. In less extreme environmental conditions a greater number of predatory insects can be supported, as was reported for gull carcasses in vegetated habitats (Lord & Burger 1984a). In this case, moderate climatic conditions increased arthropod colonisation and survival, resulting in an increase of the arthropod community. In particular, consumption and decomposition rates were reduced because there were more predator species and fewer carrion consumers in this habitat. Observations on a dolphin (Stenella sp.) carcass stranded on a beach of Southern Italy again indicated the importance of insect species in the consumption of the animal (pers. obs.). The dolphin had been stranded about 20 days before the arthropods were analysed and half of its body, with all internal soft parts, had already been consumed (i.e. the bones were half picked clean). Dipteran larvae were most abundant and represented the dominant taxa. These larvae attracted coleopteran predators, mainly staphylinids and histerids, which constituted 49% of the total taxa present on the carcass (excluding dipterans from counts). Fewer amphipod and isopod species were present (3% and 16%, respectively) as would be expected from the advanced state of decay. The remaining 32% of the taxa was represented by scavenging tenebrionids (Phaleria spp.). Not only predatory beetles but also shorebirds are opportunistic feeders on ﬂies, larvae and other insects associated with carrion. Sanderlings (Calidris alba) were reported to forage on ﬂy species associated with carcasses of laughing gull (Larus atricilla), northern gannet (Morus bassanus), common loon (Gavia immer) and striped burrﬁsh (Chilomycterus

130

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

schoepﬁ) and on one occasion birds were seen protecting a carcass from other avian predators such as the black-bellied plover (Pluvialis squatarola) (Grant 1997). Both macrophyte wracks and carrion must be considered as a “processing chain”, that is, as units of a resource passing through a sequence of condition changes over time, with consumers specialising in a speciﬁc resource in each condition and inﬂuencing the rate at which the resource is transformed (Heard 1994). In this process, a range of interspeciﬁc interactions can be found and can vary from amensal to commensal depending on a number of factors. These factors include temporal scale, species composition and relative processing rates and efﬁciencies (Heard 1994, Hodge & Arthur 1996).

Intra- and interspeciﬁc interactions in wracks and feeding preferences Macrofaunal species (amphipods, isopods and dipteran larvae) are the major primary consumers of beach-wrack macrophytes and there are many studies on individual species within the macrophyte community, and their interactions in the wracks (Craig 1970, 1973, Dobson 1974a,b, Hayes 1974, Kensley 1974, Koop & Field 1980, 1981, Robertson & Mann 1980, Stenton-Dozey & Grifﬁths 1980, Venables 1981, Behbehani & Croker 1982, Agnew & Moore 1986, Marsden 1991a,b, Phillips & Arthur 1994, Phillips et al. 1995a,b, Chown 1996, Hodge & Arthur 1997, Persson 1999, Pennings et al. 2000). Amphipod species are usually numerically dominant in wracks. Much attention has been given to the biological cycles, population densities, daily production and respiration estimates of Orchestia and Talorchestia, species commonly found associated with macrophyte deposits (Venables 1981, Behbehani & Croker 1982, Marsden 1991a,b). Persson (1999), in studying the growth and reproduction of Orchestia gammarellus in wrack beds of beaches along the Baltic Sea, showed the inﬂuence of wrack type on the growth rates and reproductive patterns of the amphipods. Growth rate of the summer generation was signiﬁcantly higher in beds dominated by the eel-grass, Zostera marina, and this increased growth was probably due to the higher temperatures that were reached, because this type of wrack was generally drier than the algal beds. Higher temperature increased molting frequency and induced higher growth rates. In wrack beds of the alga Fucus vesiculosus, overwintering individuals had a signiﬁcantly greater mean length compared with those overwintering in eel-grass, indicating that some beneﬁts were obtained by animals living in the algae. Studies on feeding preferences indicated that Orchestia gammarellus preferred to ingest algal material rather than angiosperms (Moore & Francis 1985) perhaps because it was softer and easier to handle (Bedford & Moore 1984) or because it attained a higher density of microrganisms. In fact, it was demonstrated that neither seagrass litter nor algal material were digested by the amphipods (Lopez et al. 1977, Bedford & Moore 1984) but that the beach-hoppers used the attached micro-organisms as a source of nutrient instead. Colombini et al. (2002) showed that the differences found in the non-uniform spatial distribution of two amphipod species (Talorchestia brito and Talitrus saltator) living along a Tunisian beach were due to the patch distributions of stranded food items. Another study on the feeding preferences of three North American supralittoral crustacean species (the isopod Ligia pallasii, and the two amphipods Megalorchestia californiana and Traskorchestia traskiana) showed that the three species had similar feeding preferences and suggested that these were based on general algal traits rather than on particular adaptations of speciﬁc herbivores (Pennings et al. 2000). Furthermore, the three 131

I. COLOMBINI & L. CHELAZZI

species showed a tendency to prefer aged over fresh seaweeds. Preference for older wrack was basically explained by its reduced water content. Because rate of food intake is limited by its volume, a given volume of aged seaweed will have several times the organic and mineral content of the same volume of fresh material. This concentration of organics was strongly preferred by the talitrids, especially, when time for feeding was limited (Carefoot et al. 1998). Also, the reduced quantity of defensive compounds, such as phenolics and terpenes, which leach and break down once the seaweeds begin to desiccate, probably render older wracks more palatable than fresher ones, as was demonstrated for senescent angiosperm leaves (Harrison 1982, Valiela et al. 1979). On the contrary, two other amphipod species Echinogammarus pirloti and E. obtusatus showed neither a preference for old versus fresh seaweed nor for substrata containing high or low loadings of bacteria. However, both preferred the “softest” food material, usually decaying Laminaria digitata, and Agnew & Moore (1986) suggested that the role of micro-epiphytes (bacteria, diatoms) in the diets of these amphipods may be of minor importance. Instead, the deﬁcit of nitrogen may be supplemented by feeding on meiofaunal crustaceans, dead conspeciﬁcs, moulted exoskeletons or by coprophagy. In Orchestia scutigerula, Moore et al. (1995) noticed that this beachhopper fed on a large spectrum of materials, although items like diatoms crustacean and mite fragments were in the minority. Tussock grass and green algal debris were highly consumed, although there was evidence that their attractiveness changed over time in favour of algae. This change was related to the conditions of the deposits, where different materials were available in different proportions, depending on their location on the beach and on their time of stranding. In fact, there was a decline of algal-derived fragments in the gut contents of O. scutigerula with landward shifts of the debris. Preference experiments indicated that these talitrids preferred brown algae to certain red and green algae, which are known to be protected chemically (Paul et al. 1987, Hay & Fenical 1988). It was also evident that O. scutigerula fed on vegetable material of the softest texture, avoiding harder materials like the vascular tissue of tussock grass. Observations on the feeding behaviour of the surf-zone inhabiting amphipod Allorchestes compressa also indicated that these amphipods fed on small particles of both seaweeds and seagrass but deliberately avoided red algae and preferred the decomposing tissues of the laminarian Ecklonia radiata taken from beached stacks of macrophytes (Robertson & Lucas 1983). Studies on individual isopod species in relation to macrophyte wrack deposits have been carried out by a number of workers (Hayes 1974, Kensley 1974, Koop & Field 1980, 1981, Robertson & Mann 1980). Koop & Field (1980) came to the important conclusion that the life histories of species inhabiting this protected microhabitat were governed by biological factors such as food availability more than by physical ones, as might be expected in supralittoral species. They showed an increased growth and reproductive development of Ligia dilatata during winter and related it to the clear seasonal pattern of kelp debris availability. Following periods of high food availability, in early spring ovigerous females and new cohorts appeared, showing that the surplus of assimilated energy was used for reproduction. Furthermore, ﬁeld observations indicated that L. dilatata ate only debris of the kelp Ecklonia maxima and thus depended almost entirely on cast material as a food source. In another study on the energy ﬂow through a population of the same isopod species, fast consumption and egestion rates, coupled with low assimilation and growth efﬁciencies suggested Ligia dilatata may be a major energy transformer (Koop & Field 1981). Of its total energy consumed, only 1.3% was channelled into production whereas 72% was extruded as faecal pellets. These, together with kelp fragments and dissolved organic matter released 132

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

mechanically by trituration, are then suspended and returned to adjacent aquatic ecosystems. Beach-cast material can also inﬂuence the spatial distribution patterns of isopod species on beaches. Hayes (1977), studying a Californian population of Tylos punctatus, found more isopods burrowed beneath or near piles of stranded kelp than in regions where kelp mounds were absent. On a Mexican beach, Simmons (in Brown & Odendaal 1994) found that the distribution of active T. punctatus could be related to the distribution of its primary food source Zostera marina and that the degree of wrack moisture was an important element in attracting the isopods. Kensley (1974) came to a similar conclusion for South African populations of Tylos granulatus and T. capensis, in which the distribution patterns of both burrowing and foraging individuals were studied in relation to the distribution of kelp deposition on the beach. Food preference experiments have shown that isopods were attracted to the food source by olfactory means, and food items with stronger odours were preferred. Generally, high isopod numbers corresponded with the presence of food deposits. As in other isopod species, juveniles and adults, preferred drier brown algae to fresh, and both age groups preferred animal matter when given the choice between animal and macrophyte debris (Pennings et al. 2000). In two sympatric scavenger tenebrionid species of the genus Phaleria (P. provincialis and P. bimaculata) the bacteria isolated from the gut contents were studied by physiological and molecular analysis (Barberio et al. 2001). The study showed a higher biodiversity of bacteria in P. bimaculata than in P. provincialis indicating a difference in foraging areas and, indirectly, in food preferences. Studies on dipteran species associated with kelp material washed ashore are abundant in the literature and can be reviewed according to the geographical location of the species. In the sub-Antarctic, Marion Island and South Georgia have been studied. At Marion Island (Crafford 1984, Crafford & Scholtz 1987, Klok & Chown 2001), attention has been focused on Paractora dreuxi. This species contributed to the degradation of stranded kelp, mainly Durvillaea antarctica, by consuming the stranded material directly or by tunnelling through the decaying fronds and thus increasing microbial decomposition. Recently, distinct ontogenetic differences in thermal tolerance and water balance have been elucidated for this species and related to the effects of global climatic change, which consequently produced pronounced microclimatic changes in the wracks (Klok & Chown 2001). Crafford & Scholtz (1987) estimated that the larvae of Paractora dreuxi were responsible for 35% loss of kelp dry mass during kelp degradation and that this species was important in the recycling of nutrients in the terrestrial ecosystem of the island. At Husvik Harbour, South Georgia, Pugh & MacAlister (1994) found that two kelp ﬂies, P. trichosterna and Antrops truncipennis, were obligate feeders of stranded seaweeds. Chown (1996) showed that these two ﬂies contributed signiﬁcantly to kelp degradation, with Paractora trichosterna responsible for 12% of kelp dry biomass in beds protected from vertebrate trampling and 20% in the exposed ones. In comparison, Antrops truncipennis was responsible for an additional 3% loss in the exposed and 8% in the protected beds. Paractora trichosterna was macropterous and not brachypterous like P. dreuxi at Marion Island and this difference in wing development was related to the different strategies adopted by the two species for the utilisation of stranded seaweeds. At Husvik, seaweeds were deposited as strings or small patches and thus favoured the retention of a ﬂight capacity. At Marion Island, where there was the tendency of seaweeds to form large stranded beds, brachyptery was favoured. The differences found in the biomass of the two species indicated that they were utilising kelp in different ways, with Antrops truncipennis favouring more sheltered 133

I. COLOMBINI & L. CHELAZZI

areas where it deposits its eggs. The author concluded that the dipteran assemblage associated with wrack was important in wrack decomposition and contributed signiﬁcantly to nutrient recycling. In the British Isles, much work has been carried out on the seaweed ﬂies of the genus Coelopa, which are considered to be a major component of the wrack beds (Day et al. 1980, 1982, 1983, 1987, Butlin et al. 1982, 1984, Butlin & Day 1985, 1989, Cullen et al. 1987, Phillips & Arthur 1994, Phillips et al. 1995a,b, Leggett et al. 1996, Hodge & Arthur 1997). Earlier works on shore-ﬂies were mainly based on systematic and life-history studies, particularly on two closely related species, Coelopa frigida and C. pilipes (Egglishaw 1960, Dobson 1974a,b, 1976, Simpson 1976). Dobson (1974a,b) analysed the breeding cycles of the two species in relation to the duration and type of wrack and studied the variations in population size, in species composition and the effects of the destruction of the wrack-beds on the two Coelopa species. In a study of the larvae of some dipteran species associated with wrack on Californian beaches, Kompfrner (1974) emphasised the different strategies adopted by the ﬂies in the exploitation of the available food and habitat resource. One species (C. vanduzeei) found exclusively in the lower, wetted beach wracks, developed quickly, whereas others (Fucellia costalis, F. ruﬁtibia), occurring in the mid- and upperbeach levels, were able to use the wrack for longer periods of time and consequently had a longer pupal stage. The genetics of seaweed ﬂies have also been studied in some detail. It has been demonstrated that all northern European populations of Coelopa frigida are polymorphic for a large inversion on chromosome I (Butlin et al. 1982, Day et al. 1982, 1983) and that polymorphism is maintained by selection. Laboratory experiments have indicated that at least three types of selection may occur: heterokaryotypic advantage in egg-to-adult viability (Butlin et al. 1984), variation in development time (Day et al. 1980) and sexual selection (Butlin et al. 1982, Day et al. 1987). Butlin & Day (1989) studied the connections between environmental variables (i.e. composition, surface area, depth, extent of decomposition, temperature, degree of moisture of seaweed beds) and inversion frequencies, while Leggett et al. (1996) assessed the genetic effects of competition in C. frigida and C. pilipes. Leggett et al. concluded that both intra- and interspeciﬁc competition can be ﬁerce and appear to be the force maintaining, rather than eliminating, genetic variation. In a more ecological study, Phillips et al. (1995b) showed that these two Coelopa species frequently coexisted, despite competition and that the temperature of the wrack had an important role in the distributions of the larvae. C. frigida larvae preferred cooler parts of the bed whereas C. pilipes larvae preferred the warmer ones. These distributions were shown to be caused by the behaviour of the larvae themselves rather than a selection made by ovipositing females. The different micro-distributions within beds caused competitive abilities to be frequency-dependent and thus permitted coexistence. In another study (Phillips et al. 1995a) it was shown that the different egg distribution of the two ﬂy species depended on a different preference for seaweed species, rather than on microhabitat factors such as temperature and humidity, with C. pilipes showing stronger preferences for Fucus than did Coelopa frigida. Once established in the wrack, larvae were able to move around according to their microhabitat preferences. The distributions and interspeciﬁc interactions of the Coelopa species and of the other fauna found within the wrack beds were analysed by Phillips & Arthur (1994). They found that all of the animals were non-randomly distributed in the wracks. Coelopa larvae aggregate in warmer and deeper areas, in agreement with the fact that breeding sites occurred in parts of decaying weed where temperature was raised by 134

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

anaerobic decay (Egglishaw 1960, Dobson 1974a). This contrasts with amphipod species, which presented a negative correlation with temperature and dominated the community in cool wrack beds but were not signiﬁcantly inﬂuenced by depth within the bed. Enchytraeids were different again and were dependent on depth of the wrack but not on the temperature. Finally, the distribution of the commonest predator species (the staphylinid Caﬁus xantholoma) was correlated with the distribution of its prey (dipteran larvae, enchytraeids) more strongly than with the abiotic factors. It was concluded that the different specialisations would tend to reduce the intensity of any competitive interactions between the Coelopa species and the other groups. In a more detailed study on the interactions between the three commonest seaweed species, C. frigida, C. pilipes and Thoracochaeta zostera (Hodge & Arthur 1997) it was shown that there were no systematic differences in the colonisation patterns of the wrack with all three species arriving soon after wrack deposition. However, a form of priority effect might have been produced through the oviposition behaviour of females, which differed among species. C. frigida tended to lay eggs in batches while the other two species laid eggs singly. In this case the competitive stage of the larvae of C. frigida could occur earlier and at higher densities on the wracks compared with those of other ﬂies. The other two species would then be at a disadvantage and would suffer because of their interactions with the larger and more competitive C. frigida larvae. The authors found that all pairwise interactions were extremely asymmetric with C. frigida being the strongest competitor species and Thoracochaeta zostera the weakest. Furthermore, a strong facilitatory effect of Coelopa pilipes on C. frigida was observed on the latter’s emergent population size, which was more than double when reared on chopped seaweed. The minced seaweed was used to simulate the fragmentation of seaweed by the ﬂy larvae in the ﬁeld. This facilitatory effect was explained by the resource modiﬁcation theory, according to which the larvae of C. pilipes changed the nature of the resource, perhaps physically, chemically or microbially, to favour feeding of C. frigida larvae. This facilitation, combined with the inhibition of C. pilipes by C. frigida, produced a contramensal interaction common in other interspeciﬁc interactions (Hodge & Arthur 1996).

Shorebirds and other terrestrial animals linked to wracks Beach-cast macrophytes supporting prey resources are commonly exploited by a number of shorebirds. In Australia, the ruddy turnstone (Arenaria interpres), the hooded plover (Charadrius rubicollis) and the silver gull (Larus novaehollandiae) were all found closely associated with beach-wrack accumulations (Kirkman & Kendrick 1997). The hooded plover was most abundant where there were large amounts of macroalgae. Plovers nested close to macroalgae, making small depressions in the sand, and fed on crustaceans, molluscs, insects, and polychaetes associated with the wracks (Schulz & Bamford 1987). In a few other studies, the distribution and abundance of shorebirds and other avian predators were related to densities of standing stocks of stranded algae. On beaches of the Skeleton Coast, Namibia, Tarr & Tarr (1987) reported higher shorebird densities in areas with higher densities of kelp stranded on the shores, and it has been shown that beaches with high kelp inputs supported 75% higher biomass of potential prey organisms than did beaches lacking stranded kelp (Tarr et al. 1985). The potential importance of prey associated with macrophyte wracks was also stressed by Grifﬁths et al. (1983) for beaches in the western Cape of 135

I. COLOMBINI & L. CHELAZZI

South Africa. Polis & Hurd (1996b) estimated that the overall arthropod abundance in areas receiving input from the ocean was 2.5 to 550 times greater than mainland sites away from such input. Bradley & Bradley (1993), comparing censuses of over-wintering shorebirds in southern California before (1969–73) and after kelp recovery (1984–6), found that ﬁve shorebird species (spotted sandpiper, Actitis macularia, wandering tattler, Heteroscelus incanus, whimbrel, Numenius phaeopus, black turnstone, Arenaria melanocephala and ruddy turnstone, Arenaria interpres) out of nine increased dramatically in abundance during the second census period. The authors suggested that the increase was caused by the recovery of kelp beds off the coast, which in turn increased prey availability to foraging shorebirds. This was indirectly conﬁrmed by the insigniﬁcant increases found for other species (black-bellied plovers, Pluvialis squatarola) which did not utilise algal windthrow. On two beaches in Baja California, Lopez-Uriate et al. (1997) suggested a relationship between migrating shorebirds and wrack-associated prey. In particular, at Punta Cabra, the semipalmated plover (Charadrius semipalmatus), which generally preferred amphipods over other food items, beneﬁted from the extremely abundant kelp wrack washed ashore. In this locality, a similar relationship was also found for the spotted sandpiper (Actitis macularia). The high abundance of snowy plovers (Charadrius alexandrinus), which generally prefer dune-backed beaches, was attributed to the distribution of the staphylinid Bledius sp. found in abundance on this beach (Grover & Knopf 1982). In a recent study on southern Californian shores, Dugan & Hubbard (in press) found positive correlations between the wintering abundance of two plover species (black-bellied plovers, snowy plovers) and the standing crop of wrack. The positive correlations found between the abundance of the two species and the abundance and species richness of the wrack-associated invertebrates suggested that this relationship was due to the increase of prey availability in the wrack subsidies. The authors stressed that human disturbance, such as beach grooming, has a heavy impact on the composition and trophic structure of the macrofauna community. This grooming may signiﬁcantly reduce the prey resources available to shorebirds and seriously threaten endangered species. Not only beach-cast material but also macrophytes present in the intertidal zone were shown to be important for some migratory birds (Percival & Evans 1997). A direct relationship between biomass density of macrophyte food supplies, such as eel-grass species (Zostera noltii and Z. angustifolia) and green algae (Enteromorpha spp.), and the distribution of brent geese (Branta bernicla) was assessed. It was demonstrated that food intake rate rapidly declined as food biomass density decreased through the season. Consequently, the birds responded by additional foraging during the night, and when this was no longer possible the birds moved out of the site. The lower threshold of vegetation biomass densities was about 5 g m2 and was determined by the birds’ energy requirements. Apart from shorebirds, other secondary consumers connected to macrophyte subsidies are represented by spiders, scorpions, lizards, rodents and coyotes. In the states of Baja California Norte and Sonora (Mexico) these consumers were from 3 to 24 times more abundant on the coast and on small islands compared with inland areas and larger islands (Polis & Hurd 1995, 1996a,b, Rose & Polis 1998). In particular, in the Baja system, terrestrial consumers (spiders, scorpions, and lizards) ate prey, primarily arthropod species derived from the marine food web, which made up 95–99% of their diet. These conduits of marine productivity made the abundance of potential prey signiﬁcantly higher in the supralittoral. Consequently, the population of spiders found along the coast was six times more abundant than that of inland areas (Polis & Hurd 1995). This relationship was also conﬁrmed by stable 136

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

carbon and nitrogen isotope analysis, which showed that the spiders’ diets were signiﬁcantly more marine based than those of the inland populations (Anderson & Polis 1998). Scorpions likewise beneﬁted from allochthonous ﬂow from the sea. Along the shores of the Gulf of California, the supralittoral scorpion Vaejovis littoralis reached densities of 8–12 ind. m2, whereas in inland areas the density of all species combined reached a value of only 0.2–1.0 ind. m2 (Due & Polis 1985). For lizard populations, estimates were four times more abundant in the supralittoral zone than inland areas (Polis & Hurd 1996b). Further up the food chain coastal coyote populations, subsidised by the ﬂow of abundant and diverse resources from the sea, also showed an increase in density, with populations 2.4–13.7 times more dense than in adjacent inland areas (Rose & Polis 1998). The coastal coyotes ate not only many living marine species (arthropods such as crustaceans and larvae of insects that eat algal drift, algae and molluscs) but also many terrestrial creatures that exploited marine input (arthropods, lizards, land birds and coastal rodents). This diverse and conspicuous marine input, together with the in situ terrestrial food supply had the effect of increasing the dietary spectrum and food intake of coastal coyotes compared with those of inland populations.

Offshore consumers of wracks Filter feeders, grazing gastropods and ﬁshes Once deposited on shores, macrophyte wracks are fragmentised and decomposed. These wracks form an important food resource consisting of fragments of seagrasses and algae, bacteria, meiofauna and beach macrofauna which, during storms, can be washed back to the sea and form the basis for primary production and food chains in nutrient-poor coastal waters (Robertson & Hansen 1982). Filter-feeders, grazing gastropods and ﬁshes are the major consumers of this food resource. It has been emphasised that particulate organic material from kelp detritus greatly increased the growth rates of offshore benthic and pelagic ﬁlter feeders (Duggins et al. 1989). The use of macrophyte-derived detritus by higher trophic levels, and its role in the secondary production has also been stressed by a number of other workers (Mann 1988, Duggins et al. 1989, Bustamante et al. 1995). Bustamante & Branch (1996) related the high biomass of ﬁlter-feeders on exposed shores to a higher concentration of particulate food and a more rapid replenishment than occurs on sheltered shores. Soares et al. (1997), studying populations of Donax serra, estimated the degree to which annual carbon requirements of this species were met by food originating from kelp. Pathways of this food resource were considered in both nearshore waters and after kelp stranding on the shore. In these habitats food became available to D. serra as particulate organic matter from frond erosion, drifting and stranded fragmentised kelp, pelagic bacteria in nearshore waters and intertidal bacteria and faeces from primary and secondary consumers (Newell et al. 1982, Grifﬁths et al. 1983). It was estimated that kelp detritus, bacteria and kelp consumers’ faeces available in the water column surpassed several times the carbon and nitrogen requirements for both intertidal and subtidal clam populations and suggested that beach clams may have beneﬁted from these energy subsidies (Soares et al. 1997). Other works in

137

I. COLOMBINI & L. CHELAZZI

Western Australia (Wells & Keesing 1989, Schiebling 1994) report that species such as Roe’s abalone (Haliotis roei) feed exclusively on drifting fragments of macroalgae. Lenanton et al. (1982) studied the diversity and abundance of the mobile epifauna associated with nearshore, detached macrophyte accumulations in Western Australia. The amphipod Allorchestes compressa, inhabiting wrack deposits and feeding on Ecklonia radiata (Robertson & Lucas 1983), could be returned to the surf zone during rough weather and constituted the major prey item of the juveniles of four ﬁsh species (the yellow-eyed mullet Aldrichetta forsteri, the cobbler Cnidoglanis macrocephalus, the school whiting Sillago bassensis and the Australian herring Arripis georgianus). The data showed that the arrival of juvenile ﬁshes on the open coast in late winter corresponded with the period of greatest deposition of detached macrophytes in the surf zone. The volumes of detached vegetation and associated amphipods were sufﬁcient to support the ﬁshes during summer months. Nearshore accumulations of detached macrophytes could provide an alternative feeding habitat for these benthic feeders, of which one, Aldrichetta forsteri, is normally restricted to shallow estuaries or sheltered embayments (Robertson 1980). The abundance and species composition of the surf-zone ﬁsh community also correlated positively with the quantity of detached macrophytes and with the prey densities (Robertson & Lucas 1983, Robertson & Lenanton 1984).

Effects of physical presence of stranded wrack on plants and animals Beneﬁcial effects The previous paragraphs have reviewed the effects of the presence of organic beach-cast material on both invertebrates and vertebrates of beach and nearshore communities. It has been shown that, generally, these allochthonous subsidies have beneﬁcial effects on the different communities, increasing species richness and population densities. Apart from providing food, the most obvious physical effect of wrack deposits is to provide shelter, and a suitable microhabitat for a number of different taxonomic groups of animals (Lavoie 1985, Colombini et al. 2000) and to act as important links between habitats. In fact, organic material stranded on beaches forms the basic element of food webs through which energy is transformed from one component to the other (Brown & McLachlan 1990, Polis & Hurd 1996a,b, Polis et al. 1997). Another beneﬁcial effect consists of species dispersion through ﬂoating beds, a phenomenon particularly common for coleopteran insects that can survive for several days at sea (Chevin 1998). Also, the widespread distribution of the coconut crab (Birgus latro) throughout the western tropical Paciﬁc region up to the islands of the Indian Ocean can be explained by taking into account the dispersal of tropical seeds. In fact, Harries (1983) postulated that the tiny post-larval stage (glaucothoe) was spent in the moist husk of ﬂoating coconuts and that the ancestors of today’s coconut crabs may have migrated on ﬂoating coconuts to remote islands and atolls of the South Paciﬁc. Vertebrate species may also have used drift material to reach distant islands. This seems to be the case for some reptiles of the Galapagos Islands, such as the marine iguanas (Amblyrhynchus cristatus) which many thousands of 138

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

years ago may have ridden large rafts of tangled vegetation broken loose from riverbeds of Ecuador’s Guayas River and drifted by the swift currents to the sea (De Roy Moore 1980). This theory is supported by the fact that only certain animals (reptiles and not amphibians) are present in the Galapogos, reﬂecting their ability to accomplish and survive such travels. Recently, however, phylogenetic analyses and age estimations have suggested an alternative hypothesis based on in situ speciation. According to this hypothesis the land and marine iguana inhabited the former, now sunken, islands of the Galapagos but became separated within the archipelago (Rassmann 1997). Plants also can beneﬁt from the presence of stranded seaweeds on shores through the amelioration of soil characteristics (Hesp 1991, Haslam & Hopkins 1996). Foredune plants can receive nutrients, including seaweeds and carrion, from sand movements and swash deposition or through algal foam blown onshore. These species adapt to the different input of nutrients using different strategies for the absorption of mineral nutrients, i.e. via salt spray or soil (Hesp 1991). Haslam & Hopkins (1996) investigated the effects of the brown seaweed Laminaria digitata on the pore volume and size distribution, aggregate stability, soil microbial biomass and biological activity (respiration and N mineralisation) of sandy soil. Pore volume and total water holding capacity was increased by kelp addition. Aggregate stability, together with soil microbial biomass content and soil respiration rate were also signiﬁcantly greater. The rate of potential N mineralisation also increased by kelp amendment. In another coastal environment (salt marsh) a similar beneﬁcial effect of wrack deposition to plants was reported by Pennings & Richards (1998). In a southwestern Atlantic salt marsh the plant biomass of Batis maritima was several times higher in elevated zones compared with lower ones. This increase was associated with the presence of wrack in the elevated zones, which ameliorated the soil characteristics by lowering salinities and organic content and producing better percolation rates. Thus wrack deposition had an important role in reinforcing pre-existing ﬂood and salinity gradients caused by terrestrial runoff and elevation.

Detrimental effects There are a few cases in which kelp stranded on shores has a detrimental effect on invertebrates. McLachlan (1985), for example, stated that large accumulations of beach macrophytes could limit oxygen exchange and the presence of the meiofauna could be seriously affected. Another case was reported by Soares et al. (1996) for clam communities. These authors, surveying 12 beaches in South Africa, indicated that biomass and density of adults of the wedge clam, Donax serra, were signiﬁcantly higher on beaches with lower stranded kelp cover. Adults were centred in different zones according to different kelp cover. They were found in the low intertidal to subtidal where kelp cover was high and also in the midintertidal where no kelp was found. It was supposed that stranding of kelp produced a physical disturbance of the clams and this may have affected both feeding and burrowing activities. Feeding time could be decreased by the siphons making contact with the kelp and withdrawing into the shell and by the presence of a shadow that could stimulate this same reaction in response to presumed predators. Stranding of kelp also disrupts the swash and decreases water ﬂow through the sand (McLachlan et al. 1985) and thereby affects food intake. Similarly, the surging back and forth of kelp would mechanically disturb slow burrowing adult clams and their reburrowing times would be increased. Consequently, they 139

I. COLOMBINI & L. CHELAZZI

would be dislodged downshore while smaller juveniles, which burrow faster, would be found in the uppershore, which explains the differences in the spatial distribution found for adults and juveniles (Donn 1990, Soares et al. 1996). The disturbance would be magniﬁed when coupled with other physical factors such as low temperatures and small waves on pocket beaches with steep slopes and coarse sands. Predation of dislodged animals by gulls could explain the clam-free 14–20 m above the swash observed in two of the 12 beaches studied by Soares et al. (1996). In other coastal environments disturbance of plant communities by wrack deposition has been reported (Valiela & Reitsma 1995, Brewer et al. 1998, Fischer et al. 2000). Burial by ﬂoating plant debris is one of the main factors affecting saltmarsh plant communities because it can kill underlying vegetation and create mosaic patches of different stages of secondary succession (Bertness & Ellison 1987, Valiela & Reitsma 1995). Removal of the deposited mats by the tides leaves hypersaline bare patches and creates opportunities for the establishment of other salt-tolerant species and determines different community patterns (Bertness & Ellison 1987, Brewer et al. 1998).

Geomorphological importance of wrack Together with other factors (Hellemaa 1998), accumulations of stranded macrophytes along beaches have the important role of inducing dune formation. Wracks are particularly important on exposed beaches where they stabilise the foreshore by enhancing the organic and moisture contents, allowing pioneer plants to establish (Llewellyn & Shackley 1996). In Mauritania, Hemminga & Nieuwenhuize (1990) reported massive accumulations of seagrass litter on the shores between Iouik and Ten Alloul of the Banc d’Arguin and associated it with local dune formation. The proﬁles of the beach on the seaward side showed a stepwise formation with alternating layers of sand and seagrass litter. These were covered by sand originating from the Sahara during strong winds. It was proposed that the interacting process of massive stranding of seagrass litter and the seaward transport of desert sediment particles resulted in a dynamic process of dune formation, with the tendency of this process to shift in a seaward direction. The higher parts of the dune are then stabilised by the colonisation of halophyte pioneer plants (e.g. beach vines). A similar process has been suggested for the stable dunes of Cervantes in western Australia (Hesp 1984). In this case the in situ porous sand that quickly dried out and the winds that carried the sand towards land were the factors implicated in the process. The aeolian capacity for sand transport weakened as it encountered an obstacle so that sand accumulated behind and around it, forming an embryo dune. In Kingston, South Australia, Kirkman & Kendrick (1997) indicated that Posidonia remains at the base of the coastal dunes could be implicated in dune formation through trapping and binding of the drifting sand. Prince et al. (1968), investigating the effect of artiﬁcial seaweed on beach erosion, concluded that artiﬁcial seaweed can build-up beaches by promoting an onshore transport of material. In the past decade the effect of kelp harvesting on dune erosion has been investigated along the Norwegian coast (Berg & Munkejord 1991, Løvås & Tørum 2001). Along the Jæren coastline, increased dune erosion was observed following an increase of kelp harvesting. It was shown that forests of Laminaria hyperborea had a damping effect on wave action. This attenuation was reduced when kelp was harvested and dune erosion ensued. 140

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

Furthermore, it was demonstrated that kelp modiﬁed the water velocity proﬁle. In a region above the canopy layer, the time-averaged water velocity was in a shoreward direction whereas the undertow was conﬁned to a region higher up in the water column. Hence it was concluded that sea vegetation, and especially L. hyperborea, dampen the waves signiﬁcantly depending on the wave length : water depth ratio and that intense harvesting may severely affect coastal erosion. Such vegetation may also contribute to the maintenance of the shoreline through root mats that stabilise the bottom and protect the sediment surface from erosion (Fonseca & Fisher 1986).

Human use of beach-cast material Harvesting of seaweeds and seagrasses In many parts of the world beach-cast wrack and subtidal macroalgae are harvested because they are considered economically important resources (Berg & Munkejord 1991, PachecoRuíz & Zertuche-González 1996, Kirkman & Kendrick 1997). In areas where wrack has accumulated in large quantities, its decomposition, production of hydrogen sulphide gas and ﬂy plagues have a negative impact on human use of beaches and are considered a serious problem for beach management (Blanche 1992). Thus, harvesting of stranded macrophytes is seen as a positive means of cleaning beaches that are important for recreational use and tourism, and for producing a variety of valuable products. Beach-cast and subtidal macrophytes are used as house insulation, garden fertiliser and soil improvers, stock and mariculture feeds, and alginates and agar are employed as additives to human food, medical, cosmetic and pharmacological products (Jolivet et al. 1991, Haslam & Hopkins 1996, Kirkman & Kendrick 1997). Kirkman & Kendrick (1997) reviewed harvesting activities in Australia and examined their impact on littoral and nearshore marine communities. They reported that only a few places in temperate Australia were harvested and that the major industry, based on collecting stranded bull kelp (Durvillaea potatorum), was on King Island in Bass Strait. The local population collected bull kelp from the swash zone as it washed up and air-dried it for 2 wk on racks supplied by the industry. Once dried, the kelp was processed by the factory by crushing into ﬁngernail and sand grain size pieces before shipment to Kelp Industries (KELCO) in Scotland. Here it was further dried and chipped before being sent to Alginate Industries, UK for making into alginates. It is believed that harvesting has had little impact on the populations of bull kelp but other factors, such as warm sea temperature and decreased frequency of storms, might have caused resource shifts. Other by-products of the King Island industry were smaller grains and dust sold for stock feeds and as soil improvers, respectively. In Australia, abalone mariculture and aquaculture industries are other important users of stranded and drifting macroalgae (Kirkman & Kendrick 1997). However, total reliance on wild stocks is not sustainable and artiﬁcial feeds and mixtures of artiﬁcial feeds and dried macroalgae are now being developed. Beach-cast seagrass leaves of the genus Posidonia were also harvested at Kingston, South Australia and used as soil improvers by locals or sold to the Japanese market. No quantitative estimates have been made of the harvest and therefore little is known of how beach removal can affect nutrient return to nearshore communities. 141

I. COLOMBINI & L. CHELAZZI

In the Gulf of California, Pacheco-Ruíz & Zertuche-González (1996) listed 55 potentially commercial seaweed species and found that the most abundant were the phaeophytes, Sargassum johnstonii, and S. sinicola, which were used as a source for alginates and fertilisers; the rhodophytes, Eucheuma uncinatum and Gigartina pectinata used as source of carageenan, and the chlorophytes Ulva lactuca and Enteromorpha spp. used in medicine as well as for human consumption. These species are annual, and the necessity for further studies to determine the sustainable harvest capacity and/or culture practices necessary for commercial exploitation has been stressed (Barilotti & Zertuche-González 1990). In reviewing the effects of marine algal extracts on the productivity of agricultural plants, Jolivet et al. (1991) listed several beneﬁcial effects when these extracts were added to the soil, notably an increased resistance of the plants to parasites, fungal attacks and to lower temperatures and humidities. Moreover, when sprayed on leaves the extracts improved germination of seeds, growth and development, enhanced mineral absorption from soil and increased quality of certain plant crops. Marine macroalgae and seagrasses have been used as soil fertiliser by coastal populations for many centuries. On the Tyrrhenian coast of Italy, particularly in the Maremma region at the beginning of the twentieth century, seagrasses of the genus Posidonia and Cymodocea were commonly collected by local farmers, dried and used as fertilisers or as cattle bedding. In Italy this practice has long been abandoned in favour of more practical commercial products. However, elsewhere in the Mediterranean beach-cast macrophytes are still exploited by local populations. Along the coast of Tunisia, (Gulf of Gabès) the collection of stranded Zostera sp. by individual farmers is still a common practice and is used as an additive to the soil of olive tree plantations (I. Colombini & L. Chelazzi, pers. obs.). Along the coast of the Indian Ocean, beach-cast macrophytes have been exploited by local populations for other uses. For example, at Danane, a beach locality to the south of Mogadishu, Somalia, local women bake their pottery using large heaps of burning seagrass Thalassodendron ciliatum. This technique is typical of the area and has been a tradition of these coastal tribes for a long time, as testiﬁed by the pottery found in many Palaeolithic sites along the coast of East Africa (pers. obs.).

Use of other stranded material The exploitation of other material of organic origin stranded on beaches by coastal populations is a worldwide practice. In Africa, large stranded logs have always been of value to coastal populations for their use in the construction of “sanbûqs” and of smaller “dahws”, typical ﬁsherman boats found along the Indian Ocean (Grottanelli 1955). In the past, ambergris represented an important resource and was actively sought along shores because it was more valuable than gold. Along the oriental coast of Africa, from the Bagiuni Islands to Madagascar, this material has been traded with Giava, Sumatra and China since AD 945 (Grottanelli 1955). At the beginning of the twentieth century ambergris was sold in northern Somalia at 10 times its weight in silver (Robecchi Bricchetti 1903). In the 1930s ambergris was sold by auction and half of the proceeds of the sale was property of the state (Zoli 1927). This material, produced in the alimentary tract of sperm whales (Physeter macrocephalus), consists of a greasy waxy secretion found around squid beaks, squids being the preferred food item of sperm whales (Tinker 1988). Ambergris produces a highly fragrant and spicy smell and was an important ﬁxative in the perfume industry, where it has 142

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

now been replaced by the synthetic molecule ambreina. However, searches for stranded ambergris are still carried out in many countries such as Mauritania, Somalia and India. Because stranding occurs in certain areas of the coast more than others the local toponymy has been inﬂuenced. For example, there are placenames along the coast of Somalia that reﬂect the stranding of this material (Ambershiek). Besides its use in perfumery, ambergris was also burnt as incense or eaten for medicinal purposes, where its effect was comparable to a modern estrogen cure. In the city of Mecca ambergris is still sold, and men and women in Iran eat it to cure infertility. In Morocco it is traditionally used to add aroma to tea. Stranded carrion represents another important organic material that can be exploited by coastal populations of the Indian Ocean. In Somalia, when turtles were washed ashore, it was common to extract their fat. Traditionally this was used by locals as body cream as there was the belief that it had regenerative properties and stimulated male sexual faculties. Also, carrion of the crab Ocypode ryderi was used by coastal tribes to attract the so-called “imbe”, a small gastropod (Marginella monilis) with a white porcelain shell, employed to make necklaces that were sold in the local market (Grottanelli 1955). Stranded shells of the Cypraeidae family (Mollusca Gasteropoda) acquired a historical, cultural and economical importance in past centuries, and were actively sought by coastdwelling peoples. Along the Somalian coast these were collected by women and children and sold for 20–22 shillings per hundred kg (Grottanelli 1955). Used as money by many populations until recently and as ornament for both living and dead since ancient Egyptian times, these shells are now employed in the tourism industry or sold for collections.

Effects of human removal of wracks on invertebrate populations Mechanical beach-cleaning of organic beach-cast material Recently, the increasing use of beaches as recreational areas has pressed regional authorities of many countries to remove all natural ﬂotsam, such as detached macrophytes, driftwood and carrion, together with sanitary refuse and other litter of human origin such as glass, metal, plastic and their derivatives (Ryan & Swanepoel 1996). Mechanical removal has been seen as a cost-effective way of removing unwanted debris and has been employed by many coastal authorities without considering the long-term detrimental consequences on coastal environments. There is growing concern with the issue and some studies have been carried out to evaluate the impact of mechanical beach-cleaning on shores. Davidson et al. (1991) concluded that beach-cleaning machines had a damaging impact on invertebrate populations and that in areas of high recreational pressure the stability of the dunes would also be affected. Kirby (1992) reinforced this concern, stating that the removal of driftwood and especially large jetsam could have a damaging effect on certain isopods and ground beetles that use this debris as shelters. Llewellyn & Shackley (1996) compared four mechanically cleaned sections of Swansea Bay, UK, with a control area with no mechanical cleaning or hand-picking. The survey indicated that mechanical beach-cleaning had a serious deleterious effect on the overall strandline-related species diversity and abundance. Recently, Dugan & Hubbard (in press) came to the same conclusions when comparing groomed with ungroomed beaches during a survey of 15 southern Californian beaches. Both the removal 143

I. COLOMBINI & L. CHELAZZI

of strandline debris by raking machines that scour the sand to a depth of 5 cm, and the compression of sand by heavy equipment such as tractors and trucks may have an adverse effect on the invertebrate communities buried in the sand and on sediment transport (Llewellyn & Shackley 1996). It is interesting to note that in those sections of mechanically cleaned beaches where cleaners were unable to operate, a small remnant population of strandline invertebrates could still be found. It was suggested that management strategies should compromise and clean mechanically only those sections of the beach designed for recreational use, whereas hand-picking methods should be employed to collect tourist and shipping-related litter, leaving most of the organic debris. This would permit invertebrates to recolonise mechanically cleaned areas and would reduce impact on species. In a recent study of the Polish coastline, Weslawski et al. (2000) showed a marked decline in the localities inhabited by the sandhopper Talitrus saltator and of its average density when these were compared with previous recordings. Several reasons such as pollution, climatic changes in storm frequency, severity of winters, the rise in sea level, changes in trophic conditions and increase in recreational use of beaches were all proposed to have caused the decline in the species. Again, mechanical cleaning was regarded as an important limiting factor for sandhoppers and it was stressed that amphipods could still recover if several kilometres of less frequently visited beaches between crowded areas were left untouched.

Inorganic beach-cast material Man-made marine debris Allochthonous input from the sea to shores all over the world includes inorganic debris usually associated with human activities. This material has increased signiﬁcantly in the twentieth century, and the material quality has also changed in line with socio-economic progress. Thus man-made debris has become an important component of beach-cast material. In many cases, according to its nature, it is practically inseparable from macrophyte deposits and thus its presence on beaches deﬁnitely has an inﬂuence on local faunal communities. Beach litter originates from many sources: marine, riverine and the beach itself. Industry, domestic sewage, illegal dumping at sea from ships and public negligence are all factors implicated in the phenomenon. Litter has a detrimental effect on beaches and the marine environment in many ways. Floating and submerged debris can threaten marine mammals, seabirds, turtles, ﬁshes and crustaceans through ingestion and entanglement (Fowler 1987, Laist 1987, Ryan 1987, Bugoni et al. 2001, Tomás et al. 2002). Litter washed ashore by tides or discarded by tourists also has an economic impact. This includes loss of aesthetic value of recreational areas relevant to tourism-generated income (Ryan & Moloney 1990), and loss of ﬁsh catches caused by loss of ﬁshing gear (Dixon & Dixon 1981). Additionally, it may be dangerous to human health, especially in the case of medical, military and some industrial wastes (Dixon & Dixon 1981). Marine debris is now considered a problem worldwide and much work has been accomplished in this context. Several recent studies on beach litter have been undertaken to determine the quality and quantity of man-made debris on a beach at a speciﬁc time, and how it varies in space and 144

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

time (e.g. Frost & Cullen 1997, Haynes 1997, Madzena & Lasiak 1997, Walker et al. 1997, Thornton & Jackson 1998, Velander & Mocogni 1998, Moore et al. 2001, Williams & Tudor 2001a,b). In a study on Israeli beaches in the Mediterranean the importance of beach geomorphology on coastal pollution was shown, and beach width, ridge, and runnel morphology and beach porosity were pointed out as dominant factors that may inﬂuence the deposition of litter (Bowman et al. 1998). Because the composition of litter and its location on beaches is very variable and depends on many physical processes there have been a variety of methods to describe and monitor beach litter. Data collected in different ways make comparisons very difﬁcult and standardisation impossible. Velander & Mocogni (1999) compared 10 methods for sampling beach litter on 16 beaches in Scotland to ascertain the effectiveness of the various methods. It was concluded that different methods did produce signiﬁcantly different results and each method had advantages and disadvantages. Thus selection of a collection method will depend on the information required. In other studies an evaluation of the status of debris was made using a set of indicator items to identify trends over time (Rees & Pond 1995, Ribic 1998). However, it was concluded that caution should be used when collecting data by this method because selected indicator items do not always reﬂect general trends of the remaining litter on the beach (Ribic 1998). Litter clearance on beaches is a main issue of coastal management. It can be seen as an instructive exercise when members of local communities are involved but it has been demonstrated that it is only a temporary management strategy (Williams & Tudor 2001a,b). In fact, depending on the sites, beaches become littered very rapidly again and on some occasions buried litter can re-emerge at the sand surface (Williams & Tudor 2001b). Costs of litter picking have greatly increased so beach managers are looking for alternative management options. Moreover, beach cleaning does not solve the problem, which should be tackled at the source (Velander & Mocogni 1998, Williams & Tudor 2001a,b). In some cases sources are easy to identify because they are local (Uneputty & Evans 1997, Willoughby et al. 1997) but in coastal situations sources can be international, with litter originating from other countries. For example, American litter may be deposited on western European shores (Olin et al. 1995). Management strategies therefore need to overcome international barriers and solve political and practical problems. In the past, source location has focused on the minutiae of identifying items of the litter, whereas now it has been understood that a combination of both a good taxonomy and methodologies that look at the mixtures of litter is required. Earll et al. (2000) stressed that prevention at the source will become a reality only when stronger links are established between measurement and management. Only by providing measured proﬁles of the quality of litter types, can the trends in the input of items be assessed and prevention programmes be applied directly to sources.

Inﬂuence of man-made debris on beach ecosystems The presence of anthropogenic debris on beaches is generally viewed as an aesthetic degradation of wilderness values. In developing countries, the presence of litter on beaches is seen as an important source of materials (plastic containers, nets, ropes) actively sought and reused by local coastal populations. However, in most developed countries beach litter is an economically important issue and each year management measures are taken to remove tons of litter stranded ashore. Apart from having a public service and educational value, this practice is extremely costly, does not solve the problem and, according to the methods 145

I. COLOMBINI & L. CHELAZZI

employed, it can cause great damage to beach communities resulting in a decrease in species richness and abundance. The use of large mechanical beach cleaners, trucks to collect litter and tractors to remove large debris (refrigerators, television sets, tyres, plastic and ﬁbreglass car and boat pieces, vessel ropes) all have detrimental effects on plants and animals. However, some invertebrate and vertebrate species may exploit man-made objects as shelters, which provide the microclimatic characteristics necessary for the survival of speciﬁc species. Some forms of beach litter may collect rainfall and thus offer a precious source of freshwater. Others can act as lure traps, as in the case of plastic drinking bottles, which attract invertebrates with their sugar contents and become mortal traps as these dry out or when removed by cleaners (I. Colombini & L. Chelazzi, pers. obs.). Large items of debris also attract scototatic animals from great distances and result in a massive concentration of species (Colombini & Chelazzi 1991, Colombini et al. 1994). When this debris is collected, enormous quantities of invertebrates are trapped and removed from beaches with a consequent decrease in biodiversity (I. Colombini & L. Chelazzi, pers. obs.). Furthermore, trampling on beaches and foredunes with mechanical beach cleaners physically affects buried species and pioneer vegetation of foredunes. The destruction of foredune plants, and in particular of those that have a binding function, has a destabilising effect on the sand sediment and results in the subsequent erosion of the dune (Brown & McLachlan 1990). Sometimes litter is collected and piled up by bulldozers and burned or buried directly on the beach. This also is a very destructive method and has a heavy impact on the faunal communities and on the stability of the beach itself (I. Colombini & L. Chelazzi, pers. obs.).

Conclusions This review has emphasised the importance of allochthonous input from the sea to beach and nearshore communities. Invertebrates and vertebrate species (including man), in some way or another, have always taken advantage of these subsidies. In some cases, species have become totally dependent upon their presence on beaches and have evolved biological cycles entirely linked to beach casts. In other cases, species have adapted their abundance, their spatial distribution and their feeding ecology in relation to the spatial and temporal patterns of wrack deposition. Studies on species succession in macrophyte wracks have indicated that these subsidies stand at the base of the beach food chains and that many species need an interspeciﬁc facilitation process before they can exploit the resource. All levels of the food chain are intimately linked to one another and sometimes there are great advantages for secondary consumers that, when subsidised, can reach higher population densities. The exploitation of man-made debris by certain beach species has been made possible through behavioural plasticity, an important key factor that has been rigorously selected for in the evolution of the sandy beach macrofauna living in such a harsh environment. Also, humans have beneﬁted from beach-cast material over the centuries and still make use of important macrophyte deposits for economic reasons. Beach ecotones are naturally unpredictable environments, continuously evolving and heavily inﬂuenced by both sea and land. Being restricted in width these environments can be directly affected by human encroachment from land or indirectly by marine and riverine input. Impacts on beaches can vary according to a number of factors, which include human population densities along coastlines, vicinity to river mouths and delta, economy of coun146

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

tries, presence of tourism, and geographical position. Beaches along seas (Red Sea, Black Sea, Mediterranean Sea, Baltic Sea) suffer more anthropogenic impacts compared with those of oceans because of higher human concentration. For example, in the Mediterranean, beaches have been under the inﬂuence of man and of his activities for more than 2000 yr and have been gradually modiﬁed through time. However, in the past 50 yr, changes have been extremely rapid and disruptive and consequently have seriously jeopardised the equilibrium of most Mediterranean beach ecosystems. Rapid social changes associated with the need to exploit beaches for tourism revenue or through harvesting of beach casts have produced environmental changes of beaches all over the world. An increasing awareness of the seriousness and urgency of the problem has induced many modern scientists to concentrate much of their effort on understanding how beach ecosystems operate and of the main forcing factors implicated in the system. Linkages between marine and terrestrial food webs are now becoming clearer although much work still needs to be done on the dynamics of the system and on how river plumes inﬂuence beach and nearshore communities (see Gillanders & Kingsford 2002 for review). With an increase in the knowledge of ecosystem functioning there might be a chance for future sustainable management of beach ecosystems and managers, decision makers and local authorities might become aware of the long-term consequences of their actions. There is the hope that different policies will be adopted in the removal of organic and man-made debris from beaches or in prevention at the source. An alternative could be the protection of certain areas at regular intervals along the coast that could serve as buffer areas where reasonable levels of biodiversity could be maintained. More communication needs to be achieved between coastal managers and scientists and this may represent a challenging task for the future.

Acknowledgements The authors would like to thank many authors in the reference list who promptly sent their articles to us and have made this review possible. Special thanks go to Professor Elfed Morgan for his critical reading of and useful comments on the manuscript.

References Abell, D. H., Wasti, S. S. & Hartmann, G. C. 1982. Saprophagous arthropod fauna associated with turtle carrion. Applied Entomology and Zoology 17, 301–307. Adams, S. M. 1976. The ecology of eelgrass Zostera marina (L.) ﬁsh communities. I. Structural analysis. Journal of Experimental Marine Biology and Ecology 22, 291–293. Agnew, D. J. & Moore, P. G. 1986. The feeding ecology of two littoral amphipods (Crustacea) Echinogammarus pirloti (Sexton & Spooner) and E. obtusatus (Dahl). Journal of Experimental Marine Biology and Ecology 103, 203–215. Alkemade, R. & van Rijswijk, P. 1993. Path analysis of the inﬂuence of substrate composition on nematode numbers and on decomposition of stranded seaweed at an Antarctic coast. Netherlands Journal of Sea Research 31, 63–70.

147

I. COLOMBINI & L. CHELAZZI

Anderson, W. B. & Polis, G. A. 1998. Marine subsidies of island communities in the Gulf of California: evidence from stable carbon and nitrogen isotopes. Oikos 81, 75–80. Andriashek, D., Kiliaan, P. H. L. & Taylor, M. K. 1985. Observations on foxes Alopex lagopus and Vulpes vulpes, and wolves, Canis lupus, on the off-shore sea ice of northern Labrador. Canadian Field-Naturalist 99, 86–89. Avery, G., Avery, D. M., Braine, S. & Loutit, R. 1987. Prey of coastal black-backed jackal Canis mesomeles (Mammalia: Canidae) in the Skeleton Coast Park, Namibia. Canadian Journal of Zoology 213, 81–94. Barberio, C., Di Cello, F., Chelazzi, L., Colombini, I., Fallaci, M. & Fani, R. 2001. Molecular and physiological analysis of mesophilic aerobic heterotrophic bacteria from the gut of two species of the genus Phaleria Latreille, 1802 (Coleoptera, Tenebrionidae). Annals of Microbiology 51, 95–105. Barilotti, D. C. & Zertuche-González, J. A. 1990. Ecological effects of seaweed harvesting in the Gulf of California and Paciﬁc Ocean off Baja California and California. Hydrobiologia 204, 35–40. Bedford, A. P. & Moore, P. G. 1984. Macrofaunal involvement in the sublittoral decay of kelp debris: the detritivore community and species interactions. Estuarine, Coastal and Shelf Science 18, 97–111. Behbehani, M. I. & Croker, R. A. 1982. Ecology of beach wrack in northern New England with special reference to Orchestia platensis. Estuarine, Coastal and Shelf Science 15, 611–620. Berg, B. S. & Munkejord, A. A. 1991. Forsvinner Jærstrendene? 1991 Årsrapport for miljøvernavdelingen ved Fylkesmannen i Rogaland. Rogaland Country Environmental Department, Norway, 19–26. Bertness, M. D. & Ellison, A. M. 1987. Determinants of pattern in a New England salt marsh community. Ecological Monographs 57, 129–147. Bigot, L. 1970. Contribution a l’étude écologique des peuplements halophiles de la région de Tuléar (Madagascar) II. La plage et le cordon dunaire. Annales de l’Université de Madagascar 7, 159–163. Blanche, K. R. 1992. Preliminary observations on the distribution and abundance of seaweed ﬂies (Diptera: Coelopidae) on beaches in the Gosford district of New South Wales, Australia. Australian Journal of Ecology 17, 27–34. Bodkin, J. L. & Jameson, R. J. 1991. Patterns of seabird and marine mammal carcass deposition along the central California coast, 1980–1986. Canadian Journal of Zoology 69, 1149–1155. Botton, M. L., Loveland, R. E. & Jacobson, T. R. 1994. Site selection by migratory shorebirds in Delaware Bay and its relationship to beach characteristics and abundance of horseshoe crab (Limulus polyphemus). The Auk 111, 605–616. Bowman, D., Manor-Samsonov, N. & Golik, A. 1998. Dynamics of litter pollution on Israeli Mediterranean beaches: a budgetary, litter ﬂux approach. Journal of Coastal Research 14, 418–432. Bradley, R. A. & Bradley, D. W. 1993. Wintering shorebirds increase after kelp (Macrocystis) recovery. The Condor 95, 372–376. Braﬁeld, A. E. & Llewellyn, M. J. 1982. Animal energetics. New York: Chapman & Hall. Brewer, J. S., Levine, J. M. & Bertness, M. D. 1998. Interactive effects of elevation and burial with wrack on plant community structure in some Rhode Island salt marshes. Journal of Ecology 86, 125–136. Bridgeford, P. 1985. Unusual diet of the lion (Panthera leo) in the Skeleton Coast Park. Madoqua 14, 187–188. Britton, J. C. & Morton, B. 1992. The ecology and feeding behaviour of Nassarius festivus (Prosobranchia: Nassariidae) from two Hong Kong Bays. In The marine ﬂora and fauna of Hong Kong and Southern China III. Proceedings of the Fourth International Marine Biological Workshop: the marine ﬂora and fauna of Hong Kong and Southern China, Hong Kong, 1989, B. Morton (ed.). Hong Kong: Hong Kong University Press, 395–416. Britton, J. C. & Morton, B. 1993. Are there obligate marine scavengers? In The marine biology of the

148

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

South China Sea I. Proceedings of the International Conference on the Marine Biology of Hong Kong and the South China Sea, Hong Kong, 1990, B. Morton (ed.). Hong Kong: Hong Kong University Press, 357–391. Britton, J. C. & Morton, B. 1994. Marine carrion and scavengers. Oceanography and Marine Biology: an Annual Review 32, 369–434. Brown, A. C. & McLachlan, A. 1990. Ecology of sandy shores. Amsterdam: Elsevier Science. Brown, A. C. & Odendaal, F. J. 1994. The biology of oniscid Isopoda of the genus Tylos. Advances in Marine Biology 30, 89–153. Bugoni, L., Krause, L. & Petry, M. V. 2001. Marine debris and human impacts on sea turtles in Southern Brazil. Marine Pollution Bulletin 42, 1330–1334. Burger, A. E. 1985. Terrestrial food webs in the sub-Antarctic: island effects. In Antarctic nutrient cycles and food webs, W. R. Siegfried et al. (eds). New York: Springer Verlag, 582–591. Bustamante, R. H. & Branch, G. M. 1996. Dependence of intertidal consumers on kelp-derived organic matter on the west coast of South Africa. Journal of Experimental Marine Biology and Ecology 196, 1–28. Bustamante, R. H., Branch, G. M. & Eekhout, S. 1995. Maintenance of an exceptional grazer biomass on South African intertidal shores: trophic subsidy by subtidal kelps. Ecology 76, 2314–2329. Butlin, R. K., Collins, P. M. & Day, T. H. 1984. The effect of larval density on an inversion polymorphism in the seaweed ﬂy Coelopa frigida. Heredity 52, 415–423. Butlin, R. K. & Day, T. H. 1985. Adult size, longevity and fecundity in the seaweed ﬂy Coelopa frigida. Heredity 54, 107–110. Butlin, R. K. & Day, T. H. 1989. Environmental correlates of inversion frequencies in natural populations of seaweed ﬂies (Coelopa frigida). Heredity 62, 223–232. Butlin, R. K., Read, I. L. & Day, T. H. 1982. The effects of a chromosomal inversion on adult size and male mating success in the seaweed ﬂy Coelopa frigida. Heredity 49, 51–62. Callan, E. McC. 1964. Ecology of sand dunes with special reference to the insect communities. In Ecological studies in southern Africa, J. H. Davis (ed.). The Hague: Junk, 174–185. Carefoot, T. H., Taylor, B. E. & Brett, K. 1998. A day in the life of an isopod: time and energy allocations in the semiterrestrial Ligia pallasii. Israel Journal of Zoology 44, 463–471. Castro, G. & Myers, J. P. 1993. Shorebird predation on eggs of horseshoe crabs during spring stopover on Delaware Bay. The Auk 110, 927–930. Caussanel, C. 1965. Recherches préliminaires sur le peuplement de coléoptères d’une plage sableuse Atlantique. Annales de la Societé Entomologique de France (N.S.) 1, 197–248. Caussanel, C. 1970. Contribution à l’étude du peuplement d’une plage et d’une dune Landaise. Vie et Milieu 21, 59–104. Chapman, A. R. O. & Craigie, J. S. 1978. Seasonal growth in Laminaria longicruris: relations with reserve carbohydrate storage and production. Marine Biology 46, 209–213. Chelazzi, L., Colombini, I., Bertin, G., Cianfanelli, A., Fallaci, M., Lucarelli, E. & Mascagni, A. 1990. Gli artropodi del tombolo antistante la laguna di Burano (GR): ambiente, microclima e primi dati sul popolamento animale. Redia 73, 307–345. Chelazzi, G. & Ferrara, F. 1978. Researches on the coast of Somalia. The shore and the dune of Sar Uanle. 19. Zonation and activity of terrestrial isopods (Oniscoidea). Monitore Zoologico Italiano (Nuovo Supplemento) 11, 189–219. Cheung, S. G. 1994. Feeding behaviour and activity of the scavenging gastropod Nassarius festivus (Powys). In The malacofauna of Hong Kong and Southern China III. Proceedings of the Third International Workshop on the malacofauna of Hong Kong and Southern China, Hong Kong, 1992, B. Morton (ed.). Hong Kong: Hong Kong University Press, 327–338. Chevin, H. 1998. La vie dans les varechs échoués sur la plage. Insectes 109, 9–10. Chown, S. L. 1996. Kelp degradation by Paractora trichosterna (Thomson) (Diptera: Helcomyzidae) at sub-antarctic South Georgia. Polar Biology 16, 171–178.

149

I. COLOMBINI & L. CHELAZZI

Clark, K. E., Niles, L. J. & Burger, J. 1993. Abundance and distribution of migrant shorebirds in Delaware Bay. The Condor 95, 694–705. Colombini, I., Aloia, A., Bouslama, M. F., ElGtari, M., Fallaci, M., Ronconi, L., Scapini, F. & Chelazzi, L. 2002. Small-scale spatial and seasonal differences in the distribution of beach arthropods on the north western Tunisian coast. Are species evenly distributed on the shore? Marine Biology 140, 1001–1012. Colombini, I., Aloia, A., Fallaci, M., Pezzoli, G. & Chelazzi, L. 2000. Temporal and spatial use of stranded wrack by the macrofauna of a tropical sandy beach. Marine Biology 136, 531–541. Colombini, I. & Chelazzi, L. 1991. A comparison between the life cycles of different populations of Eurynebria complanata (Linnaeus, 1767) (Coleoptera: Carabidae). Elytron Supplement 5, 5–14. Colombini, I., Chelazzi, L. & Scapini, F. 1994. Solar and landscape cues as orientation mechanisms in the beach-dwelling beetle Eurynebria complanata (Coleoptera, Carabidae). Marine Biology 118, 425–432. Connell, J. H. & Slatyer, R. O. 1977. Mechanism of succession in natural communities and their role in community stability and organisation. The American Naturalist 111, 1119–1144. Crafford, J. E. 1984. Life cycle and kelp consumption of Paractora dreuxi mirabilis (Diptera: Helcomyzidae): a primary decomposer of stranded kelp on Marion Island. South African Journal of Antarctic Research 14, 18–22. Crafford, J. E. & Scholtz, C. H. 1987. Phenology of stranded kelp degradation by the kelp ﬂy Paractora dreuxi mirabilis (Helcomyzidae) at Marion Island. Polar Biology 7, 289–294. Craig, P. C. 1970. The behaviour and distribution of the intertidal sand beetle, Thinopinus pictus (Coleoptera, Staphylinidae). Ecology 51, 1012–1017. Craig, P. C. 1973. Behaviour and distribution of the sand-beach amphipod Orchestoidea corniculata. Marine Biology 23, 101–109. Cullen, S. J., Young, A. M. & Day, T. H. 1987. Dietary requirements of seaweed ﬂies (Coelopa frigida). Estuarine, Coastal and Shelf Science 24, 701–710. Davidson, N. S., D’Alaffoley, D., Doody, J. P., Way, L. S., Gordon, J., Key, R., Drake, C. M., Pienkowski, M. W., Mitchell, R. & Duff, K. L. 1991. Nature conservation and estuaries in Great Britain. Peterborough: Nature Conservation Council. Day, T. H., Dawe, C., Dobson, T. & Hillier, P. C. 1983. A chromosomal inversion polymorphism in Scandinavian populations of the seaweed ﬂy, Coelopa frigida. Hereditas 99, 135–145. Day, T. H., Dobson, T., Hillier, P. C., Parkin, D. T. & Clarke, B. C. 1980. Different rates of development associated with the alcohol dehydrogenase locus in the seaweed ﬂy, Coelopa frigida. Heredity 44, 321–326. Day, T. H., Dobson, T., Hillier, P. C., Parkin, D. T. & Clarke, B. C. 1982. Associations of enzymic and chromosomal polymorphisms in the seaweed ﬂy, Coelopa frigida. Heredity 48, 35–44. Day, T. H., Miles, S., Pilkington, M. D. & Butlin, R. K. 1987. Differential mating success in populations of seaweed ﬂies (Coelopa frigida). Heredity 58, 203–212. Dayton, P. K. 1985. Ecology of kelp communities. Annual Review of Ecology and Systematics 16, 215–245. den Hartog, C. 1970. Seagrasses of the world. Amsterdam: North Holland. Dennis, J. V. & Gunn, C. R. 1971. Case against trans-Paciﬁc dispersal of the coconut by ocean currents. Economic Botany 25, 407–413. De Rougemont, G. 2000. Beetles in seaweed in Hong Kong. Porcupine 21, 6–7. De Roy Moore, T. 1980. Galapagos Islands lost in time. London: George Allen & Unwin. Dixon, T. R. & Dixon, T. J. 1981. Marine litter surveillance. Marine Pollution Bulletin 12, 289–295. Dobson, T. 1974a. Studies on the biology of the kelp-ﬂy Coelopa in Great Britain. Journal of Natural History 8, 155–177. Dobson, T. 1974b. Mechanisms controlling species composition in natural populations of the seaweed ﬂy Coelopa. Journal of Natural History 8, 653–673. Dobson, T. 1976. Seaweed ﬂies (Diptera: Coelopidae, etc.). In Marine insects, L. Cheng (ed.). Amsterdam: North-Holland, 447–463.

150

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

Donn, T. E. 1990, Zonation patterns of Donax serra Röding (Bivalva: Donacidae) in southern Africa. Journal of Coastal Research 6, 903–911. Due, A. D. & Polis, G. A. 1985. Biology of the intertidal scorpion, Vaejovis littoralis. Journal of Zoology 207, 563–580. Dugan, J. E. & Hubbard, D. M. in press. The response of macrofauna communities and shorebirds to macrophyte wrack subsidies on exposed sandy beaches of Southern California. Estuarine, Coastal and Shelf Science. Duggins, D. O., Simenstad, C. A. & Estes, J. A. 1989. Magniﬁcation of secondary production by kelp detritus in coastal marine ecosystems. Science 245, 170–173. Earll, R. C., Williams, A. T., Simmons, S. L. & Tudor, D. T. 2000. Aquatic litter, management and prevention – the role of measurement. Journal of Coastal Conservation 6, 67–78. Egglishaw, H. J. 1960. Studies on the family Coelopidae (Diptera). The Transactions of the Royal Entomological Society of London 112, 109–140. Field, J. G., Grifﬁths, C. L., Grifﬁths, R. J. Jarman, N., Zoutendyk, P., Velimirov, B. & Bowes, A. 1980a. Variation in structure and biomass of kelp communities along the south-west Cape coast. Transactions of the Royal Society of South Africa 44, 145–203. Field, J. G., Grifﬁths, C. L., Linley, E., Carter, R. A. & Zoutendyk, P. 1980b. Upwelling in a nearshore marine ecosystem and its biological implications. Estuarine and Coastal Marine Science 11, 133–150. Field, J. G., Jarman, N., Dieckmann, G. S., Grifﬁths, C. L., Velimirov, B. & Zoutendyk, P. 1977. Sun, waves, seaweed and lobsters: the dynamics of a west coast kelp-bed. South African Journal of Science 73, 7–10. Fielding, P. J. & Davis, C. L. 1989. Carbon and nitrogen resources available to kelp bed ﬁlter feeders in an upwelling environment. Marine Ecology Progress Series 55, 181–189. Fischer, J. M., Klug, J. L., Reed-Anderson, T. & Chalmers, A. G. 2000. Spatial pattern of localised disturbance along southeastern salt marsh tidal creek. Estuaries 23, 565–571. Fonseca, M. S. & Fisher, J. S. 1986. A comparison of canopy friction and sediment movement between four species of seagrass with reference to their ecology and restoration. Marine Ecology Progress Series 29, 15–22. Fowler, C. W. 1987. Marine debris and northern fur seals: a case study. Marine Pollution Bulletin 18, 326–335. Fretey, J. & Babin, R. 1998. Arthropod succession in leatherback turtle carrion and implications for determination of the postmortem interval. Marine Turtle Newsletter 79, 4–7. Frost, A. & Cullen, M. 1997. Marine debris on northern New South Wales beaches (Australia): sources and the role of beach usage. Marine Pollution Bulletin 34, 348–352. Fry, B. & Virnstein, R. W. 1988. Leaf production and export of the seagrass Syringodium ﬁliforme Kütz. in Indian River lagoon, Florida. Aquatic Botany 30, 261–266. Gillanders, B. M. & Kingsford, M. J. 2002. Impact of changes in ﬂow of freshwater on estuarine and open coastal habitats and the associated organisms. Oceanography and Marine Biology: an Annual Review 40, 233–309. Grant, G. S. 1997. Sanderling exploit beached animal carcasses as a source of ﬂy larvae. The Chat 61, 104–107. Green, P. T. 1999. Greta’s Garbo: stranded seeds and fruits from Greta Breach, Christmas Island, Indian Ocean. Journal of Biogeography 26, 937–946. Grifﬁths, C. L. & Stenton-Dozey, J. 1981. The fauna and the rate of degradation of stranded kelp. Estuarine, Coastal and Shelf Science 12, 645–653. Grifﬁths, C. L., Stenton-Dozey, J. & Koop, K. 1983. Kelp wrack and the ﬂow of energy through a sandy beach ecosystem. In Sandy beaches as ecosystems, A. McLachlan & T. Erasmus (eds). The Hague: Junk, 547–556. Grottanelli, V. L. 1955. Pescatori dell’Oceano Indiano. Roma, Cremonese. Grover, P. B. & Knopf, F. L. 1982. Habitat requirements and breeding success of charadriiform birds

151

I. COLOMBINI & L. CHELAZZI

nesting at Salt Plains National Wildlife Refuge, Oklahoma. Journal of Field Ornithology 53, 139–148. Gunn, C. R. & Dennis, J. V. 1973. Tropical and temperate stranded seeds and fruits from the Gulf of Mexico. Contributions in Marine Science 17, 111–121. Hammond, A. L. (ed.) 1990. World Resources: 1990–91. New York: Oxford University Press. Hansen, J. A. 1983. Macrophyte decomposition in beach wrack and surf-zone accumulations along the Western Australian coast. Bulletin of the Australian Marine Science Association 83, 19 only. Hansen, J. A. 1984. Accumulation of macrophyte wrack along sandy beaches in Western Australia: biomass, decomposition rate and signiﬁcance in supporting nearshore production. PhD thesis, University of Western Australia, Nedlands, WA. Harries, H. C. 1983. The coconut palm, the robber crab and Charles Darwin: April fool or a curious case of instinct? Principes 27, 131–137. Harrison, P. G. 1982. Control of microbial growth and of amphipod grazing by water-soluble compounds from leaves of Zostera marina. Marine Biology 67, 225–230. Harrison, P. G. & Mann, K. H. 1975. Detritus formation from eelgrass (Zostera marina): the relative effects of fragmentation, leaching and decay. Limnology and Oceanography 20, 924–934. Hartwig, E. O. 1978. Factors affecting respiration and photosynthesis by the benthic community of subtidal siliceous sediment. Marine Biology 46, 283–293. Haslam, S. F. I. & Hopkins, D. W. 1996. Physical and biological effects of kelp (seaweed) added to soil. Applied Soil Ecology 3, 257–261. Hay, M. E. & Fenical, W. 1988. Marine plant-herbivore interactions: the ecology of chemical defence. Annual Review of Ecology and Systematics 19, 111–145. Hayes, W. B. 1974. Sand beach energetics: importance of the isopod, Tylos punctatus. Ecology, New York 55, 838–847. Hayes, W. B. 1977. Factors affecting the distribution of Tylos punctatus (Isopoda, Oniscoidea) on beaches in Southern California and Northern Mexico. Paciﬁc Science 31, 165–186. Haynes, D. 1997. Marine debris on continental islands and sand cays in the far northern section of the Great Barrier Reef Marine Park, Australia. Marine Pollution Bulletin 34, 276–279. Heard, S. B. 1994. Processing chain ecology: resource condition and interspeciﬁc interactions. Journal of Animal Ecology 63, 451–464. Heip, C., Vincx, M. & Vranken, G. 1985. The ecology of marine nematodes. Oceanography and Marine Biology: an Annual Review 23, 399–489. Hellemaa, P. 1998. The development of coastal dunes and their vegetation in Finland. Fennia 176, 111–221. Hemminga, M. A., Gwada, P., Slim, F. J., de Koeyer, P. & Kazungu, J. 1995. Leaf production and nutrient contents of the seagrass Thalassodendron ciliatum in proximity of a mangrove forest (Gazi Bay, Kenya). Aquatic Botany 50, 159–170. Hemminga, M. A. & Nieuwenhuize, J. 1990. Seagrass wrack-induced dune formation on a tropical coast (Banc D’Arguin, Mauritania). Estuarine, Coastal and Shelf Science 31, 499–502. Hemminga, M. A. & Nieuwenhuize, J. 1991. Transport, deposition and in situ decay of seagrasses in a tropical mudﬂat area (Banc D’Arguin, Mauritania). Netherlands Journal of Sea Research 27, 183–190. Hersteinsson, P. & MacDonald, D. W. 1982. Some comparisons between red and Arctic foxes, Vulpes vulpes and Alopex lagopus, as revealed by radio tracking. Symposium of the Zoological Society of London 49, 259–289. Hesp, P. A. 1984. The formation of sand beaches, ridges and foredunes. Search 15, 289–291. Hesp, P. A. 1991. Ecological processes and plant adaptations on coastal dunes. Journal of Arid Environments 21, 165–191. Hiscocks, K. & Perrin, M. R. 1987. Feeding observations and diet of black-backed jackals in an arid coastal environment. South African Journal of Wildlife Reserves 17, 55–58.

152

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

Hobday, A. J. 2000. Abundance and dispersal of drifting kelp Macrocystis pyrifera rafts in the Southern California Bight. Marine Ecology Progress Series 195, 101–116. Hodge, S. & Arthur, W. 1996. Contramensal interactions between species. Oikos 77, 371–375. Hodge, S. & Arthur, W. 1997. Asymmetric interaction between species of seaweed ﬂy. Journal of Animal Ecology 66, 743–754. Hodge, S., Arthur, W. & Mitchell, P. 1996. Effects of temporal priority on interspeciﬁc interactions and community development. Oikos 76, 350–358. Inglis, G. 1989. The colonisation and degradation of stranded Macrocystis pyrifera.(L.) C. Ag. by the macrofauna of a New Zealand sandy beach. Journal of Experimental Marine Biology and Ecology 125, 203–217. Jarman, N. G. & Carter, R. A. 1981. The primary producers of the inshore regions of the Benguela. Transaction of the Royal Society of South Africa 44, 321–326. Jolivet, E., de Langlais-Jeannin, I. & Morot-Gaudry, J. F. 1991. Les extraits d’algues marines: propriété phytoactives et intérêt agronomique. L’Année Biologique 30, 109–126. Josselyn, M. N., Cailliet, G. M., Niesen, T. M., Cowen, R., Hurley, A. C., Connor, J. & Hawes, S. 1983. Composition, export, and faunal utilisation of drift vegetation in the salt river submarine canyon. Estuarine, Coastal and Shelf Science 17, 447–465. Kensley, B. 1974. Aspects of the biology and ecology of the genus Tylos Latreille. Annals of the South African Museum 65, 401–471. Kikuchi, T. 1974. Japanese contribution on consumer ecology in eelgrass (Zostera marina L.) beds with special reference to trophic relationships and resources in inshore ﬁsheries. Aquaculture 4, 145–160. Kirby, P. 1992. Habitat management for invertebrates: a practical handbook. Joint Nature Conservation Committee. Publication of the Royal Society for the Protection of Birds, London. Kirkman, H. 1984. Standing stock and production of Ecklonia radiata (C.Ag.) J. Agardh. Journal of Experimental Marine Biology and Ecology 76, 119–130. Kirkman, H. 1985. Community structure in seagrasses in southern Western Australia. Aquatic Botany 21, 363–375. Kirkman, H. & Kendrick, G. A. 1997. Ecological signiﬁcance and commercial harvesting of drifting and beach-cast macro-algae and seagrasses in Australia: a review. Journal of Applied Phycology 9, 311–326. Kirkman, H. & Kuo, J. 1990. Pattern and process in southern Western Australian seagrasses. Aquatic Botany 37, 367–382. Klok, C. J. & Chown, S. L. 2001. Critical thermal limits, temperature tolerance and water balance of a sub-Antarctic kelp ﬂy Paractora dreuxi (Diptera: Helcomyzidae). Journal of Insect Physiology 47, 95–109. Kompfrner, H. 1974. Larvae and pupae of some wrack dipterans on a California beach. The PanPaciﬁc Entomologist 50, 44–52. Koop, K. & Field, J. G. 1980. The inﬂuence of food availability on the population dynamics of a supralittoral isopod Ligia dilatata Brandt. Journal of Experimental Marine Biology and Ecology 48, 61–72. Koop, K. & Field, J. G. 1981. Energy transformation by the supralittoral isopod Ligia dilatata Brandt. Journal of Experimental Marine Biology and Ecology 53, 221–233. Koop, K. & Grifﬁths, C. L. 1982. The relative signiﬁcance of bacteria, meio- and macrofauna on an exposed sandy beach. Marine Biology 66, 295–300. Koop, K., Newell, R. C. & Lucas, M. I. 1982a. Microbial regeneration of nutrients from the decomposition of macrophyte debris on the shore. Marine Ecology Progress Series 9, 91–96. Koop, K., Newell, R. C. & Lucas, M. I. 1982b. Biodegradation and carbon ﬂow based on kelp (Ecklonia maxima) debris in a sandy beach microcosm. Marine Ecology Progress Series 7, 315–326. Koskela, H. & Hanski, I. 1977. Structure and succession in a beetle community inhabiting cow dung. Annales Zoologica Fennici 14, 204–223.

153

I. COLOMBINI & L. CHELAZZI

Kuo, J. & McComb, A. J. 1989. Biology of seagrasses. A treatise on the biology of seagrasses with special reference to the Australian region. In Aquatic Plant Studies 2, A. W. D. Larkum et al. (eds). Amsterdam: Elsevier Science, 6–73. Laist, D. W. 1987. Overview of the biological effects of lost and discarded plastic debris in the marine environment. Marine Pollution Bulletin 18, 319–326. Lavoie, D. 1985. Population dynamics and ecology of beach wrack macroinvertebrates of the central California coast. Bulletin of the Southern California Academy of Science 84, 1–22. Leach, J. H. 1970. Epibenthic algal production in an intertidal mudﬂat. Limnology and Oceanography 12, 514–521. Leggett, M. C., Wilcockson, R. W., Day, T. H., Phillips, D. S. & Arthur, W. 1996. The genetic effects of competition in seaweed ﬂies. Biological Journal of the Linnean Society 57, 1–11. Lenanton, R. C. J., Robertson, A. I. & Hansen, J. A. 1982. Nearshore accumulations of detached macrophytes as nursery areas for ﬁsh. Marine Ecology Progress Series 9, 51–57. Llewellyn, P. J. & Shackley, S. E. 1996 The effects of mechanical beach-cleaning on invertebrate populations. British Wildlife 7, 147–155. Lopez, G. R., Levinton, J. S. & Slobodkin, L. B. 1977. The effect of grazing by detritivore Orchestia grillus on Spartina litter and its microbial cover. Oecologia 30, 111–127. Lopez-Uriate, E., Escofet, A., Palacios E. & González, S. 1997. Migrant shorebirds at sandy beaches located between two major wetlands on the Paciﬁc coast of Baja California (Mexico). Natural Areas Journal 17, 212–218. Lord, W. D. & Burger, J. F. 1984a. Arthropods associated with herring gull (Larus argentatus) and great black-backed gull (Larus marinus) carrion on islands in the Gulf of Maine. Environmental Entomology 13, 1261–1268. Lord, W. D. & Burger, J. F. 1984b. Arthropods associated with harbour seal (Phoca vitulina) carcasses stranded along the New England coast. International Journal of Entomology 26, 282–285. Løvås, S. M. & Tørum, A. 2001. Effect of the kelp Laminaria hyperborea upon sand dune erosion and water particle velocities. Coastal Engineering 44, 37–63. Madzena, A. & Lasiak, T. 1997. Spatial and temporal variations in beach litter on the Transkei coast of South Africa. Marine Pollution Bulletin 34, 900–907. Mann, K. H. 1972. Macrophyte production and detritus food chains in coastal waters. Memorie dell’Istituto Italiano di Idrobiologia 29, 353–383. Mann, K. H. 1973. Seaweeds: their productivity and strategy for growth. Science 182, 975–981. Mann, K. H. 1982. Ecology of coastal waters. A system approach. Berkeley: University of California Press. Mann, K. H. 1988. Production and use of detritus in various freshwater, estuarine, and coastal marine ecosystems. Limnology and Oceanography 33, 910–930. Marsden, I. D. 1991a. Kelp-sandhopper interactions on a sand beach in New Zealand. I. Drift composition and distribution. Journal of Experimental Marine Biology and Ecology 152, 61–74. Marsden, I. D. 1991b. Kelp-sandhopper interactions on a sand beach in New Zealand. II. Population dynamics of Talorchestia quoyana (Milne-Edwards). Journal of Experimental Marine Biology and Ecology 152, 75–90. McGwynne, L. E. 1980. A comparative ecophysiological study of three sandy beach gastropods in the eastern Cape. MSc thesis, University of Port Elisabeth, South Africa. McGwynne, L. E., McLachlan, A. & Furstenburg, J. P. 1988. Wrack break-down on sandy beaches. Its impact on interstitial meiofauna. Marine Environmental Research 25, 213–232. McIntyre, A. D. 1968. The meiofauna and macrofauna of some tropical beaches. Journal of Zoology, London 156, 377–392. McKinnerney, M. 1978. Carrion communities in the northern Chihuahuan Desert. Southwest Nature 23, 563–576. McLachlan, A. 1977. Composition, distribution, abundance and biomass of the macrofauna and meiofauna of four sandy beaches. Zoologica Africana 12, 276–306.

154

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

McLachlan, A. 1980. Intertidal zonation of macrofauna and stratiﬁcation of meiofauna on high energy sandy beaches in the Eastern Cape, South Africa. Transactions of the Royal Society of South Africa 44, 213–223. McLachlan, A. 1985. The biomass of macro- and interstitial fauna on clean and wrack-covered beaches in Western Australia. Estuarine, Coastal and Shelf Science 21, 587–599. McLachlan, A., Eliot, I. G. & Clarke, D. J. 1985. Water ﬁltration through microtidal beaches and shallow sublittoral sands and its implications for an inshore ecosystem in Western Australia. Estuarine, Coastal and Shelf Science 21, 91–104. McLachlan, A., Erasmus, T., Dye, A. H., Wooldridge, T., van der Horst, G., Rossouw, G., Lasiak, T. A. & McGwynne, L. E. 1981. Sand beach energetics: an ecosystem approach towards a high energy interface. Estuarine, Coastal and Shelf Science 13, 11–25. McLachlan, A. & McGwynne, L. E. 1986. Do sandy beaches accumulate nitrogen? Marine Ecology Progress Series 34, 191–195. Messana, G., Chelazzi, G., Chelazzi, L., Ercolini, A., Ferrara, F., Messeri, P., Pardi, L. & Vannini, M. 1977. Researches on the coast of Somalia. The shore and the dune of Sar Uanle. 12. Physical environment: microclimate and soil. Monitore Zoologico Italiano (Nuova Serie Supplimento IX) 7, 147–181. Moggi, G. 1987. Researches on the coast of Somalia. The shore and the dune of Sar Uanle. 41. Vegetation. Monitore Zoologico Italiano (Nuova Serie Supplimento XXII) 17, 291–330. Moore, I. & Legner, E. F. 1973. Succession of the coleopterous fauna in wrack. The Wasmann Journal of Biology 31, 289–290. Moore, I. & Legner, E. F. 1976. Intertidal rove beetles (Coleoptera: Staphylinidae). In Marine insects, L. Cheng (ed.). Amsterdam: North-Holland, 521–551. Moore, P. G. 2002. Mammals in intertidal and maritime ecosystems: interactions, impacts and implications. Oceanography and Marine Biology: an Annual Review 40, 491–608. Moore, P. G. & Francis, H. C. 1985. Some observations on food and feeding of the supralittoral beachhopper Orchestia gammarellus (Pallas) (Crustacea: Amphipoda). Ophelia 24, 183–197. Moore, P. G., MacAlister, H. E. & Taylor, A. C. 1995. The environmental tolerances and behavioural ecology of the sub-Antarctic beach-hopper Orchestia scutigerula Dana (Crustacea: Amphipoda) from Husvik, South Georgia. Journal of Experimental Marine Biology and Ecology 189, 159–182. Moore, S. L., Gregorio, D., Carreon, M., Weisberg, S. B. & Leecaster, M. K. 2001. Composition and distribution of beach debris in Orange County, California. Marine Pollution Bulletin 42, 241–245. Morritt, D. 2001. The crab carrion-scavenging amphipod, Orchomene nanus in Lough Hyne, Co. Cork, Ireland. Journal of the Marine Biological Association of the United Kingdom 81, 1059–1060. Morton, B. 1990. The physiology and feeding behaviour of two marine scavenging gastropods in Hong Kong: the subtidal Babylonia lutosa (Lamarck) and the intertidal Nassarius festivus (Powys). The Journal of Molluscan Studies 56, 275–288. Morton, B. & Yuen, W. Y. 2000. Feeding behaviour and competition for carrion between two sympatric scavengers on a sandy shore in Hong Kong: the gastropod, Nassarius festivus (Powys) and the hermit crab, Diogenes edwardsii (De Haan). Journal of Experimental Marine Biology and Ecology 246, 1–29. Newell, R. C., Field, J. G. & Grifﬁths, C. L. 1982. Energy balance and signiﬁcance of microorganisms in a kelp bed community. Marine Ecology Progress Series 8, 103–113. Newell, S. Y., Fallon, R. D., Cal Rodriguez, R. M. & Groene, L. C. 1985. Inﬂuence of rain, tidal wetting and relative humidity on release of carbon dioxide by standing-dead salt marsh plants. Oecologia 68, 73–79. Obura, D. O. 2001. Kenya. Marine Pollution Bulletin 42, 1264–1278. Ochieng, C. A. & Erftemeijer, P. L. A. 1999. Accumulation of seagrass beach cast along the Kenyan coast: a quantitative assessment. Aquatic Botany 65, 221–238. Olin, R., Carlsson, B. & Stahre, B. 1995. The west coast of Sweden – the rubbish tip of the North Sea.

155

I. COLOMBINI & L. CHELAZZI

In Coastal and riverine litter: problems and effective solutions, R. Earll (ed.). Candle Cottage, Kempley, Gloucester: Marine Environmental Management and Training, 12–14. Pacheco-Ruíz, I. & Zertuche-González, J. A. 1996. The commercially valuable seaweeds of the Gulf of California. Botanica Marina 39, 201–206. Paul, V. J., Hay, M. E., Duffy, J. E., Fenical, W. & Gustafson, K. 1987. Chemical defence in the seaweed Ochtodes secundiramea (Montagne) Howe (Rhodophyta): effects of its monoterpenoid components upon diverse coral-reef herbivores. Journal of Experimental Marine Biology and Ecology 114, 249–260. Pennings, S. C., Carefoot, T. H., Zimmer, M., Danko, J. P. & Ziegler, A. 2000. Feeding preferences of supralittoral isopods and amphipods. Canadian Journal of Zoology 78, 1918–1929. Pennings, S. C. & Richards, C. L. 1998. Effects of wrack burial in salt-stressed habitats: Batis maritima in a southwest Atlantic salt marsh. Ecography 21, 630–638. Percival, S. M. & Evans, P. R. 1997. Brent geese Branta bernicla and Zostera; factors affecting the exploitation of a seasonally declining food source. Ibis 139, 121–128. Persson, L. E. 1999. Growth and reproduction in two brackish water populations of Orchestia gammarellus (Amphipoda: Talitridae) in the Baltic Sea. Journal of Crustacean Biology 19, 53–59. Philipp, R., Pond, K. & Rees, G. 1993. Litter and medical waste on bathing beaches in England and Wales. British Medical Journal 306, 1042 only. Philipp, R., Pond, K. & Rees, G. 1994. Medical wastes increasing on the coastline. British Medical Journal 309, 471 only. Phillips, D. S. & Arthur, W. 1994. Observations on the distribution of seaweed ﬂy larvae and other invertebrates within a wrack-bed. The Entomologist 113, 154–163. Phillips, D. S., Arthur, W., Leggett, M. & Day, T. H. 1995a. Differential use of seaweed species by British seaweed ﬂies, Coelopa spp. (Diptera: Coelopidae) with a description of the egg morphology of the two species. The Entomologist 114, 158–165. Phillips, D. S., Leggett, M., Wilcockson, R., Day, T. H. & Arthur, W. 1995b. Coexistence of competing species of seaweed ﬂies: the role of temperature. Ecological Entomology 20, 65–74. Polis, G. A., Anderson, W. B. & Holt, R. D. 1997. Towards an integration of landscape and food web ecology: the dynamics of spatially subsidised food webs. Annual Review of Ecology and Systematics 28, 289–316. Polis, G. A. & Hurd, S. D. 1995. Extraordinarily high spider densities on islands: ﬂow of energy from the marine to terrestrial food webs and the absence of predation. Proceedings of the National Academy of Science of the United States 92, 4382–4386. Polis, G. A. & Hurd, S. D. 1996a. Linking marine terrestrial food webs: allochthonous input from the ocean supports high secondary productivity on small islands and coastal land communities. The American Naturalist 147, 396–423. Polis, G. A. & Hurd, S. D. 1996b. Allochthonous input across habitats, subsidised consumers and apparent trophic cascades: examples from the ocean-land interface. In Food webs: integration of patterns and dynamics, G. A. Polis & K. O. Winemiller (eds). London: Chapman & Hall, 275–285. Prince, W. A., Tomlinson, K. W. & Hunt, J. N. 1968. The effects of artiﬁcial seaweed in promoting the build-up of beaches. Proceedings of the 11th International Conference on Coastal Engineering, London, American Society of Civil Engineering, 570–578. Pugh, P. J. A. & MacAlister, H. E. 1994. Acari of the supralittoral zone on sub-Antarctic South Georgia. Pedobiologia 38, 552–565. Rassmann, K. 1997. Evolutionary age of the Galapagos iguana predates the age of the present Galapago islands. Molecular Phylogenetics and Evolution 7, 158–172. Ray, G. C. & Hayden, B. P. 1992. Coastal zone ecotones. In Landscape boundaries: consequence of biotic diversity and ecological ﬂows, A. J. Hansen & F. di Castri (eds). New York: Springer, 403–420. Rees, G. & Pond, K. 1995. Marine litter monitoring programmes, a review of methods with special reference to national surveys. Marine Pollution Bulletin 30, 103–108.

156

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

Ribic, C. A. 1998. Use of indicator items to monitor marine debris on a New Jersey beach from 1991–1996. Marine Pollution Bulletin 36, 887–891. Ricci, S., Colombini, I., Fallaci, M., Scoccianti, C. & Chelazzi, L. 1998. Arthropods as bioindicators of the red fox foraging activity in a Mediterranean beach-dune system. Journal of Arid Environments 38, 335–348. Robecchi Bricchetti, L. 1903. Nel paese degli aromi. Diario di una esplorazione nell’Africa Orientale da Obbia ad Alula. Milano: Tipograﬁa Editrice L. F. Cogliati. Robertson, A. I. 1980. The structure and organisation of an eelgrass ﬁsh fauna. Oecologia 47, 76–82. Robertson, A. I. & Hansen, J. A. 1982. Decomposing seaweeds: a nuisance or a vital link in coastal food chains? CSIRO Division of Fisheries Research Report 1980–81, Australia, 75–83. Robertson, A. I. & Lenanton, R. C. J. 1984. Fish community structure and food chain dynamics in the surf-zone of sandy beaches: the role of detached macrophyte detritus. Journal of Experimental Marine Biology and Ecology 84, 265–283. Robertson, A. I. & Lucas, J. S. 1983. Food choice, feeding rates and the turnover of macrophyte biomass by a surf-zone inhabiting amphipod. Journal of Experimental Marine Biology and Ecology 72, 99–124. Robertson, A. I. & Mann, K. H. 1980. The role of isopods and amphipods in the initial fragmentation of eelgrass detritus in Nova Scotia, Canada. Marine Biology 59, 63–69. Rose, M. D. & Polis, G. A. 1998. The distribution and abundance of coyotes: the effects of allochthonous food subsidies from the sea. Ecology 79, 998–1007. Rosengarten, F. 1986. Coconut. Principes 30, 47–62. Ryan, P. G. 1987. The effects of ingested plastic on seabirds: correlations between plastic load and body condition. Environmental Pollution 46, 119–125. Ryan, P. G. & Moloney, C. L. 1990. Plastic and other artefacts on South African beaches: temporal trends in abundance and composition. South African Journal of Science 86, 450–452. Ryan, P. G. & Swanepoel, D. 1996. Cleaning beaches: sweeping the rubbish under the carpet. South African Journal of Science 92, 275–276. Schiebling, R. E. 1994. Molluscan grazing and macroalgal zonation on a rocky intertidal platform at Perth, Western Australia. Australian Journal of Ecology 19, 141–149. Schoenly, K. 1983. Microclimatic observations and diel activities of certain carrion arthropods in the Chihuahuan Desert. New York, Entomological Society 91, 342–347. Schoenly, K. & Reid, W. 1983. Community structure of carrion arthropods in the Chihuahuan Desert. Journal of Arid Environments 6, 253–263. Schoenly, K. & Reid, W. 1987. Dynamics of heterotrophic succession in carrion arthropod assemblages: discrete seres or a continuum of change? Oecologia 73, 192–202. Schulz, M. & Bamford, M. 1987. The hooded plover, A RAOU Conservation Statement. Royal Australasian Ornithologists Union Report 35, 1–11. Seiderer, L. J. & Newell, R. C. 1985. Relative signiﬁcance of phytoplankton, bacteria and plant detritus as carbon and nitrogen resources for the kelp bed ﬁlter-feeder Choromytilus meridionalis. Marine Ecology Progress Series 22, 127–139. Sheldon, J. W. 1991. Wild dogs: the natural history of the nondomestic Canidae. New York: Academic Press. Shibata, Y. 1993. On the staphylinid beetles living in the seashore. Insects and Nature 28, 23–27. Simpson, K. W. 1976. Shore ﬂies and brine ﬂies (Diptera: Ephydridae). In Marine insects, L. Cheng (ed.). Amsterdam: North-Holland, 465–495. Smith, B. D. & Foreman, R. E. 1984. An assessment of seaweed decomposition within a southern Strait of Georgia seaweed community. Marine Biology 84, 197–205. Soares, A. G., McLachlan, A. & Schlacher, T. A. 1996. Disturbance effect of stranded kelp on populations of the sandy beach bivalve Donax serra (Röding). Journal of Experimental Marine Biology and Ecology 205, 165–186. Soares, A. G., Schlacher, T. A. & McLachlan, A. 1997. Carbon and nitrogen exchange between sandy

157

I. COLOMBINI & L. CHELAZZI

beach clams (Donax serra) and kelp beds in the Benguela coastal upwelling region. Marine Biology 127, 657–664. Stenton-Dozey, J. M. E. & Grifﬁths, C. L. 1980. Growth, consumption and respiration by larvae of the kelp-ﬂy Fucellia capensis (Diptera: Anthyomyiidae). South African Journal of Zoology 15, 280–283. Stenton-Dozey, J. M. E. & Grifﬁths, C. L. 1983. The fauna associated with kelp stranded on a sandy beach. In Sandy beaches as ecosystems, A. McLachlan & T. Erasmus (eds). The Hague: Junk, 557–568. Stuart, V., Field, J. G. & Newell R. C. 1982. Evidence for absorption of kelp detritus by the ribbed mussel Aulacomya ater using a new 51Cr-labelled microsphere technique. Marine Ecology Progress Series 9, 263–271. Tarr, J. G., Grifﬁths, C. L. & Bally, R. 1985. The ecology of three sandy beaches on the Skeleton Coast, South West Africa. Madoqua 14, 78–90. Tarr, J. G. & Tarr, P. W. 1987. Seasonal abundance and the distribution of coastal birds on the northern Skeleton Coast, South West Africa, Namibia. Madoqua 15, 63–72. Taylor, J. D. & Lewis, M. S. 1970. The ﬂora, fauna and sediments of the marine grass beds of Mahé, Seychelles. Journal of Natural History 4, 199–220. Thayer, G. W., Bjorndal, K. A., Ogden, J. C., Williams, S. L. & Zieman, J. C. 1984. Role of large herbivores in seagrass communities. Estuaries 7, 351–376. Thornton, L. & Jackson, N. L. 1998. Spatial and temporal variations in debris accumulation and composition on a estuarine shoreline Cliffwood Beach, New Jersey, USA. Marine Pollution Bulletin 36, 705–711. Tinker, S. W. 1988. Whales of the world. Leiden: E. J. Brill, Bess Press. Tomás, J., Guitart, R., Mateo, R. & Raga, J. A. 2002. Marine debris ingestion in loggerhead sea turtles, Caretta caretta, from the western Mediterranean. Marine Pollution Bulletin 44, 211–216. Tomlinson, P. B. 1982. Helobiae (Alismatidae). In Anatomy of the moncotyledons, C. R. Metcalfe (ed.). Oxford: Clarendon Press. Tsipoura, N. & Burger, J. 1999. Shorebird diet during spring migration stopover on Delaware Bay. The Condor 101, 635–644. Uneputty, P. A. & Evans, S. M. 1997. Accumulation of beach litter on islands of the Pulau Seribu Archipelago, Indonesia. Marine Pollution Bulletin 34, 652–655. Valiela, I. 1974. Composition, food webs and population limitation in dung arthropod communities during invasion and succession. American Midland Naturalist 92, 370–385. Valiela, I., Koumjian, L., Swain, T., Teal, J. M. & Hobbie, J. E. 1979. Cinnamic acid inhibition of detritus feeding. Nature 280, 55–57. Valiela, I. & Reitsma, C. S. 1995. Disturbance of salt marsh vegetation by wrack mats in Great Sippewissett Marsh. Oecologia 102, 106–112. Valiela, I., Teal, J. M. & Allen, S. D. 1985. Decomposition in salt marsh ecosystems: the phases and the major factors affecting disappearance of the above ground organic matter. Journal of Experimental Marine Biology and Ecology 89, 29–54. van Pelt, T. I. & Piatt, J. F. 1995. Deposition and persistence of beach cast seabird carcasses. Marine Pollution Bulletin 30, 794–802. Velander, K. A. & Mocogni, M. 1998. Maritime litter and sewage contamination at Cramond beach Edinburgh – a comparative study. Marine Pollution Bulletin 36, 385–389. Velander, K. A. & Mocogni, M. 1999 Beach litter sampling strategies: is there a “best” method. Marine Pollution Bulletin 38, 1134–1140. Velimirov, B., Field, J. G., Grifﬁths, C. L. & Zoutendyk, P. 1977. The ecology of kelp bed communities in the Benguela upwelling system. Analysis of biomass and spatial distribution. Helgoländer Wissenschaftliche Meeresuntersuchungen 30, 495–518. Venables, B. J. 1981. Aspects of the population biology of a Venezuelan beach amphipod, Talorches-

158

MARINE ALLOCHTHONOUS INPUT ON SANDY BEACH COMMUNITIES

tia margaritae (Talitridae), including estimates of biomass and daily production and respiration rates. Crustaceana 41, 271–285. Walker, D. I., Masini, R. J. & Paling, E. I. 1988. Comparison of the annual production and nutrient status of the primary producers in a shallow limestone reef system (Rottnest Island), Western Australia. Proceedings of Australian Marine Science Association Silver Jubilee Conference, Australia, 1–5. Walker, T. R., Reid, K., Arnould, J. P. Y. & Croxall, J. P. 1997. Marine debris surveys at Bird Island, South Georgia 1990–1995. Marine Pollution Bulletin 34, 61–65. Wells, F. E. & Keesing, J. K. 1989. Reproduction and feeding in the abalone Haliotis roei Gray. Australian Journal of Marine and Freshwater Research 40, 187–197. Weslawski, J. M., Stanek, A., Siewert, A. & Beer, N. 2000. The sandhopper (Talitrus saltator, Montagu 1808) on the Polish Baltic coast. Is it a victim of increased tourism? Oceanological Studies 24, 77–87. Williams, A. T. & Tudor, D. T. 2001a. Temporal trends in litter dynamics at a pebble pocket beach. Journal of Coastal Research 17, 137–145. Williams, A. T. & Tudor, D. T. 2001b. Litter burial and exhumation: spatial and temporal distribution on a cobble pocket beach. Marine Pollution Bulletin 32, 1031–1039. Willoughby, N. G., Sangkoyo, H. & Lakaseru, B. O. 1997. Beach litter: an increasing and changing problem for Indonesia. Marine Pollution Bulletin 34, 469–478. Wilson, D. S. & Knollenberg, W. G. 1984. Food discrimination and ovarian development in burying beetles (Coleoptera: Silphidae: Nicrophorus). Annals of the Entomological Society of America 77, 165–170. Zabel, C. J. & Taggart, S. J. 1989. Shift in red fox, Vulpes vulpes, mating system associated with El Niño in the Bering Sea. Animal Behaviour 38, 830–838. Zoli, C. 1927. Oltre Giuba. Notizie raccolte dal Commissario Generale nel primo anno di occupazione italiana (1925–1926). Roma: Sindacato Italiano Arti Graﬁche.

159

Oceanography and Marine Biology: an Annual Review 2003, 41, 161–236 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

THE EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES LAURA AIROLDI Centro Interdipartimentale di Ricerca per le Scienze Ambientali di Ravenna & Dipartimento di Biologia Evoluzionistica Sperimentale, University of Bologna, Italy & Marine Biological Association of the United Kingdom, Plymouth, UK e-mail: [email protected] Correspondence address: Scienze Ambientali, Universita di Bologna, via dell’Agricoltura 5, I-48100 Italy

Abstract Sedimentation is a widespread and increasing process on most rocky coasts. The literature on its effects is reviewed and support is found for the general conclusion that sedimentation is an important ecological factor for hard bottom organisms. Sediments deeply affect the composition, structure and dynamics of rocky coast assemblages, and increased sediment load as a consequence of anthropogenic activities can be a threat to their diversity and functioning. Sediments that accumulate on rocky substrata are important agents of stress and disturbance. They can cause burial, scour and profound modiﬁcations to the characteristics of the bottom surface, and interact with other important physical and biological processes. The effects of sedimentation are complex, because they involve both direct outcomes on settlement, recruitment, growth or survival of individual species and indirect outcomes through mediation of competitive and/or predator–prey interactions. Not all species and assemblages are equally affected by sedimentation and responses vary over space and time, depending on the characteristics of the depositional environment, life histories of species and the stage of development of individuals and assemblages, and in relation to variable physical factors, including hydrodynamics, light intensity and bottom topography. Recent studies have much improved our ability to detect and understand the effects of sedimentation on rocky coast assemblages. However, little is still known about the underlying mechanisms. Overall, our present ability to make generalisations and predictions is limited by a paucity of quantitative and experimental research, and by the scant attention devoted to measuring the regime of perturbation by sediments and responses of organisms at relevant spatial and temporal scales. Predicting the magnitude of the effects that different sedimentation regimes have on rocky coast organisms and the critical levels above which detrimental effects become manifest remains a key issue for the ecology of rocky coasts and a challenge for future studies.

Introduction Context and aim of the study Over the past few decades, there have been increases in water turbidity and sediment deposition in coastal areas. Sedimentation has occurred at unprecedented rates all over the world as a consequence of anthropogenic activities, such as deforestation, dredging, industrial and domestic discharges, construction activities and land reclamation. Such increase of sediment 161

LAURA AIROLDI

loads has been recognised as a major threat to marine biodiversity at a global scale (United Nations Environmental Programme 1995). Changes in sedimentation have been dramatic for some coral reefs, where accelerated deposition caused by unsustainable land practices and dredging resulted in rapid shifts in species composition and abundance, and eventually in irreversible deterioration and loss of coral reefs and associated ﬁshery resources (e.g. Johannes 1975, Amesbury 1982, Cortés & Risk 1985, van Katwijk et al. 1993, Hunter & Evans 1995, Chou 1996, McClanahan & Obura 1997, White et al. 2000; but see Larcombe & Woolfe 1999). The increasing concern about degradation of coastal habitats as a result of anthropogenic increase of sediment loads is reﬂected in the amount of research that has been directed in recent years towards these problems in both temperate and tropical regions. Several reviews have appeared that discuss the effects of sedimentation in different habitats, including coral reefs (Dodge & Szmant-Froelich 1985, Rogers 1990, Richmond 1993), mangroves (Ellison 1998), seagrasses (Vermaat et al. 1997), freshwater systems (Barko et al. 1991, Henley et al. 2000), and estuaries (Ryan 1991). As the development of many human activities is likely to result in the release of suspended sediments in coastal areas and/or in changes to the regime of sedimentation, understanding the effects of sediments on coastal assemblages and predicting threshold levels of impact for different habitats are fundamental to identifying sustainable management strategies. Rocky coasts are amongst the habitats potentially most sensitive to increased sediment loads, as excessive deposition of sediments may cause dramatic alterations in the characteristics of the benthos. Nevertheless, the impact of sediment loads on rocky coast assemblages has rarely been examined directly. Despite the scarcity of direct observations and experiments, there is an extensive body of literature that indicates the ecological role of sediments in rocky coasts is of major importance. The purpose of this review is to assimilate much of this literature, and attempt to address a number of questions: How does sedimentation affect rocky coast organisms? How do individual species and assemblages on rocky coasts respond to changes in the characteristics of the regime of sedimentation? Are there physiological and morphological adaptations that enhance tolerance of species in rocky coasts to disturbance by sediment? Are there critical levels of change in the regime of sedimentation that will irreversibly damage the biodiversity and/or functioning of rocky coast assemblages?

The review The review is organised into three major sections. In the ﬁrst, the variability and major sources of sediments to rocky coasts are described, the different mechanisms by which sedimentation can affect rocky coast assemblages discussed, and problems in measurement and comparison of sedimentation rates outlined. This section is not meant to be inclusive, as it is a vast subject that would require a dedicated review; rather it is intended to provide the reader with a few examples that highlight the complexity of the effects that sediments can have on rocky coast assemblages, the importance of quantifying the spatial and temporal scales of disturbance by sediments, and the necessity of improving comparison of sedimentation rates across different habitats and regions. In the second section, the available information concerning the effects of sedimentation on biota of rocky coasts at the levels of individuals species and assemblages is critically documented. It is anticipated that, although sedimentation is invoked as an important factor for rocky coast organisms in many papers, direct research to quantify these effects and the 162

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

underlying mechanisms is relatively scarce, and much of the available information is scattered, fragmental and often incidental. It is not the purpose of this section to cover all sources that mention sediments as a potentially important ecological factor for rocky coast organisms. Rather an attempt is made to cover the pertinent literature that is most frequently cited in international journals, with particular attention to the direct experimental evidence that supports or refutes the hypothesis that sediment accumulation inﬂuences the structure of rocky coast assemblages. Much of this literature deals with assemblages on rocky coasts that are naturally subject to the presence of sediments and, in the absence of suitable data, examples have also been drawn from other habitats. The information has been synthesised in tables, which are an essential part of this section. Some observations are also available from assemblages subjected to enhanced sediments loads as a consequence of human activities. However, a thorough coverage of this literature has not been attempted because in most cases discharge of sediments was one of many stresses undergone by the assemblage, and effects due to sedimentation were not separated from effects caused by potentially toxic organic and/or chemical pollutants. In the last section, information is used to discuss: (a) the knowledge of the role of sedimentation in inﬂuencing the structure and functioning of rocky coast assemblages, including direct and indirect effects; (b) the present limited abilities to understand mechanisms of response of individual species and assemblages and to predict critical levels of disturbance by sediments; (c) which factors most hinder generalisations and predictions, including insufﬁcient description of the regime of perturbation by sediments and scarce consideration of the spatial and temporal context; and (d) which questions need to be addressed by future studies. Organic particulate matter may constitute an important source of food for suspension and detritus feeding benthic animals. On rocky coasts, however, there is a general consensus that high loads of suspended and settling inorganic particles represent a factor of stress and disturbance for both hard-bottom algae and animals. This review focuses on the physical effects of burial and scour by settled inert inorganic particles, because sediments that accumulate to the bottom, as a consequence of either natural or anthropogenic processes, are generally mostly inorganic. Related effects due to the presence of suspended organic and inorganic particles or to the presence of pollutants are touched upon only lightly, in view of the availability of other reviews related to these topics (e.g. Kinne 1971, Moore 1977, Pearson & Rosenberg 1978, Capuzzo et al. 1985, Wotton 1990, Walker & Kendrick 1998). Sediments accumulated on rocky substrata, or trapped within mats of algae, mussels or other invertebrates, will affect the composition and diversity of assemblages on rocky coasts also in terms of providing habitat to a variety of soft-bottom organisms (e.g. macroinfauna and meiofauna). Although the importance of this factor is recognised, effects of sediments on soft-bottom species have not been covered, as this would deserve a review of its own. The review devotes particular attention to assemblages dominated by macroalgae, reﬂecting the author’s background experience and the extensive nature and global distribution of these habitats. Nomenclature of macroalgae was updated following Guiry & Dhonncha (2002). Coverage of “grey literature” has been kept to a minimum, to ensure the review is of accessible literature. Reference to work before the 1970s is limited, and readers are directed to the extensive coverage of Moore (1977).

163

LAURA AIROLDI

Sediment load to rocky coasts Some sediments occur on most rocky coasts, to the point that some authors have recently recognised the artiﬁciality of the traditional separation between rocky and sandy shores (Bally et al. 1984, Jørgensen & Gulliksen 2001). Sediments are added to rocky coasts by many natural and human-related processes (Fig. 1), and are redistributed as a function mainly of sediment characteristics, hydrodynamic conditions, bottom proﬁle, and biological factors (Moore 1972, 1973a, Hiscock 1983). These factors operate over a range of scales. Thus, while sediment deposition can be relatively predictable at large spatial and temporal scales, depending on the source and magnitude of sediment loads, at small scales, relevant to individual organisms, patterns may be highly heterogeneous and unpredictable (Trowbridge 1996, Airoldi & Virgilio 1998). This section, does not attempt to give detailed information on these topics, but rather to provide some baseline information relevant to understanding the possible effects of sedimentation on rocky coast organisms, and to analyse problems related to the measurement and comparison of sedimentation rates across different habitats.

Sources Natural processes Sediments on rocky coasts (in the forms of clay, silt, detritus, or more frequently sand) derive from a variety of natural processes. The principal sources include discharges by rivers, erosion of cliffs, and resuspension and transport of sediments. In some coastal areas, detritus from benthic and pelagic organisms and atmospheric transport can also be important

Figure 1

Schematic diagram of inputs and transport of inorganic particles to rocky coasts.

164

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

sources of sediments (e.g. Moore 1977 and references therein, Fornos et al. 1992), but there do not appear to be speciﬁc examples relative to rocky coasts. River catchment of land-derived materials from natural soil erosion and runoff is a major input of sediments to coastal areas (Moore 1973a, French 1997). The rate at which terrestrial sediments are washed into rivers is a function of many environmental factors, such as intensity of rainfall, type of soil, and cover by vegetation. The major supplies of river-derived sediments to rocky coasts occur during ﬂoods associated with storms; as a consequence, for short time periods (hours to several days) sediment loads can be orders of magnitude higher than normal. Branch et al. (1990), for example, reported that, during a massive ﬂood that occurred for several days in March 1988, the South African Orange River transported up to 55 t silt s1 out of the mouth to the surrounding rocky coastal areas. The sediment did not deposit immediately on the rocky shore but was slowly trapped by algae, and after 2 months amounts of silt as high as over 1000 g m2 were accumulated at midshore levels. While sediment inundations by ﬂooding may be considered extreme and rare events, erosion and runoff of cliffs by rain, wind, ocean waves and ice are relatively frequent local sources of sediments to rocky coasts (French 1997). Vogt & Schramm (1991), for example, reported that in Kiel Bay (Germany) each year about 75 000 m3 of rock were eroded by wave action and washed into the bay from about 35 km out of 90 km of cliff; about 35% to 40% of such material, in particular coarse sand, gravel and stones, were deposited near the shore line at the base of the cliff, resulting in the formation of sandy ﬂoors and sandbanks which in some places have signiﬁcantly reduced the amount of rocky substrata available for settlement of rocky coast organisms. Sediment traps placed along vertical cliffs have allowed some quantiﬁcation and analysis of the composition of detritus rolling down the cliffs (Evans et al. 1980, Gulliksen 1982, Bavestrello et al. 1995). Along rocky cliffs in the Ligurian Sea (Italy), for example, ﬂuxes of sediment were closely related to wave action and rain, with peaks up to more than 100 g m2 day1 during the spring and autumn (Bavestrello et al. 1991). Sediments were mainly composed of inorganic particles from the cliff but debris from animals and plants was also present. In some areas, such as the Paciﬁc coasts of North America, or the coasts of New Zealand, landslides are a relatively common natural phenomenon that can cause local inundation of rocky shores by sediments. Examples of the extent and impact of landslides are discussed by Shaffer & Parks (1994), Konar & Roberts (1996), Slattery & Bockus (1997), and Smith & Witman (1999). The contribution of sediments from these events, however, although locally important and persistent, has been considered small when compared with the annual contribution of sediments from other sources (Shaffer & Parks 1994). The most frequent natural source of sediments to rocky coasts is re-suspension and transport of sediments from nearby soft-bottom areas (Storlazzi & Field 2000). Periodical inundations of sand by coastal currents or the action of storms are a very common feature of rocky coasts throughout the world, including California (Littler et al. 1983, Stewart 1983), Oregon (Markham 1973, D’Antonio 1986, Menge et al. 1994, Trowbridge 1996), New Hampshire (Daly & Mathieson 1977), Maine (Moring 1996), North Carolina (Renaud et al. 1996, 1997), British Columbia (Mathieson 1982), Mexico (Pineda & Escofet 1989), Ireland (Cotton 1912), Egypt (Aleem 1993), Morocco (Birje et al. 1996), Ghana (Towsend & Lawson 1972, Evans et al. 1993), Namibia (Engledow & Bolton 1994), and South Africa (Stephenson 1943). The degree of inundation by sand can be extremely variable, and Bally et al. (1984) have proposed a scheme of classiﬁcation of “mixed shores” in relation to the spectrum of relative abundances of rock and sediments. 165

LAURA AIROLDI

Human-related processes There are many reports of major increases of water turbidity and sediment load that occur in rocky reefs as a consequence of human activities. Sediments may originate directly from industrial and domestic discharges (Boney 1978, Eagle et al. 1979, Schroeter et al. 1993, Gorostiaga & Díez 1996, Raimondi et al. 1997, Kim et al. 1998), mining activities (Castilla & Nealler 1978, Ellis 1988, Bernier et al. 1997, Hyslop et al. 1997, Fariña & Castilla 2001, Pulfrich et al., in press), construction of roads, bridges, tunnels, harbours and residential developments (Meinesz et al. 1991, Bach et al. 1993, Iannuzzi et al. 1996, MacDonald et al. 1997, Turner et al. 1997), dredging operations (Eagle et al. 1979), replenishment of beaches (Guidetti 2001), and aquaculture (Holmer et al. 2001). Most often, however, man’s activities affect supply of sediments to rocky coasts in indirect ways, such as by accelerating natural soil erosion, by modifying the coastline and river catchments thus changing hydrodynamic and bottom characteristics, or by altering the abundance of important species that control the distribution of sediments (Seapy & Littler 1982, Meinesz et al. 1991, French 1997, SaizSalinas & Urkiaga-Alberdi 1999, Gillanders & Kingsford 2002). Paucity of long-term quantitative data sometimes makes it difﬁcult to quantify the trends, and unequivocally attribute the causes to human activities (Lumb 1990). However, there are many lines of evidence that acceleration of natural soil erosion in relation to changes in land use both inland and along the coast is one of the most likely causes of enhanced inputs of sediments to coastal areas in both tropical and temperate regions (Cortés & Risk 1985, Rogers 1990, van Katwijk et al. 1993, McClanahan & Obura 1997, MacDonald et al. 1997). The greatest impacts are felt when forests are cleared for timber, agriculture, or urban developments. Deforestation increases soil erosion, water runoff, and occurrence of landslides, with dramatic results for the amount of sediments washed into rivers and brought downstream to the coast. It has been estimated that, in tropical woodlands, forest clearance and cultivation have increased natural losses of soil from 3 t ha1 yr1 up to 54–334 t ha1 yr1 (French 1997). Similarly, MacDonald et al. (1997) estimated that erosion from urban land uses and development of roads caused at least a 4-fold increase in sediment yields and unprecedented sedimentation rates in coastal areas around St John, US Virgin Islands. Pronounced increase of water turbidity and sediment loads in rocky coasts have also been observed in relation to ﬁres of coastal vegetation (Airoldi et al. 1996), which in some regions are most often accidentally or deliberately started by humans. Sediments deriving from these different activities may vary in chemical composition and grain size, may contain various organic and inorganic pollutants, and may range from slurries with a high water content to highly compacted sediments (e.g. Eagle et al. 1979). Sometimes the discharges occur over limited space and/or timescales, and sediments are dispersed relatively quickly by natural processes. However, more often the discharges are long lasting, resulting in persistent accumulations of sediments. Along the coasts of England, for example, considerable discharges of sediments from mining and industrial activities occurred continuously at some sites for about 95 yr (Eagle et al. 1979): although in the 1990s dumping had been substantially reduced, colliery waste was still washed up on to rocky shores near former dumping sites (Hyslop et al. 1997). Similarly, asbestos excavations at Canari Mine (Corsica, France) from 1948–65 greatly modiﬁed the natural rock escarpments, ultimately determining the appearance of artiﬁcial shores up to 300 m in width at the bottom of existing steep cliffs over a distance of more than 5 km (Bernier et al. 1997). Schroeter et al. (1993) reported important increase of sediment ﬂuxes and water turbidity as far as 1.4 km 166

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

from the diffusers of a coastal nuclear power plant in southern California: muddy sediments accumulated on the bottom and became armoured with coarser materials, accreting over time up to covers above 40%. In recent years, at a few locations there has been reported a decrease of sediment load, which has been attributed to a recession of industries and mining activities and to better treatment of domestic discharges (Gorostiaga & Díez 1996, Hyslop et al. 1997). Although this is still a limited process, and although available data do not allow us to draw any conclusions about the possible timings for recovery of rocky coast assemblages, this seems an auspicious trend.

Spatial and temporal variability of sediment deposition Identifying the main pathways of dispersal and accumulation of sediments is important for predicting the impact of discharges from natural and human-related processes. Overall, the distribution of sediments on rocky coasts is inﬂuenced by the characteristics of the sediments themselves and hydrodynamic conditions (Hiscock 1983). Whereas larger particles tend to settle quickly (e.g. close to the source points), the very ﬁne particles can be kept in suspension for long periods by water turbulence, and thereby be transported over long distances by prevailing currents (Moore 1972, Capuzzo et al. 1985, Bach et al. 1993). In a report on the effects of the disposal of solid wastes off the northeast coast of England, for example, Eagle et al. (1979) showed how the area of impact of sediments released by different industrial developments extended far away from the dumping sites: sediments were transported and redistributed as a function of particle characteristics (i.e. composition, size, shape and density), tidal and wave-induced currents, and bottom characteristics. Patterns of deposition and movements of sediments can be variable over space and time (Airoldi & Cinelli 1996a). For example, accumulation of sediments, especially the ﬁnest fractions, is in general most pronounced at sheltered locations (Lilly et al. 1953, Mathieson 1982, Hiscock 1983, D’Antonio 1986), whereas at exposed locations, sediments, if any, tend to be coarser, and generally persist accumulated in small crevices and depressions or trapped by local assemblages (Gotelli 1988, Airoldi & Virgilio 1998). Movements of sediments have greater effects in terms of abrasion at exposed than sheltered sites (Mathieson 1982, Hiscock 1983, Engledow & Bolton 1994), also due to the larger size of particles at the former sites. Sedimentation is generally greatest following strong rainfall and storms, due to increased runoff and resuspension of sediments (Bavestrello et al. 1995, Airoldi et al. 1996). Similarly, on some rocky shores, inundations by sand can assume seasonal, neap-spring tide, or even daily cycles (e.g. Daly & Mathieson 1977, Littler et al. 1983, Stewart 1983, D’Antonio 1986). Along the coasts of British Columbia, for example, Mathieson (1982) and Markham & Newroth (1972) recorded seasonal ﬂuctuations of the levels of sand of 1 m to 1.5 m: generally the largest deposition occurred in late spring, and the sand persisted until the ﬁrst autumn storms, which then removed the sand completely, sometimes in less than 24 h. Superimposed on these large-scale, relatively predictable patterns, sediments may be redistributed within each shore contingent on the microtopography of the bottom and local proﬁles of ﬂow speed. This process can result in highly heterogeneous spatial and temporal patterns of distribution of sediments within each shore. Sediments, for example, tend to be more abundant on horizontal than sloping surfaces (Whorff et al. 1995, Jørgensen & Gulliksen 2001). Trowbridge (1996) observed that on intertidal shores in Oregon, patterns of 167

LAURA AIROLDI

Figure 2 Example of heterogeneous spatial and temporal patterns of distribution of sediments on rocky coasts (modiﬁed from Airoldi & Virgilio 1998, published with permission). Data are average ( 1 SD) dry weights of sediment (total amount, including coarse and ﬁne fractions) deposited above and trapped into ﬁlamentous turf-forming algae at three nearby sites (about 100 m apart) on subtidal reefs south of Livorno, Italy.

accumulation of sand were more regular on planar benches than on heterogeneous substrata with surge channels. Furthermore, at a local scale, relevant to individual algal thalli or invertebrates, patterns of sand burial (frequency, duration, depth) were unpredictable, and shores that appeared to be buried for long periods were actually locally uncovered for short periods, affecting the extent of sand burial as well as conditions during burial. Similar observations were reported for subtidal rocky reefs in the Mediterranean Sea (Airoldi & Virgilio 1998): accumulation of sediments signiﬁcantly differed among nearby sites 100 m apart, and was highly patchy at a scale of metres within each site (Fig. 2). The importance of sediments as a major source of spatial and temporal heterogeneity for rocky coast organisms has been fully recognised only recently (Daly & Mathieson 1977, Littler et al. 1983, McQuaid & Dower 1990, Trowbridge 1996, Airoldi 1998). In particular, Airoldi & Virgilio (1998) and Airoldi (2000b) have shown how the responses of rocky coast assemblages to stress and disturbance by sediments may vary with changes in spatial and temporal scales. Thus the perception of the effects of sediments on rocky coast assemblages may be inﬂuenced by the spatial and temporal extent of the study. So far, however, little work has been undertaken to identify the relevant spatial and temporal scales of interactions between sedimentation and rocky coast assemblages.

Control by biological factors Biological factors may control the distribution of sediments on rocky coasts. For example, several authors (e.g. Scofﬁn 1970, Stewart 1983, Airoldi & Virgilio 1998) have shown how turf-forming algal assemblages can bind and stabilise sediments even on exposed coasts, main168

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Figure 3 Example of biological control of the distribution of sediments on rocky shores (redrawn from Bertness 1984, published with permission). The experimental removal of the snail Littorina littorea from a sheltered New England rocky beach resulted in rapid accumulation of sediments, and the development of foliose algae, which further accelerated sedimentation. Data are average sediment dry weights ( 1 SD) measured at two different times in each of unmanipulated controls, caged controls, and three L. littorea removal treatments. The box insert shows the relationship between weight of accumulated sediment and per cent cover of foliose algae (Enteromorpha intestinalis and Ulva lactuca) measured in one Littorina littorea removal treatment.

taining relatively constant accumulations of sediments despite marked temporal variations in sediment load (see p. 187). In some areas, kelp canopies have been reported to accelerate sediment deposition and prevent sediments being washed away (e.g. Moore 1972, 1973a, Eckman et al. 1989), while in others there have been reports of little to no accumulation of sediments under kelp canopies because of whiplash by fronds (e.g. Kennelly 1989, Melville & Connell 2001). In both cases, kelps have been shown to exert an important control on sediment dynamics and Estes & Palmisano (1974) have suggested that, by controlling the abundance of kelps, sea urchins may indirectly control the sedimentation regime of many subtidal habitats. Branch et al. (1990) observed that accumulation of sediments transported to rocky shores following the Orange River ﬂoods was not immediate but started only after the disappearance of patellid limpets and the consequent development of algal beds that trapped silt. Bertness (1984) showed that some herbivorous snails may prevent accumulations of sediments either directly, by bulldozing surfaces, or indirectly by removing algal ﬁlms that trap sediments (Fig. 3). Bertness speculated that biological factors can mediate sedimentation rate and sediment binding sufﬁciently to dictate whether a habitat is primarily soft- or hard-bottomed. Although the effects of biological activity in controlling sediment transport, deposition and accrual rates have long been recognised as an important phenomenon in soft-bottom marine and lentic environments (e.g. Scofﬁn 1970, Fonseca & Fisher 1986, Power 1990, Pringle & Blake 1994, French 1997, Gacia & Duarte 2001), the role of biological factors in inﬂuencing the presence and distribution of sediments on rocky coasts has been so far largely neglected. 169

LAURA AIROLDI

Mechanisms by which sediments affect organisms on rocky coasts Because of the variety of possible sources, and of the synergistic effects of environmental and biological processes in controlling the distribution and dynamics of sediment particles, the nature and amount of sediments that are present on rocky coasts is highly variable in space and time. Sediments can be clay, silt, sand or detritus, may have variable composition and grain size distributions, and may or may not include pollutants. They may occur as a thin stratum, or form deposits centimetres to tens of centimetres thick, and may or may not turn into compact, impermeable layers. They may accumulate locally or be resuspended and transported above the substrata, depending on local hydrodynamic and topographic conditions. Furthermore, high rates of sedimentation are generally accompanied by high levels of turbidity from suspended sediments, and may often be associated with elevated inputs of nutrients or lower levels of salinity, as in the case of river ﬂoods (Gillanders & Kingsford 2002). This heterogeneity results in a variety of possible effects on rocky coast organisms, which are often difﬁcult to separate from each other or from the concomitant effects of other environmental factors. Indeed, while in recent years there has been an increasing number of studies aimed at testing experimentally the effects of sediments on rocky coast organisms and assemblages, very few attempts have been made to try to clarify and separate the mechanisms of action of sediments (but see Devinny & Volse 1978, Marshall & McQuaid 1989, Kendrick 1991, Airoldi 1998, Chapman & Fletcher 2002), and identify the relevant spatial and temporal scales of impact of these different and interacting processes (Airoldi & Virgilio 1998). With these problems in mind, and leaving out the possible interacting effects related to other environmental factors often covarying with sediments (e.g. hydrodynamics, turbidity, salinity, organic and inorganic pollutants), at least three major mechanisms by which inert inorganic sediments may directly affect rocky coast organisms have been postulated by a number of authors (e.g. Lilly et al. 1953, Daly & Mathieson 1977, Devinny & Volse 1978, Littler et al. 1983, Turner 1985, D’Antonio 1986, Kendrick 1991, Airoldi et al. 1996, Airoldi 1998, Chapman & Fletcher 2002). Namely, (1)

(2) (3)

Burial/smothering, which may involve reduced availability of light, oxygen, nutrients, or accumulation of hydrogen sulphide and metabolic waste products, thus resulting in major changes in the characteristics of the chemical microenvironment. Scour/abrasion by moving sediments, that may damage and remove whole organisms or their parts. Changes in the physical characteristics of the bottom surface, which occur as a consequence of the replacement of stable hard substrata with unstable particles, and can result in a loss of habitat suitable for settlement.

Probably these mechanisms often occur together. For example, on exposed rocky coasts, sediment burial and scour are often combined, owing to high turbulent ﬂow near the bottom (Devinny & Volse 1978, Mathieson 1982, Airoldi et al. 1996) but the effects they have on benthic communities are different and should be distinguished (Taylor & Littler 1982, D’Antonio 1986, Kendrick 1991, Airoldi 1998). Overall, the characteristics of sediment particles (e.g. grain size, shape, density, mineral and chemical composition), the extent, degree, location, frequency and duration of sediment burial, and the strength of water motion all 170

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

contribute to determine the regime of perturbation by sediments, which can range from a large-scale, sub-lethal, chronicle stress, to an abrupt severe disturbance that locally disrupts the assemblage by removing organisms (Airoldi et al. 1996, Airoldi 1998). In this context, “stress” and “disturbance” are deﬁned as in Grime (1977): stress refers to mechanisms that preclude or limit the growth of the assemblage, whereas disturbance refers to mechanisms causing the partial or total removal of the assemblage. Given that, with notable exceptions, ecologists have frequently referred to the effects of “sedimentation” (or analogous terms) ambiguously, in the absence of data on the regime of perturbation by sediments (see next section), and without explicit distinction of the mechanisms involved, the term “perturbation” is used here to refer to the complex range of effects related to presence of sediments, and more speciﬁc deﬁnitions (e.g. burial, scour, stress, disturbance) are used whenever explicit information is available.

Problems in measurement and comparison of sedimentation It is often difﬁcult to interpret and compare results from studies on the effects of sedimentation on rocky coast assemblages because either no or very limited information is given about the magnitude, characteristics and spatial and temporal variability of sedimentation regime, or different methods are used to quantify sedimentation, sometimes without an explicit consideration of which aspect of “sedimentation” is measured (i.e. sediment deposition, accumulation or movement, scour, or turbidity). This difﬁculty makes it rather arbitrary to identify which levels of sedimentation should be considered as “high” or “low” for rocky coast organisms or, just as importantly, which effects of sedimentation should be considered as “stressful” or “non stressful”. The majority of the ﬁeld studies included in a later section of the review (p. 174) reported no or limited information on sediment abundance or reported qualitative subjective visual estimates (e.g. “little”, “moderate”, “heavy”, “very heavy”) without any reference to actual levels of sediments (Fig. 4). Less than 50% of the papers reported quantitative estimates of sediment abundance, either measured directly or through reference to previously published work. Unexpectedly, sometimes information on abundance of sediments was lacking even from experimental work in the laboratory, where no information was given about the type of sediment treatment applied (see “Laboratory experiments”, p. 196), and lack of quantitative information on the effectiveness of manipulations of sediment is a major shortcoming of most ﬁeld experiments (see “Field experiments”, p. 207). Less than 10% of the papers reported data on sediment characteristics (e.g. type, size or composition of sediment particles). Even less information was available on temporal variability of sedimentation, and only very few papers reported quantitative information on spatial variability of sediment deposition or acknowledged the potential importance of this factor.

Measurement of sediment deposition and accumulation Methods used to quantify sedimentation to rocky coasts have been very variable. The use of different methods is, in part, related to the characteristics of the habitat where researchers operate. Trowbridge (1996), for example, suggested that in areas where sand ﬂuctuations are greater than 1 m, and the underlying rocks have high topographic relief, quantifying 171

LAURA AIROLDI

No quantitative estimates

69

Estimates by using sediment traps

22

Estimates of sediment thickness

10

Estimates of sediment weight

9

Estimates of sediment cover

7

Estimates of turbidity

6

Information on sediment discharges

5

Estimates of suspended sediments

4

Composition and size of sediments

11

Figure 4 Estimates of sediment load reported in selected studies on the effects of sedimentation on rocky coast assemblages. Bars represent number of studies.

sediment cover, depth or mass may be difﬁcult to apply or not informative. As an alternative, she used qualitative estimates of the degree of burial of the surveyed algae (i.e. thalli not buried, some thalli uncovered and some buried, or thalli totally buried). However, sometimes, the choice seems dictated more by personal preference rather than by objective constraints. Rocky intertidal habitats In rocky intertidal habitats sedimentation is generally estimated in terms of accumulated sediments. The principal methods used include: (1) (2) (3) (4) (5) (6)

estimates of distribution and per cent surface cover of sediments by using visual, photographic or video techniques (e.g. Stewart 1982, Littler et al. 1983, Renaud et al. 1996), estimates of sediment depth by direct measurements or by using reference marks (e.g. Markham 1973, Daly & Mathieson 1977, Mathieson 1982, Moring 1996, Renaud et al. 1996), estimates of sediment mass per unit area (e.g. Emerson & Zedler 1978, Stewart 1983, Branch et al. 1990, Engledow & Bolton 1994, Whorff et al. 1995), and estimates of sediments accumulated over time on panels or trapping surfaces (Bertness 1984). Sometimes sedimentation has also been estimated indirectly by using measurements of sediment suspended in waters close to the shore, by ﬁltering known volumes of water (e.g. Little & Smith 1980, Hyslop et al. 1997, Fariña & Castilla 2001), and estimates of water turbidity (Mettam 1994, Iannuzzi et al. 1996). It should be noted, however, that high levels of suspended sediments do not necessarily result in high levels of sediment accumulation on the bottom.

Rocky subtidal habitats The principal methodologies used to quantify sedimentation in rocky subtidal habitats include: 172

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

(1)

(2) (3)

the estimation of sediment deposition by using sediment traps, which collect settling sediments over a given time (e.g. Moore 1972, Gulliksen 1982, Deysher & Dean 1986, Bavestrello et al. 1991, 1995, Kendrick 1991, Schroeter et al. 1993, Airoldi et al. 1996, Airoldi & Virgilio 1998, Maughan 2001), measurement of sediment cover and/or thickness, or mapping of sediments (e.g. Kennelly 1983, Gotelli 1988, Renaud et al. 1997, Slattery & Bockus 1997), and measurements of suspended solids or water turbidity (Saiz-Salinas & UrkiagaAlberdi 1999).

Sediment traps can have variable designs, which can affect the performance of estimates of sedimentation rates in moving waters (Bloesch & Burns 1980). In recent years, laboratory and in situ experiments have shown that stable, vertically suspended, smooth cylinders, with an inner diameter of 45 mm and an aspect ratio of 3 give the most reliable results (Blomqvist & Håkanson 1981, Butman et al. 1986), which has lead to a major uniformity in the design of traps used. However, there is still variability in the way traps are positioned and in the methods used to quantify sediment trapped. In some cases, due to experimental or habitat constraints, traps have been replaced by panels or other small trapping structures, which have been generally used for relative comparisons of sedimentation rates rather than to estimate absolute ﬂuxes of sediments (e.g. Eckman et al. 1989, Airoldi & Cinelli 1997). Sediment traps are invaluable as research tools for documenting gross sediment inputs to rocky coasts. Caution has been recommended, however, as results are greatly affected by resuspension and movement of sediments, particularly in highly energetic habitats (Moore 1972, Blomqvist & Håkanson 1981, Gulliksen 1982, Bavestrello et al. 1995, Airoldi et al. 1996, Lund-Hansen et al. 1997). Thus, the downward ﬂux of particles does not necessarily equal the rate of accumulation of sediments to the bottom and the sessile biota (Gardner 1980). Ideally, studies should incorporate direct sampling of sediments accumulated on rocky substrata and organisms (Purcell 1996) but, probably due to the constraints of working underwater, estimates of sediment mass per unit area have been scarce in subtidal rocky habitats (Kendrick 1991, Herrnkind et al. 1988, Airoldi & Virgilio 1998).

Measurement of scour In highly energetic habitats, movement of sediments, particularly the coarse fractions, can have important effects on organisms in terms of scour. The intensity of scour by sediments, however, has generally been inferred indirectly from observations on either movements of sediments and/or damage to and removal of organisms (e.g. Daly & Mathieson 1977, Robles 1982, Littler et al. 1983, D’Antonio 1986, McGuinness 1987a, Menge et al. 1994, Airoldi et al. 1996, Trowbridge 1996, Airoldi 1998, Underwood 1998). Craik (1980) proposed a method for measuring scour in intertidal areas based on the rate of dissolution of cement blocks anchored to the substratum; this method, however, could not separate the scour due to movement of blocks over the substratum by wave action from scour due to abrasion from suspended sediments. Estimates of scour have also been obtained by using tiles painted with a thin coat of colour that is removed by contact with particles (Thompson et al., unpubl. data), similar to those used to measure scour by kelp fronds on understorey species in kelp forests (e.g. Kennelly 1989).

173

LAURA AIROLDI

Effects of sedimentation Historically, studies on the effects of sedimentation on rocky coast organisms have focused on reports of lists of species from areas naturally affected by sediments, or more rarely on examinations of life histories and structural adaptations of single species or groups of species. A consistent body of literature has reported observations on species that tend to bind and trap sediments but quantitative data have been scarce. In the last two decades quantitative observations on responses of individual species or assemblages to sedimentation have become relatively more frequent, probably in response to the increasing concern for the trend of enhanced sediment loads to coastal areas. These studies have sometimes been supported by laboratory experiments. However, it is only recently that systematic attempts have been made to investigate experimentally, both in the ﬁeld and in the laboratory, the causal mechanisms by which sediments may affect rocky coast organisms. In this review an attempt is made to summarise the major ﬁndings from this considerable body of literature, and analyse whether observations stimulating the hypothesis that sediments have an important ecological role on rocky coasts are supported by experimental evidence. Most of the papers that are included in the present section of the review have been published in international journals indexed in Current Contents. Many represent papers that reported qualitative, and sometimes fortuitous, observations (Fig. 5). This information included records of species that appeared to be more frequent or less frequent in habitats characterised by high levels of sediments, or observations done during studies not speciﬁcally aimed at analysing the effects of sedimentation on rocky coast organism. In most cases, these papers invoked sedimentation as a possible explanation for patterns observed in the assemblage, but provided little or no supporting data. About 35% of the papers reported quantitative observations of changes in the abundance of rocky coast organisms that were attributed to variations in the regime of sedimentation. The reliability of these studies was variable, and the papers that reported rigorous observations repeated over time or with an adequate level of spatial replication were few. In many cases, effects of sediments were difﬁcult to separate from the possible inﬂuence of other environmental factors covarying with

Figure 5 Percentage of qualitative and quantitative observations and/or laboratory and ﬁeld experiments in studies on the effects of sedimentation on rocky coast assemblages.

174

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

sediments (e.g. variations in the hydrodynamic regime, changes in bottom topography and depth, natural seasonal patterns) or from factors often associated with high sediment loads (e.g. turbidity, presence of organic or chemical pollutants). In this group, also included are “natural experiments”, in which sediments were not manipulated directly, but effects were studied by repeating experiments or transplanting organisms across habitats “impacted” or “non-impacted” by sediments. Papers reporting results from experiments in which sedimentation was directly manipulated were in a minority: most work was done in the laboratory, while experiments in the ﬁeld made up 10% of studies.

Qualitative and quantitative observations The effects most frequently attributed to the presence of sediments on rocky coasts are summarised in Figure 6. Effects were most generally based on comparisons among areas with “low” and “high” sediment load, or on changes observed over time and presumably related to variations in sediment load. Interpretation of results as responses to “high” or “low” levels of sedimentation, however, was limited in many studies by lack of quantitative information on the characteristics and variability of the regime of sedimentation (see p. 171), by scarce consideration of spatial and temporal issues, and by confounding effects because the distribution of sediments on rocky coasts generally covaries with other important physical and biological factors (see pp. 167 and 168). Furthermore, it is important to note that identifying relationships between the distribution of sediments and species on rocky coasts, while important for the formulation of predictive hypotheses, cannot be considered an evidence of causality.

Changes in species composition and distribution

57

Inhibition of settlement and recruitment

27

Decline/mortality/removal of species

26

Prevalence of distinctive morphological, or life-history traits

25

Reduced species diversity/monopolization of space

23

Weakened competition and/or predation

22

Associations turf-sediments

20

No, limited or non persistent effects

10 8

Inhibition of growth or fertility Enhanced species diversity Enhanced development, growth or survival Changes in species morphology

7 3 2

Figure 6 Effects or lack of effects most frequently attributed to the presence of sediments on rocky coast organisms and/or assemblages. More than one effect was often attributed to sediments in any study. Bars represent number of studies.

175

LAURA AIROLDI

Changes in species composition and abundance Early indication that sediments are an important factor inﬂuencing the composition of rocky coast assemblages can be found in past taxonomic accounts of the species occurring in rocky coast habitats naturally affected by sediments (e.g. Cotton 1912, Hoyt 1920, Kitching et al. 1934, Rees 1935, Doty 1947a,b, Mathieson & Fralick 1972, Mathieson 1979, Mathieson 1982). These habitats appeared to be peculiar systems, characterised by distinctive assemblages of plants and animals; taxonomic records included species that appeared to be restricted to sediment affected areas, species that seemed to occur at both sediment affected and unaffected areas, and species that occurred more frequently at unaffected areas but were occasionally present also in areas with sediments. More recent quantitative works on assemblages from sediment impacted rocky coasts have conﬁrmed the observation that the composition and distribution of species often seems to be closely related to spatial and/or temporal changes in sediment load. Daly & Mathieson (1977), for example, reported that the composition, abundance and distribution of intertidal species in New Hampshire were related to ﬂuctuations of sand levels: assemblages at shores most affected by sediments were characterised by the dominance of opportunistic species (e.g. Enteromorpha spp. and Ulva lactuca), and perennial “psammophytic” species (e.g. Ahnfeltia plicata and Sphacelaria radicans), and by the notable absence of species common on nearby rocky shores and considered intolerant to sediments (e.g. Ascophyllum nodosum). Furthermore, the lower limits of distribution of some species (e.g. Mytilus edulis, Semibalanus balanoides (as Balanus balanoides), and Porphyra umbilicalis) appeared to be related to the historical sequences of sand inundations in the area, as they approximated the zone of highest summer elevations of sand. Littler et al. (1983) described how patterns of species abundance on intertidal rocky shores in southern California were related to the relative degree of sand stress on different portions of the study site: opportunistic macrophytes (e.g. Chaetomorpha linum, Cladophora columbiana, Ulva lobata and Enteromorpha intestinalis) and highly reproductive macroinvertebrates (e.g. Tetraclita rubescens, Chthamalus ﬁssus/dalli and Phragmatopoma californica) dominated areas routinely buried by sand; long-lived species (e.g. Mytilus californianus, Haliotis cracherodii and Lottia gigantea) dominated areas where rock contours provided a refuge from sand deposition, and sand tolerant species (e.g. Anthopleura elegantissima and Phyllospadix scouleri) dominated areas with greatest sediment deposition. Gorostiaga et al. (1998) reported that sedimentation, probably related to wave exposure, was the environmental factor that best matched the gradient in the distribution of sublittoral benthic algae along the eastern Basque coast. Along a gradient of increasing levels of sediments (characterised by using a semi-quantitative scale of cover), they observed changes in the relative abundance of some species; for example, the abundance of Gelidium corneum (as G. sesquipedale) and Mesophyllum lichenoides was negatively related to sediments, whereas Halopitys incurvus and Chondracanthus acicularis were most abundant in areas with high cover of sediments. The general pattern was a decrease of vertical structure due to the loss of canopy-forming species, as also reported from other areas impacted by sediments (Seapy & Littler 1982, Vogt & Schramm 1991, Airoldi et al. 1995, Airoldi 1998, Eriksson et al. 2002). Konar & Roberts (1996) reported differences in the composition of species between areas close to or distant from plumes originated by landslides. Patterns differed among sites, however, and the only consistent trend was a greater abundance of brown algae at the areas distant from the sediment plumes. Renaud et al. (1996, 1997) observed that the distribution and composition of macroalgae on 176

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

subtidal rocky reefs affected by sediments were related to ﬂuctuations in sediment characteristics and depth; cover of macroalgae was consistently greatest in areas with low cover of sediments, and a notable increase in the abundance of macroalgae was observed after a storm removed sediments from some areas. Saiz-Salinas & Urdangarin (1994) described a gradient of disappearance of macroalgae and increased abundance of opportunistic ﬁlterfeeding animals along the outer part of Bilbao harbour and discussed how estuarine sedimentation appeared to be the main environmental factor responsible for such gradient. Similarly, several authors have described patterns of zonation of species, or differences among assemblages on surfaces of different inclinations, and identiﬁed sedimentation as one of the environmental factors most closely related to those patterns (e.g. Chapman 1943, Stephenson 1943, Lilly et al. 1953, Lewis 1964, Pérès & Picard 1964, Clarke & Neushul 1967, Moore 1973b, Norton et al. 1977, Little & Smith 1980, Farrow et al. 1983, CastricFey 1988, Ballesteros 1992, Brattström 1992, Santos 1993, Mettam 1994, Johansson et al. 1998, Gabriele et al. 1999, Saiz-Salinas & Urkiaga-Alberdi 1999, Pedersén & Snoeijs 2001). Caution is needed, however, because in many studies distribution of sediments was correlated with other important environmental factors, such as wave action, depth or salinity. Low density of grazers and concomitant dominance of turf-forming and/or opportunistic foliose algae have been observed on several rocky coasts affected by sediments (e.g. Stewart 1989, Airoldi et al. 1995, Airoldi 1998, Pulfrich et al., in press, see pp. 187 and 189). Branch et al. (1990), for example, reported that following the ﬂooding of the Orange River in South Africa, entirely different assemblages developed in areas most affected by abnormal dilution of water and high load of sediments and the shore changed from being dominated by patellid limpets to being dominated by opportunistic foliar algae. Among others, Daly & Mathieson (1977), Emerson & Zedler (1978), Robles (1982), Seapy & Littler (1982), Littler et al. (1983), Stewart (1983), D’Antonio (1986), Aleem (1993), Evans et al. (1993), Birje et al. (1996) and Kim et al. (1998) described changes in the composition and abundance of species following occasional, seasonal or long-term ﬂuctuations in the levels of sediments. On a protected New England rocky beach, Bertness (1984) showed how accumulation of sediments, due to the removal of the snail Littorina littorea and consequent development of foliose algae, ultimately increased the abundance of organisms characteristic of soft-bottom habitats, such as polychaetes, tubiculous amphipods, mud crabs, and mud snails, and decreased the success of organisms characteristic of hard-bottom habitats, such as barnacles and encrusting algae. Changes in the abundance and composition of rocky coast assemblages have been frequently reported following enhanced sediment loads as a consequence of human activities (see p. 193). Finally, accumulation of sediments within mats of algae, mussels, or other invertebrates is known to be related to the abundance and diversity of species typically associated with soft bottoms, including macroinfauna and meiofauna (e.g. Gibbons 1988 and references therein, Grahame & Hanna 1989). A few authors have also reported limited or non-persistent effects of sedimentation on the composition of rocky coast assemblages. Carballo et al. (1996), for example, found that the composition and distribution of sponges in subtidal hard-bottom habitats in Algeciras Bay, Spain, were not apparently related to rates of sedimentation. Baynes (1999) reported that differences in the composition of assemblages between horizontal and vertical surfaces on rocky reefs in southern Gulf of California were not related to differences in sediment deposition. Shaffer & Parks (1994) showed that a landslide had immediate effects on the abundance of algae in kelp beds adjacent to the slide area, but such effects did not persist over time, and differences between affected and unaffected areas were no longer evident after few 177

LAURA AIROLDI

months. Similarly, Moring (1996) reported immediate responses of assemblages to severe disturbance by sediments following a hurricane but by the following year excess sand was removed from the area by winter turbulence, and assemblages recovered to their pre-disturbance appearance. Many observations and experiments on the effects of sedimentation on rocky coast assemblages are carried out over very short times. However, the short-term nature of the effects reported by Shaffer & Parks (1994) and by Moring (1996), as well as the evidence from observations and experiments that effects of sediments vary over time in relation to other environmental and biological factors (e.g. Airoldi & Cinelli 1997, Airoldi 1998), suggest that timing is a critical and overlooked factor in studies on the impact of sedimentation on rocky coast assemblages.

Responses of individual species: evidence of direct and indirect effects Although it appears evident that sediments may inﬂuence the species composition of assemblages on rocky coasts, by limiting the abundance of some species and favouring the development of others, in most cases the underlying mechanisms are not known. With few exceptions, the above reported relationships between sediments and species compositional changes have been attributed to direct effects on individual species that appeared to respond “negatively”, “positively”, or “indifferently” to their presence. Thus, observations have been focused on hard-bottom species that seemed to tolerate or even be enhanced by sediments (see pp. 179 and 187), or that seemed to undergo negative effects on their recruitment, growth, or survival (see p. 189). However, the lack of a clear understanding of the mechanisms by which sedimentation affects rocky coast organisms is reﬂected by the notably contrasting effects sometimes attributed to sediments. Species, or groups of hard-bottom species, that have been ranked from sensitive to tolerant to sediments are numerous, including, among others, species belonging to the genera Ulva, Enteromorpha and Gelidium, mussels, encrusting coralline algae, and sponges. These observations will not be analysed case by case, because quantitative data supporting evidence of either tolerance or intolerance are limited, and because scarcity of data on sediment loads makes it difﬁcult to compare and interpret results as response to “high” or “low” levels of sedimentation (see p. 171). The case of encrusting coralline algae is taken as an example because there are lines of evidence that such apparently contrasting observations are not only related to different susceptibility among congeneric species, or to variations in the regime of perturbation by sediments, but are also connected with the complex direct and indirect effects of sedimentation. Responses of encrusting coralline algae to presence of sediments are controversial. Some observations suggest that encrusting coralline algae may be negatively affected by sediments (e.g. Ayling 1981, Kennelly 1991, Moring 1996, Gorostiaga et al. 1998, Maughan 2001, Melville & Connell 2001, Pulfrich et al., in press, S. D. Connell, pers. comm.), whereas others suggest that encrusting coralline algae are often abundant, or even dominant, in a variety of sediment-impacted habitats (Littler 1973, Kendrick 1991, Saiz-Salinas & Urdangarin 1994, Airoldi et al. 1995, Falace & Bressan 1995, Konar & Roberts 1996, Airoldi 2000a). Such different responses are certainly in part related to the variable ecology of this vast and diversiﬁed group of species (e.g. Dethier 1994). Furthermore, observed effects may depend on the local characteristics of sedimentation regime. For example, recent experiments by Matsunaga et al. (1999) have shown that the presence of forest-derived, humic substances in sediments may inhibit the germination of encrusting coralline spores. There is 178

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

evidence, however, that the contrasting responses observed for encrusting coralline algae may be related also to the complex indirect effects of sedimentation. Steneck et al. (1997), for example, suggested that most coralline algae are sensitive to sedimentation, although the encrusting coralline alga Neogoniolithon strictum formed unique algal ridges in the Bahamas, in environments characterised by low wave energy, high rates of sedimentation and low rates of herbivory. The authors discussed how this encrusting coralline alga tolerated sediment burial possibly as a result of anatomical adaptations (abundant multiple cell fusions and unbranched morphology). Even this species, however, required relatively sediment-free, hard substrata for successful germination and growth. The authors suggested that the success of N. strictum in such sediment-affected environments may be inﬂuenced by the low rates of herbivory observed in those habitats. Kendrick (1991) demonstrated that recruitment and growth of coralline crusts were enhanced by treatments simulating scour, whereas crusts were overgrown and outcompeted by turf-forming algae in treatments simulating erosion and accretion. Kendrick suggested that positive effects of scour were related indirectly to negative effects on the abundance of overgrowing algae (i.e. burial and abrasion by sand may provide the cleaning effects normally provided by grazers, and described as important to maintain encrusting coralline algae). Similar hypotheses have been suggested by other authors (e.g. Stewart 1989). Airoldi (2000a), however, demonstrated that some encrusting corallines, which occurred abundantly on sediment-stressed subtidal reefs, were more tolerant to overgrowth by turf-forming algae. It was hypothesised that crusts could beneﬁt from being overgrown by turf through protection from abrasion by sediments, which seemed to negatively affect recruitment and growth of crusts. Furthermore, overgrowth by turf and trapped sediment could relieve crusts from competition for primary substrata with erect algae. These observations suggest that the effects of sediments on rocky coast organisms may be complex, probably involving not only direct effects (e.g. from smothering and scour) on individual species but also indirect effects mediated by competitive or predator/prey outcomes. Indeed, results of recent ﬁeld experiments have conﬁrmed the potential indirect effects of sedimentation (see p. 207), the importance of which has long been overlooked. Consideration of indirect effects is fundamental to understanding responses of species to variations in sediment load.

Morphological, physiological and life-history attributes of “psammophytic” “sand-tolerant” species Rocky coasts affected by sediments have always been viewed as extreme environments for hard-bottom species, characterised by highly stressful physical conditions. Therefore, morphological and physiological attributes of hard-bottom species occurring at and often dominating these habitats have long been the object of observation and discussion, particularly in the case of long-lived species that were able to persist and maintain spatial dominance from year to year in the presence of sediments. These species have often been indicated as “psammophytic” or “sand loving”, which, as Littler et al. 1983 pointed out, implies that they are somehow directly favoured (e.g. in terms of enhanced growth and/or reproduction) by sediments. However, in recent years, observations and experiments have indicated that most “psammophytic” species are rather “sand-tolerant” species: they may be negatively affected by sediments but not as severely as other species, and the costs imposed by living in 179

LAURA AIROLDI

sediment-stressed habitats are probably compensated for by indirect advantages in terms of reduced competition and/or predation (Taylor & Littler 1982, Littler et al. 1983, Turner 1985, D’Antonio 1986, Kendrick 1991, Trowbridge 1992, Airoldi & Cinelli 1997, Airoldi & Virgilio 1998; see also pp. 196 and 207). The role of tolerance as an important mechanism inﬂuencing community structure relative to “negative” and “positive” interactions has been highlighted by the results of recent studies, which have suggested that the prevalence of space monopolising forms, such as algal turfs and crusts, may be related to their abilities to withstand a variety of physical and biological challenges (Airoldi 1998, 2000a). This emphasises the need for attention to life-history traits that enhance tolerance of species. Littler et al. (1983) observed that rocky coasts impacted by sediments are colonised primarily by three groups of species with different life-history traits: (1) (2) (3)

long-lived species (either “psammophytic” or “sand-tolerant”) that seem to be capable of adjusting to stresses imposed by the presence of sediments, opportunistic species, that rapidly recolonise space following mortalities caused by burial and scour, and migratory species, that move in and out of the habitat depending on the level of burial by sediments.

Rocky coasts affected by sediments are also often dominated by another group of species (Airoldi 1998, Airoldi & Virgilio 1998 and references therein), consisting of (4)

species that bind and trap sediments, and appear to tolerate burial and scour. Many of these species have life histories intermediate to long-lived and opportunistic species, and show a strong association with sediments. Because of its importance particular attention is dedicated to this group in the following section.

Very little experimental work has been done to test what attributes of a species result in differentiating tolerance or susceptibility to the presence of sediments (see Pineda & Escofet 1989). Nevertheless, observations from a variety of sediment impacted rocky habitats consistently suggest the prevalence of certain morphological, reproductive and physiological attributes (Table 1). These include: • • • • • • •

the regeneration of upright portions from remnant bases tolerant to sediment burial and scour; opportunistic cycles of reproduction and growth or the capacity to propagate vegetatively; tough and wiry thalli or bodies; growth, reproductive cycles, and/or migrating behaviours synchronised with ﬂuctuations of sediments; apical meristems that maintain dividing cells above sediment; erect morphology that prevents settlement of sediment; and physiological adaptations to withstand darkness, anaerobic conditions and high hydrogen sulphide concentrations.

For example, Markham (1972) observed that Laminaria sinclairii, which grows in habitats affected by sediments, has a longer stipe and narrower blades than its congener L. longipes which grows in sediment-free habitats, and has distinctive deciduous blades. Markham & 180

Morphological, physiological, life-history traits

Terete branched thalli of tough construction.

Peak growth and reproduction synchronised with seasonal ﬂuctuations of sand, deciduous blades, regeneration of upright fronds from perennial holdfast resistant to damage from sand burial.

Peak growth synchronised with seasonal ﬂuctuations of sand, long stipe and narrow deciduous blades, reproduction by vegetative propagation, regeneration of upright fronds from perennial holdfast resistant to damage from sand burial.

Thallus plasticity, considerable growth of rhizoids, presumed reduction of respiratory rates and tolerance to hydrogen sulphide.

Extensive rhizome system, tough

Species

Numerous algae, including Rhodothamniella ﬂoridula (as Rhodochorton ﬂoridulum), Vaucheria velutina (as V. thuretii), Polyides rotundus, and Gracilaria gracilis (as G. confervoides)

Phaeostrophion irregulare

Laminaria sinclairii

Zonaria farlowii

181

Halimeda spp., Udotea spp.,

Species belonging to these genera

Intertidal and subtidal rocky habitats, often in areas subject to seasonal burial by sand. Alive thalli were collected from 150–100 mm thick layers of black anoxic sand. California, USA.

Normally restricted to low intertidal rocky habitats, moderately to fully exposed to wave action, and subject to seasonal ﬂuctuations of sand (up to 2 m). Paciﬁc coast of North America.

Normally restricted to low intertidal rocky habitats subject to seasonal ﬂuctuations of sand (up to 2 m). Paciﬁc coast of North America.

Restricted to or very abundant on intertidal and shallow subtidal rocks covered by sand or in sandy tide pools. Clare Island (Ireland).

Habitat/Location

Scofﬁn 1970, Williams et al. 1985,

Dahl 1971

Markham 1968, 1972, 1973

Mathieson 1967, 1982, Turner 1983, 1985

Cotton 1912

References

Table 1 Morphological, physiological or life-history traits that have been suggested to confer tolerance to presence of sediments in species most frequent on rocky coasts with high levels of sediments (species sometimes deﬁned as “psammophytic”).

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Morphological, physiological, life-history traits

thalli.

Peak growth synchronised with seasonal ﬂuctuations of sand, clump growth, terete branched thalli of tough construction due to thick cortical layers.

Tough and wiry thalli, regeneration of upright fronds from perennial basal holdfasts, reproduction by vegetative propagation, incomplete alternation of generations.

Reproduction by vegetative propagation, ability to elongate above the sediment surface, low oxygen demand, migratory behaviour (in solitary specimens).

Large size, tough thalli, rhizomatous growth, regeneration of upright portions from remnant bases, opportunistic life histories, migratory behaviour.

Penicillus spp., Caulerpa spp.

Ahnfeltiopsis linearis (as Gymnogongrus linearis), Ahnfeltia plicata, A. concinna

Numerous algae, including Ahnfeltia plicata and Sphacelaria radicans

Anthopleura elegantissima

Numerous plants and animals, including Phyllospadix torreyi and Tegula funebralis

continued

Species

Table 1

182

Daly & Mathieson 1977

Markham & Newroth 1972

Littler et al. 1988, Piazzi et al. 1997

References

Abundant to dominant in intertidal rocky habitats periodically inundated by sediments. California, USA.

Littler et al. 1983

Common in rocky intertidal habitats Taylor & Littler 1982, Littler et al. subject to seasonal burial by sand. 1983, Pineda & Escofet 1989 California, USA, and Baja California, Mexico.

Dominant in intertidal rocky habitats subject to irregular ﬂuctuations of sand (up to 1 m). New Hampshire, USA.

Normally restricted to low intertidal rocky habitats, moderately to fully exposed to wave action, and subject to seasonal ﬂuctuations of sand (up to 2 m). Paciﬁc coast of North America.

occur in both rocky and sandy habitats in a variety of tropical and temperate environments.

Habitat/Location

LAURA AIROLDI

183

Reproduction by vegetative propagation, regeneration of upright fronds from perennial basal holdfast resistant to burial.

Small size, low sexual reproduction, reproduction by vegetative propagation.

Neorhodomela larix (as Rhodomela larix)

Sargassum sinicola

Dominant in shallow subtidal rocky habitats, also in areas close to the mouth of streams and subject to high sand deposition. Southern Gulf of California, Mexico.

Frequent on moderately exposed, horizontal, intertidal rocky habitats, particularly in areas subject to seasonal ﬂuctuations of sand (up to 60 cm). Absent from vertical surfaces or areas with high urchin densities. Paciﬁc coast of North America.

Adaptations were discussed for specimens in shallow subtidal sheltered bays, but the species can also occur in rocky habitats with sand. Chile.

Regeneration of upright fronds from underground thallus system resistant to burial.

Habitat/Location

Gracilariopsis lemaneiformis (as Gracilaria lemaneiformis)

Morphological, physiological, life-history traits Abundant to dominant in intertidal rocky habitats periodically inundated by sediments. California, USA.

continued

Numerous algae, including Apical meristems that maintain Pterocladia capillacea, Dictyota sp., dividing cells above sediment, Sphacelaria rigidula (as S. furcigera) regeneration of upright portions from remnant bases resistant to burial, peak growth and reproduction synchronised with seasonal ﬂuctuations of sand, tough and calciﬁed thalli, vegetative propagation.

Species

Table 1

Espinoza & Rodriguez 1987

D’Antonio 1986

Santelices et al. 1984

Stewart 1983

References EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Morphological, physiological, life-history traits

Regeneration of upright branches from perennial creeping stolons.

Low oxygen demand, use of anaerobic pathways.

Regeneration of upright fronds from perennial basal crust resistant to burial and scour, calciﬁed thalli, apical meristems, large reproductive outputs, lateral vegetative propagation, slow growth rates.

Reproduction by vegetative propagation, regeneration of upright fronds from perennial basal parts resistant to burial and scour, tough thalli, opportunistic life histories.

No obvious demographic, phenological, anatomical or morphological features that indicate resistance to sand burial. The alga grows adherent to rocky surfaces, forming discrete crustose pads or

Several species of Dictyotales

Siphonaria capensis

Corallina spp. (C. pinnatifolia and C. vancouveriensis)

Several algae including Womersleyella setacea (as Polysiphonia setacea), Halimeda tuna and Dictyota dichotoma

Codium setchellii

continued

Species

Table 1

Marshall & McQuaid 1989

King & Farrant 1987

References

184

Airoldi et al. 1995, Airoldi & Cinelli 1997, Airoldi 1998

Present at low densities in low Trowbridge 1996 intertidal rocky habitats subject to seasonal up to nearly continuous burial by sand (up to 1 m). Growth was possible at non-sandy sites, if shielded from herbivory. Buried thalli

Abundant to dominant in subtidal rocky habitats, also in areas characterised by high deposition (2 g m2 d1 to 178 g m2 d1) and movement of ﬁne sediments. Mediterranean Sea.

Dominant on mid intertidal shores Stewart 1989 periodically disturbed by sand burial and scour. Southern California, USA.

Frequent on sand-free, intertidal rocky areas, but can extend its distribution to sediment-affected areas. South Africa.

Abundant in shallow water, sheltered rocky habitats, covered permanently by a layer of moving sand. Sydney Harbour, Australia.

Habitat/Location

LAURA AIROLDI

continued

Sargassum microphyllum

Fucus vesiculosus forma mytili

Species

Table 1

Regeneration from perennial basal holdfast.

Adaptation to sediments were discussed for intertidal specimens in coral reefs, but species belonging to this genus are also frequent in rocky habitats.

On mussel beds in soft-bottom habitats. The form growing in soft-bottom habitats is so different from that in hard-bottom habitats that it was originally characterised as a distinct species (F. mytili). Wadden Sea, Germany.

did not appear to be subject to anoxic conditions. Paciﬁc coast of North America.

irregular cushions. The author suggested that persistence may be due to slow growth and long lifespan.

Lack of holdfast, lack of air vesicles, lack of sexual reproduction, propagation by vegetative reproduction by means of drifting fragments of adult thalli.

Habitat/Location

Morphological, physiological, life-history traits

Umar et al. 1998

Albrecht 1998

References EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

185

LAURA AIROLDI

Newroth (1972) observed that Ahnfeltiopsis linearis (as Gymnogongrus linearis), which can survive sediment burial over 6 months, has thicker cortical layers and a more terete thallus than the congener G. platyphyllus, which cannot survive in sediment affected areas. Solitary animals have been found to be less susceptible to sediments than colonial species (reviewed in Jackson 1977). Invertebrates with an erect morphology have been suggested to be less susceptible to sediments than prostrate forms (e.g. Saiz-Salinas & Urdangarin 1994, Irving & Connell 2002). Marshall & McQuaid (1989) demonstrated in the laboratory that the pulmonate limpet Siphonaria capensis, which extends its distribution to sediment-affected areas, has speciﬁc physiological adaptations to reduced oxygen tensions that presumably allow it to survive burial for considerably longer periods than does the prosobranch Patella granularis which is restricted to sediment-free areas. Albrecht (1998) reported that Fucus vesiculosus (forma mytili) growing on intertidal mussel beds on soft bottoms in the Wadden Sea differed in growth habit, morphology and reproductive ability from F. vesiculosus growing on mussel beds on rocky shores (Fig. 7). The former was characterised by lack of holdfast, lack of air vesicles, and especially by lack of sexual reproduction: reproduction only occurred vegetatively, by means of drifting fragments of adult thalli. In a detailed description of assemblages on intertidal rocky shores periodically inundated by sand at Bound Rock, California, Daly & Mathieson (1977) reported that several species, including F. vesiculosus, Chondrus crispus and Mastocarpus stellatus (as Gigartina stellata), exhibited extensive holdfast regeneration. Many species of algae that occur on both sheltered soft bottoms and rocky coasts, such as species belonging to the genera Halimeda, Caulerpa, Penicillus and Udotea, typically have tough thalli and develop extensive systems of rhizomes (Scofﬁn 1970, Williams et al. 1985, Littler et al. 1988, Airoldi & Cinelli 1997, Piazzi et al. 1997). Furthermore, the ability to reproduce vegetatively seems to be one of the features most consistently reported for species in sediment-impacted areas (see Table 1). Preva-

Figure 7 Growth habit and morphology of fertile Fucus vesiculosus colonising mussels on rocky substrata (left) and sterile F. vesiculosus forma mytili colonising mussels on mud and sand ﬂats in the Wadden Sea (modiﬁed from Albrecht 1998, published with permission).

186

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

lence of vegetative propagation in these habitats may be related to the fact that this form of reproduction removes dependence upon spore attachment to buried, unstable substrata, and allows a rapid recovery from damage (Airoldi 1998). Indeed, observations and experiments suggest that vegetative propagating stages are less vulnerable to a variety of physical and biological factors than are sexually reproductive stages, and recover very quickly after most common disturbances, especially when damage to organisms is patchy (see Airoldi 2000b and references therein). In contrast to these apparent adaptations, Trowbridge (1996) discussed how Codium setchellii, which forms low-density populations on sand-inﬂuenced rocky benches in Oregon, does not exhibit obvious anatomical or morphological traits that suggest tolerance to sand. She hypothesised that the alga persists in sand-stressed habitats because of its slow growth and long life-span, and stressed the need for more studies to understand better the adaptation of species to environmental conditions during burial.

Species that trap and bind sediments Rocky intertidal and subtidal coasts affected by sediments appear to be often dominated by species which have a morphology that tends to accumulate and trap sediments. These species may include sessile invertebrates, such as the mud ectoprot Cryptoarachnidium argilla, which has been observed forming extensive mats up to 3 mm thick that entrap ﬁne sand and silt (Palmer-Zwahlen & Aseltine 1994), or mussels which are known to accumulate large amounts of sediments (e.g. Albrecht 1998 and references therein). However, the most common and abundant group of species is represented by densely packed, small macroalgae which form mats that are frequently referred to as turfs (Airoldi 1998). The composition and structure of turfs may be extremely variable, because they may be formed by ﬁlamentous, calcareous-articulated or corticated-terete algae (Airoldi 2001). Despite their variable morphology, however, they all tend to be associated with trapped sediments. The dense network of ramiﬁcations within turfs forms, in fact, a cohesive surface layer that tends to alter the ﬂow microenvironment and entrap particles moving on the bottom (Neumann et al. 1970, Scofﬁn 1970, Carpenter & Williams 1993). Cotton (1912), for example, reported that on intertidal shores at Clare Island (Ireland), Rhodothamniella ﬂoridula (as Rhodochorton ﬂoridulum) was the most important species of a group of ﬁnely branched, upright growing algae which accumulated and retained sand. Boney (1980) observed stratiﬁed mineral accumulations of sand grains and particles of coal dust within small turfs formed by the ﬁlamentous alga Rhodochorton purpureum (as Audouinella purpurea) along the coasts of Scotland. Sousa et al. (1981) reported the presence of layers of anoxic sediments within turfs formed by perennial red algae, mainly Chondracanthus canaliculatus (as Gigartina canaliculata), Laurencia paciﬁca and Gastroclonium coulteri, on intertidal shores of southern California. Kennelly (1989) reported that a conglomerate of silt, microscopic ﬁlamentous algae and microinvertebrates built up in patches experimentally cleared of kelps. Among others, turfs formed by various species of red (e.g. Corallina spp., Polysiphonia spp., Ceramium spp., Laurencia spp., Gelidium spp., Acrochaetium spp.), brown (e.g. Giffordia spp.) and ephemeral green algae (e.g. Ulva spp., Enteromorpha spp., Cladophora spp.) have been frequently reported to trap sediments in rocky intertidal and shallow subtidal habitats in various geographical areas (Scofﬁn 1970, Towsend & Lawson 1972, Emerson & Zedler 1978, Ayling 1981, Stewart 1983, 1989, Herrnkind et al. 1988, 187

LAURA AIROLDI

Grahame & Hanna 1989, Branch et al. 1990, Kendrick 1991, Airoldi et al. 1995, Whorff et al. 1995, Piazzi & Cinelli 2001). Sediment is considered to be a structural constituent of algal turfs (Stewart 1983, Kendrick 1991, Airoldi & Virgilio 1998). However, interactions between turf-forming algae and sediments have rarely been analysed directly and quantitative data on the amount or dynamics of sediments accumulated are scarce. The amount of sediment trapped within turfs formed by Corallina spp. on intertidal shores in southern California varied seasonally from 5 mm to 4.5 cm, and was closely related to the species composition and structure of the turf itself (Stewart 1983). Sediment trapped in ﬁlamentous algal turfs occurring on a shallow subtidal reef in the Galapagos archipelago ranged from 180 g m2 to 1850 g m2, which corresponded to up to ﬁve times more than the biomass of the turf itself (Kendrick 1991). Whorff et al. (1995) found that algal turfs formed by a mixture of ﬁlamentous and corticated-terete algae on San Jose Island, Texas, trapped on average from about 270 g m2 to 2600 g m2 of sediment, depending on substratum slope and wave exposure. Airoldi & Virgilio (1998), working on turfs formed by the ﬁlamentous alga Womersleyella setacea (as Polysiphonia setacea) on exposed subtidal rocky reefs in the Ligurian sea (Italy), measured average amounts of sediment ranging from 86 g m2 to 924 g m2, which represented up to 96% of the total mass of the turf (Fig. 8). The amount of sediment accumulated remained relatively constant over a year, despite signiﬁcant temporal variations of rates of sediment deposition, supporting the hypothesis that turfs exert an important control on sediments. Furthermore, while the vertical growth of the turf was sensitive to the quantity and grain-size of sediment accumulated, its cover was unaffected: prostrate basal axes were resistant to sedi-

Total sediment: turf algae

40

30 20 10

1994

18SEP

27JUL

21JUN

5MAY

7APR

11MAR

28NOV

8JAN

0

1995

Figure 8 Sediment trapped in turfs formed by the ﬁlamentous alga Womersleyella setacea in subtidal rocky reefs south of Livorno, Italy (modiﬁed from Airoldi & Virgilio 1998, published with permission, photo by the author). Data are average ratios of dry mass of sediment trapped to biomass of turf-forming algae at three nearby sites (indicated by different symbols).

188

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

ment smothering, and, if damaged by severe scour, the turf was able to regain spatial dominance by quick vegetative propagation and regrowth of surviving axes (Airoldi 1998, Airoldi 2000b). In recent years there seems to have been a worldwide trend of increasing abundance of turfs that has been hypothesised to be related to enhanced perturbations, including perturbation by sediments, in coastal areas (Airoldi et al. 1995, Airoldi 1998 and references therein). Observations and experiments suggest that the abilities to propagate by vegetative reproduction and to entrap and withstand sediments are among the possible determinants of the persistence and spatial dominance of turfs (Sousa et al. 1981, Airoldi 1998, 2000b, Airoldi & Virgilio 1998, but see Irving & Connell 2002). By propagating vegetatively, turfs may exploit opportunistically and pre-empt space. Subsequent accumulation of sediment is thought to inhibit both recruitment of algae that form canopies and grazing by sea urchins and other herbivores, thus favouring the prevalence of turfs (but see the review by Vadas et al. 1992 and work by Boaventura et al. 2002 for suggestions of facilitation of recruitment of canopy algae by turfs). So far, little experimental data either support or refute these hypotheses.

Inhibition of recruitment, growth, and survival Sediments have been reported to be detrimental to a variety of rocky coast organisms. The most frequently postulated mechanisms are smothering and/or scouring of adult or juvenile stages, prevention from settling of larvae and propagules, and interference with normal foraging and feeding activities. Sometimes effects on growth, fertility and/or morphology have also been reported (e.g. Burrows & Pybus 1971, Dahl 1971, Espinoza & Rodriguez 1987, Sfriso & Marcomini 1996). Although direct observations of such effects are limited, several lines of indirect evidence suggest that the trends are real. The suggestion is that even the most tolerant hard-bottom organisms would eventually suffer inhibition and mortality above certain degrees of sedimentation. However, paucity of information on the regime of sedimentation greatly limits the interpretation and generalisation of the results, and prevents identiﬁcation of the critical levels above which such detrimental effects of sediments become manifest. Mortality Mortality of hard-bottom organisms as a consequence of severe smothering and scour by sediments has been reported frequently. Menge et al. (1994), for example, observed that at one study site affected by sand, mussels, barnacles and other invertebrates suffered severe mortality from burial by sand. Quantitative records of survival of mussels after different times and degrees of burial indicated that both partial and total burial by sediments reduced survival of mussels. In particular, Mytilus californianus was not able to survive total burial longer than 2 months, and in most cases mortality occurred within 12–18 days. Slattery & Bockus (1997) reported massive die-offs of the soft coral Alcyonium paessleri and other invertebrates along the shoreline of Ross Island: removal of organisms occurred from bands of substratum as extensive as 200 m2, and the disturbance was clearly related to localised sediment slides. Laboratory experiments conﬁrmed the intolerance of A. paessleri to a regime of sediment deposition and movement as that produced by sediment slides. Similar observations of dramatic mortality of epifaunal assemblages as a consequence of scour and smothering by landslides have been reported by Smith & Witman (1999). Seapy 189

LAURA AIROLDI

& Littler (1982) and Branch et al. (1990) observed pronounced declines in covers of algae, and especially invertebrates, following ﬂooding and sediment inundation events of rocky intertidal assemblages. In both cases, rates of mortality differed among species, resulting in shifts in the composition of dominant organisms. Robles (1982) reported that recurrent abrasion caused by shifting sediments was severe enough to remove most sessile organisms from local intertidal areas ranging in size from a few square metres to hectares. Similarly, extensive losses of invertebrates and algae from intertidal shores have been frequently reported in relation to severe sand scour and burial during storms (e.g. Seymour et al. 1989, Moring 1996, Underwood 1998). Burial by sediments was accounted to be the major cause of mortality of oysters and other sessile organisms in subtidal oyster reefs (Lenihan 1999). Several species of invertebrates, including cnidarians, sponges and ascidians, have been reported to suffer reduced growth and survival from sedimentation by burial, clogging of canals and chambers, or scour (reviewed by Moore 1977; see also Round et al. 1961, Bakus 1968, Gabriele et al. 1999). Sediments have also been reported to enhance survival of some species of algae. For example, burial into sediments has been shown to be a good overwintering location for Ulva spp. (Kamermans et al. 1998) and Gracilariopsis lemaneiformis (as Gracilaria lemaneiformis, Santelices et al. 1984). It should be noted, however, that such observations referred to species most frequently occurring in soft-bottom habitats. Inhibition of settlement and recruitment Effects of sediments on settlement or recruitment of rocky coast organisms have rarely been observed directly in the ﬁeld. However, there is a great deal of circumstantial observation to suggest that detrimental effects of sediments on many species may be related to inhibition of their larval or juvenile stages. For example, the ability of transplanted adult sporophytes of Hedophyllum sessile to survive and prosper at one site where the species was absent suggested that lack of adults was due to negative effects of sedimentation on the gametophytes (Dayton 1975). Steneck et al. (1997) observed that whereas adults of the branching encrusting coralline alga Neogoniolithon strictum were capable of surviving and growing even when covered by sediments for long times, spores appeared to require hard, relatively sediment-free substrata for successful germination and growth. One of the postulated mechanisms by which many turf-forming algae and invertebrates, including mussels, have been demonstrated to interfere with recruitment and survival of other hard-bottom species is through enhanced deposition of sediments (e.g. Dayton 1973, Sousa et al. 1981, Airoldi 1998, Albrecht 1998). Reed et al. (1988) suggested that high rates of mortality of gametophytes of the kelp Macrocystis pyrifera, observed within the ﬁrst week of settlement at one site in southern California, may have been due to high rates of sedimentation. Similarly, Deysher & Dean (1986) observed that a small but signiﬁcant proportion of the variance in recruitment of sporophytes of M. pyrifera on artiﬁcial substrata deployed in the water column was correlated with sedimentation rates measured by using sediment traps. Santos (1993) observed that abundance of Gelidium corneum (as G. sesquipedale) was negatively related to sediment load, and suggested that sedimentation was more likely to inﬂuence spore settlement and recruitment of this species than adult plants. Enhanced algal recruitment in previously unsuitable habitats was observed as a consequence of an episodic storm that removed ﬁne sediments from rocky reefs (Renaud et al. 1996, 1997) but such enhancement was not observed in those areas in which sediment persisted after the storm. Vogt & Schramm (1991) suggested that one of the causes of the decline of populations of Fucus spp. in Kiel Bay was the loss of substrata suitable for settlement 190

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

through deposition of sediment from the eroded cliffs. Moran (1991) analysed changes in the rates of recruitment of fouling assemblages on artiﬁcial panels during dredging operations in a port in Australia. After 2 wk panels submerged before the dredging contained almost twice as many species as those that were submerged during dredging. The author suggested that direct physical removal of larvae was the probable limiting factor, because organisms appeared to survive the turbid and toxic conditions once settlement and metamorphosis had taken place. Observations made during experiments involving manipulation of rates of sediment deposition suggested that early recruitment of encrusting algae was negatively inﬂuenced by sediment scour (Airoldi 2000a) but rapid cover by turf made it difﬁcult to quantify the importance of this process over appropriately long periods. Yoshida et al. (1997) transplanted embryos of Sargassum horneri at two sites, one where the species was naturally absent and another where it occurred at high densities. They observed that at the site where S. horneri was absent, the growth of germlings was greatly inhibited due to accumulation of ﬁne sediments and concluded that sedimentation was probably the most important factor affecting the settlement, growth and consequently the distribution of S. horneri at their study sites. The prevalence in sediment affected areas of species that reproduce by vegetative propagation or that have reproductive cycles synchronised with seasonal ﬂuctuations in the levels of sediments (see p. 179) also supports the hypothesis that early settlement stages are sensitive to mortality from sediments. Indeed, conﬁrmation of the hypotheses that sediments can inhibit the settlement and recruitment of species that propagate by sexual reproduction has come from many laboratory and ﬁeld experiments, which have shown the general susceptibility of larval and juvenile stages to sediments (see pp. 196 and 207). It has been suggested that sometimes sediments also have positive effects on settlement and recruitment of species. Kennelly (1983), for example, reported a positive correlation between sedimentation and recruitment of macroalgae in a subtidal kelp forest. He suggested that sedimentation might have been beneﬁcial to early growth of algae possibly by supplying nutrients or by protecting algae from disturbance from water movements or grazers. The author, however, also discussed the possibility that more silt might have been trapped where there is more algal cover, or that different effects might have been observed with greater sediment cover. Inhibition of grazing and predation Suggestions that herbivorous organisms are deterred by sediments are numerous. This suggested deterrence is particularly important because one of the postulated mechanisms by which sediments may control the algal vegetation on rocky coasts is through inhibition of grazing. Scarcity of both large herbivores (e.g. sea urchins, limpets, chitons, herbivorous ﬁshes) and small herbivores (e.g. amphipods, isopods, small gastropods) has been reported frequently from areas characterised by the presence of high levels of sediments both in rocky coasts (e.g. Emerson & Zedler 1978, Stewart 1982, 1989, Ebeling et al. 1985, Miller 1985, D’Antonio 1986, McGuinness 1987a, Trowbridge 1992, Schroeter et al. 1993, Airoldi & Virgilio 1998, Airoldi 2000a, Pulfrich et al., in press) and coral reef environments (e.g. Lim & Chou 1988, Steneck et al. 1997, Umar et al. 1998). However, direct evidence of negative effects of sediments on herbivores is scarce, and the mechanisms by which sediments might deter herbivores (e.g. direct damage to tissues by scour, smothering by clogging of respiratory apparatus or other physiological stresses, interference with movements or feeding activities, prevention of ﬁrm attachment, inhibition of recruitment) are unclear. D’Antonio (1986) hypothesised that the prevalence of Neorhodomela larix (as Rhodomela larix) in rocky shores affected by sand was related to 191

LAURA AIROLDI

Figure 9 Effects of sand burial on survival of transplanted individuals of the chiton Katharina tunicata (from D’Antonio 1986, published with permission). Open circles represent chitons transplanted to areas that did not experience sand coverage, while solid circles represent chitons transplanted to areas which experienced sand coverage.

the notably scarcity of sea urchins, small herbivores, and especially chitons (Katharina tunicata) deterred by sediments. Experiments in which adults of K. tunicata were transplanted to areas that either experienced or did not experience burial by sand conﬁrmed that these chitons were not able to tolerate sand burial, as their mantle suffered severe abrasion during coverage by sand (Fig. 9). Long-term monitoring at three sites that experienced different degrees of disturbance by sediments suggested that both abundance of herbivorous sea slugs and rate of attack of their algal prey Codium setchellii decreased with increased disturbance by sand (Trowbridge 1992). Experiments in which limpets were transplanted to areas affected or unaffected by sediments (Robles 1982), or in which they were exposed to burial conditions in the laboratory (Marshall & McQuaid 1989), suggested that some species of limpets are not able to tolerate burial by sediments. This result is consistent with observations of negative correlations between density of limpets and presence of sediments (Engledow & Bolton 1994, L. Airoldi & S. J. Hawkins, unpubl. data). Furthermore, laboratory experiments suggest that rate of grazing of limpets may be reduced by the presence of even a thin layer of sediment (L. Airoldi & S. J. Hawkins, unpubl. data). There are also observations suggesting that sediments have negative effects on predator organisms. Schroeter et al. (1993), for example, reported a decline of abundance of sea stars following enhanced sediment load by a power plant. Pulfrich et al. (in press) observed signiﬁcantly lower densities of predators, such as Burnupena spp. and Nucella spp., at sheltered sites affected by ﬁne discharges from diamond mines compared with unaffected sites. Menge et al. (1994) reported weak predation by sea stars at a wave-protected site regularly buried by sand. They hypothesised that weak predation was due to direct negative effects of 192

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

sediments on sea stars (e.g. interference with foraging and feeding activity, or physical and physiological stresses), or to indirect effects due to mortality and reduced abundance of prey.

Observations of effects of sediments from human discharges There are observations of effects of sediments either indirectly related to enhanced erosion and runoff or as a direct consequence of discharges of industrial and urban wastes (see p. 187). These observations have consistently reported dramatic effects on rocky coast assemblages, including changes in the composition and distribution of species, and/or impoverishments in the richness and abundance of hard-bottom organisms, sometimes resulting in patterns of spatial dominance by a few monopolising species (e.g. Castilla & Nealler 1978, Ellis 1988, Moran 1991, Saiz-Salinas & Urdangarin 1994, Gorostiaga & Díez 1996, Konar & Roberts 1996, Turner et al. 1997, Kim et al. 1998, Roberts et al. 1998, SaizSalinas & Urkiaga-Alberdi 1999). An example of such changes is the decline in cover of canopy algae reported in the past decades by many authors from different parts of the world (e.g. Littler & Murray 1975, Thom & Widdowson 1978, Seapy & Littler 1982, Vogt & Schramm 1991, Benedetti-Cecchi et al. 2001, Eriksson et al. 2002). This pattern appears to be paralleled by a trend of increasing abundance of turf-forming algae (Airoldi et al. 1995 and references therein) that, once established, seem to inhibit invasion of canopy forming algae and other organisms (Sousa et al. 1981, Airoldi 1998, see also p. 187). Paucity of longterm quantitative data, however, makes it difﬁcult to quantify the trends and unequivocally attribute the causes to enhanced sediment load from human activities. Furthermore, in many cases discharge of sediments was just one of many stresses affecting rocky coast assemblages, and effects due to sedimentation have not been separated from effects caused by potentially toxic organic and/or chemical compounds. For this reason, the present section is restricted to a few examples of effects of sediments from human discharges in which effects of sediments per se were considered predominant. In a study on the ecological effects of cooling water discharges from a coastal nuclear power plant in southern California, Schroeter et al. (1993) reported signiﬁcant reductions in density of snails, sea urchins and sea stars from the rocky substrata close to the diffusers over 2 yr after the power plant became operative. A reduction of the size and density of the kelp forest was also observed. Such decreases were paralleled by increases in abundance of two ﬁlter-feeding species, a gorgonian coral and a sponge. The most plausible cause of the changes in the composition and abundance of species was identiﬁed in the offshore discharging of turbid nearshore waters, which created turbidity plumes and enhanced sediment deposition over the kelp forest as far as 1.4 km from the diffusers. As a consequence, muddy sediments accumulated on the bottom at the closest affected site (0.4 km from the diffusers) and became armoured with coarser materials. Hyslop et al. (1997) compared the composition, abundance and distribution of dominant plants and animals at several rocky shores affected or unaffected by dumping of colliery wastes along the coastline of northeast England. They reported that while the distribution of animals was not related to colliery wastes, diversity of macroalgae was signiﬁcantly negatively correlated with colliery waste inputs and particularly dramatic reductions in cover at the affected sites were observed for the species Palmaria palmata and Ulva lactuca. The authors suggested that, because colliery waste leaches much of its toxic chemical content into the sea, detrimental effects were most 193

LAURA AIROLDI

likely related to the physical presence of sediments. Indeed, a laboratory experiment conﬁrmed the prevalent role of abrasion by sediments in affecting negatively the abundance of U. lactuca (Hyslop & Davies 1998). Dramatic declines of diversity and abundance of species on rocky coasts have been consistently reported as a consequence of the discharges of copper mine wastes (e.g. Castilla & Nealler 1978, Fariña & Castilla 2001 and references therein). In particular, Fariña & Castilla (2001) reported patterns of notably low species diversity on affected shores owing to the absence of many species of algae and ﬁlter-feeding animals, and monopolisation of space by Enteromorpha compressa. The authors emphasised that while previous work attributed such effects of copper mine wastes only to the high concentrations of trace metals, the mechanical effects due to the presence of sediments appeared to be an important and overlooked factor. Pulfrich et al. (in press) reported marked differences in species composition between rocky coasts close to and distant from the site of disposal of ﬁne sediments from diamond mines in Namibia. The effects of ﬁne deposits were restricted to sheltered sites, whereas no effects were observed at exposed sites where sediments were probably dispersed by wave action. Effects observed included reductions in the densities of grazers and predators, and proliferation of opportunistic foliose algae. Increased dominance of ﬁlter feeders was also observed, but only at intertidal sites.

Relationships between sediments and species diversity Sediments are believed to affect the diversity of assemblages on rocky coasts. The prevalent opinion is that “high” sediment loads are detrimental to the overall diversity of rocky coast organisms, through inhibition of recruitment and mortality of less tolerant species (e.g. Devinny & Volse 1978) and/or through enhancement of spatial dominance by a few tolerant, space-monopolising species (Airoldi et al. 1995). Observations consistent with this hypothesis have been reported by many authors (e.g. Daly & Mathieson 1977, Little & Smith 1980, Mathieson 1982, Seapy & Littler 1982, Kennelly 1991, Evans et al. 1993, Crothers & Hayns 1994, Saiz-Salinas & Urdangarin 1994, Falace & Bressan 1995, Birje et al. 1996, Naranjo et al. 1996, Kim et al. 1998, Saiz-Salinas & Urkiaga-Alberdi 1999, Smith & Witman 1999), and have been emphasised in studies on the impact of human-related sediments on rocky coast organisms (see p. 193). Particularly interesting are the results of Engledow & Bolton (1994) who analysed patterns of diversity of macroalgal species at several low shore sites along the coasts of Namibia, and quantiﬁed several physical and biological factors that might potentially affect species diversity. Their results showed that species diversity (Shannon-Wiener index) within each plot was negatively correlated with the amount of sand present but only at levels of sediment greater than 5.6 kg m2 (Fig. 10). This pattern was related to changes in equitability (Simpson’s Dominance index) rather than species richness. The authors suggested that sediments inﬂuenced the diversity of the assemblage by excluding less tolerant species and by favouring monopolisation of space by most tolerant species relieved from competition. On the other hand, several authors have provided data that support the hypothesis that, in some cases, presence of sediments may promote diversity of species on rocky coasts by increasing patchiness and habitat heterogeneity, preventing monopolisation of space by competitively dominant species, controlling the balance between sand-tolerant and sandintolerant species and providing new habitat to infaunal species typical of soft bottoms (Foster 1975, Robles 1982, Taylor & Littler 1982, Littler et al. 1983, Gibbons 1988, 194

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Figure 10 Relationship between diversity of macroalgae and degree of sand inundation in the lower eulittoral zone at 18 sites along the coast of Namibia (from Engledow & Bolton 1994, published with permission). The vertical line indicates the level of sand accumulation (350 g corresponding to 5.6 kg m2, size of quadrat was 25 cm 25 cm) below which sand did not have an effect.

Jørgensen & Gulliksen 2001). McQuaid & Dower (1990), in particular, showed that along the coasts of South Africa, total faunal species richness was higher for rocky shores regularly inundated by sediments than for rocky shores unaffected by sediments and sandy shores combined. They attributed the causes of this pattern to two factors. The ﬁrst was that the species recorded covered a spectrum of tolerance to sediments that varied from complete intolerance (species restricted to permanently sand-free areas, such as vertical cliffs), to complete dependence (species typical of soft-bottom habitats and thus restricted to the sand itself). The second was that within shores, patterns of accumulation of sediments were unpredictable and heterogeneous: species were thus often excluded locally by the presence of sand but patchiness of deposits resulted in a few being eliminated from the shore as a whole. Interpreting these contrasting views is difﬁcult, because in many studies the temporal and spatial context is not explicit, and because of lack of quantitative standardised measurements of the regime of perturbation by sediments. It is suggested here that differences arise because effects of sediments on rocky coast organisms vary in space and time, depending on the characteristics of the regime of sedimentation and their interactions with variable environmental and biological factors (see pp. 196 and 207). Thus, for example, the sub-lethal chronic smothering by a moderate layer of sediment may have different effects on the diversity of rocky coast organisms than severe and unpredictable scour events that create patches of open space (Airoldi 1998). Similarly, while a “moderate” regime of disturbance by sediments may promote diversity of species, “excessive” disturbance may result in dramatic declines in species diversity (Seapy & Littler 1982). Rocky coasts affected by sand may 195

LAURA AIROLDI

have fewer species of seaweeds than adjacent unaffected coasts: even so, within sandabraded coasts, microhabitats may occur, causing differential diversity and/or abundance of species (Daly & Mathieson 1977, Littler et al. 1983, Airoldi & Cinelli 1997, Airoldi 1998). Effects of sediments could also vary across different habitats and geographical locations, depending on local environmental and biological characteristics, and could be influenced by the “vulnerability” (sensu Sousa 2001) of the organisms affected. For example, accumulation of sediments within mussel beds or turf-forming species may enhance richness and abundance of infauna (e.g. Tsuchiya & Nishihira 1985, Gibbons 1988, Grahame & Hanna 1989), while inhibiting growth of macroalgae (e.g. Dayton 1973, Sousa et al. 1981, Airoldi 1998, Albrecht 1998). Relationships between perturbations and diversity in natural systems are complex, and reflect the differential expression of lifehistory attributes under different regimes of disturbance (Petraitis et al. 1989, Airoldi 1998, Sousa 2001). Deeper knowledge of the causal mechanisms by which sediments interact with affected organisms, greater consideration of scale issues, and quantitative information on the characteristics of the regime of perturbation by sediments are necessary to improve our present understanding of the relationships between sedimentation and species diversity on rocky coasts.

Laboratory experiments The ﬁrst laboratory experiments on the effects of sedimentation on rocky coast organisms were carried out in the 1970s, and became relatively frequent in the 1990s (Table 2). Experiments have focused on the effects of smothering by sediments on settlement, recruitment, growth, survival, or behaviour of a variety of species, whereas only a little work has been done on the effects of scour or substratum instability. Despite the short-term nature of laboratory experiments and their shortcomings (including the frequent arbitrary choice of treatment levels), these studies have detected a variety of responses of rocky coast organisms to sediment load, and have given insight into the possible mechanisms of the action of sediments. A general trend that emerges from these experiments is that adult individuals of many species (e.g. Zonaria farlowii, Ahnfeltiopsis linearis, Neorhodomela larix, Gracilariopsis lemaneiformis, Laminaria saccharina, Codium setchellii, Ulva spp., Siphonaria capensis) can survive some degree of burial, and resume growth or regenerate vegetatively from remaining fragments, despite often remarkable negative effects on biomass, growth, or photosynthetic activity (Table 2). Conversely, larvae, propagules, early post-settlement stages and juveniles generally suffered severe stress and mortality from sediments. Moss et al. (1973), for example, observed that growth of zygotes of Himanthalia elongata was inhibited under a layer of silt 1–2 mm thick, and that attachment on silt was insecure. Similarly, Norton (1978) showed that an underlying layer of silt prevented attachment of young sporophytes of kelps, whereas an overlying layer of silt inhibited the development of gametophytes. Devinny & Volse (1978) showed that sedimentation both prevented spore settlement and smothered gametophytes of Macrocystis pyrifera although survival increased markedly when spores were allowed to attach for 24 h (Fig. 11). Arakawa & Matsuike (1992) showed that presence of sediments inhibited the insertion, germination, survival and maturation of gametophytes of the kelps Ecklonia cava and Undaria pinnatifida. In these species adhesion of zoospores and maturation of gametophytes were the most sensitive phases, whereas germination and survival of gametophytes were more 196

197

Sand burial (natural

Ahnfeltiopsis linearis

Plants of both species attached to rocks

Burial by natural sand. S Plants were exposed to burial for increasing intervals (up to 6 months). Experiments were not replicated. Reported effects were qualitative. Degree of sand burial was not speciﬁed.

Disks of U. lactuca were grown, both in the light and in dark, in presence of control ﬁltered sea water or water from Dublin Harbour with or without mud. Details of experiments and concentration of mud were not speciﬁed. Effects on growth (disk diameter) were measured after 14 and 21 days.

Presence of mud, polluted by sewage, from Dublin Harbour. S

Zonaria farlowii

Settlement of L. saccharina on slides was followed in absence and presence of a layer of silt. Detail of experiment and abundance of silt were not speciﬁed. Reported effects were qualitative.

Presence of silt. ?

Laminaria saccharina, Ulva lactuca

Experimental set-up

Factor/type of sediment (size)

Species

After 1 month, G. platyphyllus plants were

Apical growth ceased short after burial. Living tissue gradually reduced as burial time increased. Only basal proliferations survived over 6 months. Resumed growth was often observed, but it was vulnerable to light exposure. Burial also affected thallus morphology. Effects were similar to those observed in natural populations in the ﬁeld.

Growth of U. lactuca was enhanced by the presence of mud but only in the presence of light. Effects were likely related to organic enrichment due to sewage.

Settlement of L. saccharina was inhibited in the presence of silt covering the slides.

Results

Markham &

Dahl 1971

Burrows 1971

Reference

Table 2 Summary of selected laboratory experiments on the effects of sedimentation on rocky coast organisms. S experiment done under still conditions, T experiment done under turbulent conditions, ? unknown water movement conditions. *The algae of in this study occurred as free-ﬂoating thalli in a lagoon with a sandy bottom. However, the species also occurs at rocky coasts. EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Spores were exposed to addition of increasing amounts of sediments (0 mg cm2 to 108 mg cm2) before and after settlement. Effects on spore density were quantiﬁed. An additional experiments, in which treatments were exposed or not to light, tested whether effects were due to light deprivation.

Fertilised eggs were exposed to addition of a layer of silt before and after settlement. Details of experimental design were not given. Reported effects were qualitative.

Macrocystis pyrifera Sand burial, scour, and presence of underlying sand (construction sand, 74 m). S and T

Presence of an underlying or overlying layer of silt, 1–2 mm thick. S

Himanthalia elongata

were subjected to burial of the lower portions and shaken. Details of experimental design were not given. Reported effects were qualitative.

Experiments were done before the construction of an unloading terminal. Ore dust (0.5 g in 100 ml ﬁltered sea water) was shaken with algal thalli. Amount of ore dust retained and transmission of radiant energy through thalli were measured. Details of experimental design were not given. Concentration was chosen arbitrarily, due to lack of information on amount of spillage.

sand). T

(as Gymnogongrus linearis), G. platyphyllus

Experimental set-up

Porphyra umbilicalis, Presence of four Plumaria plumosa different types of iron (as P. elegans), ore dust (8–25 m). T Polysiphonia lanosa, Cladophora rupestris, Pelvetia canaliculata

Factor/type of sediment (size)

continued

Species

Table 2

198

Boney 1978

Moss et al. 1973

Newroth 1972

Reference

Spore attachment was reduced by 90% in Devinny & the presence of 8 mg cm2 of sediments and Volse 1978 was prevented at 10 mg cm2. Survival of established germlings was reduced by 90% at 108 mg cm2. Effects were not attributable to light deprivation. Effects of sediments were more severe in moving water.

All algae retained considerable amounts of iron ore dust. Differences in retention ability between algae were observed.

Zygotes settled on a silt substratum, but attachment was insecure and they assumed a different shape. Growth of settled zygotes ceased under a layer of silt.

dead. Conversely G. linearis plants survived and produced new branches, fronds and cystocarps. Growth was still observed after 6 months. Results were in agreement with observations in the ﬁeld and results of transplant experiments.

Results

LAURA AIROLDI

Presence of an underlying or overlying layer of natural silt. S

Saccorhiza polyschides, Laminaria saccharina

199 Branched and unbranched pieces (about 10 cm long) of thalli were buried under 0 cm, 2 cm and 10 cm of sand, either in vertical or horizontal position with respect to the sand surface. Growth of thalli was measured over 90 days. Two experiments in which plants were placed under 15–20 cm of ﬁne sand for over 3 months, and survival was compared with controls. Observations on survival continued after removal of the plants from sand.

Neorhodomela larix Burial by natural sand (as Rhodomela larix) (125 m to 250 m). S and its epiphytes, Microcladia borealis, Cryptosiphonia woodii, Ulva sp., Mastocarpus papillatus (as Gigartina papillata), Coralline crust

Spores of both species were exposed to addition of natural non toxic sediments (0.3 mg mm2) before and after settlement. Details of experimental design were not given. Reported effects were qualitative.

Effects of sediments in moving water were also tested. Experiments were run for 4 days, as pilot studies suggested that most effects occurred during the ﬁrst days of germination.

Experimental set-up

Burial by natural sand. S

Gracilariopsis lemaneiformis (as Gracilaria lemaneiformis)

Factor/type of sediment (size)

continued

Species

Table 2

After 3 months of burial, holdfasts and basal crusts of R. larix were intact, but growth was inhibited and upright portions had decayed. Epiphytes disappeared after 6 wk. Buried coralline crusts were similar to control specimens, but reduced in size and slightly discoloured. Ulva sp. suffered severe mortality, less than 10% surviving. Other species did not survive 1 month of burial. Results were consistent with ﬁeld observations.

Thalli survived burial, but none increased in weight or length.

Spores of both species germinated and grew normally if settled on top of silt, but could not attach. If spores were covered by sediment after attachment, they germinated, but few gametophytes were produced and no sporophytes were formed. Results could only in part explain the distribution of the two species in the ﬁeld.

Results

D’Antonio 1986

Santelices et al. 1984

Norton 1978

Reference

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Factor/type of sediment (size)

Effects of natural silt trapped by algal turfs. S

Effects of sand burial and reduced oxygen tension. S

Effect of sand burial. S

Panulirus argus

Patella granularis, Siphonaria capensis

Anthopleura elegantissima

continued

Species

Table 2

200 Experiments tested the effects of burial (0 cm or 12 cm of sand) on behaviour of replicated aggregated clonal, wandering clonal, small solitary and large solitary forms of A. elegantissima. Experiment lasted 4 days. Effects were measured as proportion of individual per experimental unit that remained buried, were detached or successfully escaped

Repeated experiments tested the effects of burial (0 mm or 15 mm of sand) and reduced concentrations of oxygen (0.8 ml l1, 2.0–2.8 ml l1 and 3.6–4.8 ml l1) on survival of replicated limpets. An additional experiment separated the effects due to presence or weight of sand. Experiments lasted 7 days. The oxygen consumption at reduced tensions and the capacity of anaerobiosis of the two species were also quantiﬁed.

Pueruli or juvenile spiny lobsters were exposed to clumps of Laurencia spp. with silt removed or unmanipulated, in order to test the effects of silt trapped into algal turfs on time-to-metamorphosis and early postsettlement survival, and on postlarval settlement and juvenile habitat selection.

Experimental set-up

Wandering and aggregated forms showed little mobility when buried by sand, small solitary forms escaped burial by increasing their mobility, while large solitary forms avoided burial by elongating their columns above the sand. It was concluded that solitary forms are better adapted to sand-inﬂuenced habitats, because of their large ﬁnal size or high mobility. Results

S. capensis survived for signiﬁcantly longer periods than P. granularis in sand burial and reduced oxygen tensions. Negative effects on P. granularis were only in part attributable to weight of sand. Differences in mortality between the two species could be explained by differential respiratory responses during exposure to reduced oxygen tensions. Results could explain the distribution of the two species on rocky shores inundated by sand.

More pueruli settled in unsilted than silted algal clumps, and similar habitat preference was shown by juveniles. Silt had no effects on survival of pueruli through time to metamorphosis or metamorphosis. Scarcity of spiny lobsters observed in habitats with silt was attributed to low rates of postlarval settlement in these habitats.

Results

Pineda & Escofet 1989

Marshall & McQuaid 1989

Herrnkind et al. 1988

Reference

LAURA AIROLDI

Experimental set-up is difﬁcult to understand as the article is in Japanese (abstract in English).

Plants were incubated after exposure to a sediment suspension for 2 h. No information is given about the type of treatment (e.g. amount of sediment tested or how treatment was applied to plants). Effects on growth,

Suspended particles of kaolinite.

Laminaria saccharina Presence of ﬁne grained sediments from test dredging. ?

Settled particles of kaolinite.

Plants were treated with increasing concentration of particles. Experimental set-up is difﬁcult to understand as the article is in Japanese (abstract in English). Four experiments in which plants were treated with increasing amounts of settled particles (from 0 mg cm2 up to 50 mg cm2). Experimental set-up is difﬁcult to understand as the article is in Japanese (abstract in English).

Suspended particles of kaolinite.

201

Growth, nitrogen uptake and chlorophyll a concentrations were signiﬁcantly reduced in treated plants with respect to controls. No effects were observed for photosynthetic capabilities or phosphorous uptake. Direct physical damage to the

Settling velocity of zoospores was negatively affected by adsorption on suspended particles. Insertion of zoospores on the base-plate was also negatively affected by settled particles. Adhesion of zoospores decreased to 50% at 0.5 mg cm2 and was prevented at 3 mg cm2. Germination of gametophytes decreased to 40% at 1 mg cm2 and was prevented at 10 mg cm2. Survival of gametophytes decreased to 50% at 1 mg cm2 and was prevented at 10 mg cm2. Maturation of gametophytes decreased to 50% at 0.25 mg cm2 and was prevented at 1 mg cm2. Settling of zoospores decreased to 0.023% at levels of turbidity greater than 2 mg l1. A reduction of kelp bed extension to 10% at levels of turbidity greater than 10 mg l1 was predicted.

partially explained the distribution of colonial and solitary forms on rocky shores inundated by sand.

from sand burial (anemones with their oral disc above the sand while still attached to the substratum).

Ecklonia cava, Undaria pinnatiﬁda

Results

Experimental set-up

Factor/type of sediment (size)

continued

Species

Table 2

Lyngby & Mortensen 1996

Arakawa & Morinaga 1994

Arakawa & Matsuike 1992

Arakawa & Matsuike 1990

Reference EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

202 Replicated specimens of A. paessleri from rocky and soft-bottom habitats were exposed to sediments (100 g) added every 4 h for either 24 or 96 h.

Effects of periodic deposition of sediments from a slide. T

Alcyonium paessleri

Replicated pieces of C. setchellii were buried in 5 cm of sand. Controls were unburied thalli and thalli buried in 5 cm coarse aquarium gravel, in order to separate the effects of reduced light and reduced water ﬂow. Biomasses of thalli were measured for 5 wk. A second similar experiment, tested responses of the 3 sympatric species to burial in 2 cm sand for 4 wk.

A. paessleri exhibited signiﬁcant mortality as a consequence of sediment addition regardless of its habitat origin. Mortality increased over time up to values

Drilling muds did not affect fertilisation or early development of H. rufescens. Settlement of competent larve was sometimes weakly positively affected. Settlement of larvae on coralline crusts (natural settlement inducer) was strongly negatively affected. Survivorship and viability of P. stearnsii were strongly negatively affected, due to increased tissue mortality: at highest concentrations individuals died after 6 days.

C. setchellii lost substantial biomass during 4 wk to 5 wk of sand burial due to both reduced light and reduced water ﬂow. Conversely the other two species lost little or no biomass. The high loss in biomass of C. setchellii in the laboratory was in apparent discrepancy with high survival to burial observed on the shore, and could be possibly explained by the use of algal discs in experiments.

treated plants was observed, as a consequence of thick layers of ﬁne-grained material deposited on the thalli.

photosynthesis, nitrogen and phosphorous uptake and chlorophyll a concentrations were quantiﬁed at weekly intervals. A similar study was repeated in the ﬁeld.

Experiments tested the effects of different concentrations (0 mg l1 to 200 mg l1) of drilling muds on fertilisation, early development, survivorship and settlement of H. rufescens and on adult survivorship, viability and tissue loss of P. stearnsii. Concentrations were chosen based on ﬁeld data. Effects due to toxicity of dissolved fractions were not separated.

Burial by natural sand. S

Codium setchellii, Ahnfeltiopsis linearis (as Gymnogongrus linearis), Neorhodomela larix

Results

Experimental set-up

Haliotis rufescens, Effects of water-based Paracyathus stearnsii drilling muds from an active platform. T

Factor/type of sediment (size)

continued

Species

Table 2

Slattery & Bockus 1997

Raimondi et al. 1997

Trowbridge 1996

Reference

LAURA AIROLDI

Factor/type of sediment (size)

Effects of biodeposition by mussels under high concentrations of suspended sediments. S

Presence of colliery waste of 3 grain sizes (500 m, 500–2000 m and 0–2000 m). S and T

Fucus vesiculosus forma mytili

Ulva lactuca

continued

Species

Table 2

203 Replicated plants were exposed to colliery-waste treatments (0 g l1 or 1 g l1) in both still and turbulent (shaking for 1 h each day) conditions, for 8 days. The physical and chemical effects of colliery waste on weight of

U. lactuca lost weight signiﬁcantly in presence of colliery waste; weight loss was most pronounced with particles 500–2000 m in size, under shaken conditions. Signiﬁcant weight gain with respect to controls was observed in still

Germling survival was lowest on both l and mf shells in aquaria with the highest proportion of live mussels (although differences were not signiﬁcant), suggesting that biodeposition was the most critical factor. It was hypothesised that lack of sexual recruits observed on mussel beds in soft bottoms compared to hard bottoms might be related to post-settlement detrimental effects due to large biodeposition in the former habitats related to high particle concentrations in the water column.

above 70%. A. paessleri was able to package silt into mucus to be sloughed away, which might explain the ability of this species to colonise soft-bottom habitats. However, scour by sediments produced necrotic wounds leading to mortality. Results were consistent with observations in the ﬁeld.

Sediment was released along a Plexiglas “slide” directed towards the soft corals, in order to mimic the effects of a landslide. Control specimens were mechanically agitated with a water ﬂow at similar time intervals. Effects on soft coral behaviour (expansion or contraction of the polyps) and conditions (tissue necrosis) were followed over 4 wk. Sediment packaged in mucus or accumulated to the bottom of the aquarium were quantiﬁed. Survival of Fucus germlings inoculated on live (l) and mortar ﬁlled (mf) mussel shells was compared among 3 treatments with different combinations of l and mf shells (25 l 5 mf, 15 l 15 mf, 5 l 25 mf), in order to separate effects of biodeposition from mussels (affecting germlings on both l and mf shells) from effects of shell cleaning activity by mussels (affecting germling only on l shells). The experiment lasted 5 wk. Amount of biodeposits in different treatments was not speciﬁed.

Results

Experimental set-up

Hyslop & Davies 1998

Albrecht 1998

Reference

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Differential effects of burial (i.e. light deprivation, sediment type, and sediment chemistry). Sediments from a tidal inlet. S

Burial under anoxic (from lagoon) and oxic (bird-cage sand) sediments, under freezing conditions. S

Ulva spp. (mixture of U. scandinavica, U. curvata, U. rigida and U. lactuca)

Fucus serratus

Factor/type of sediment (size)

continued

Species

Table 2

204 Zygotes settled on slides were exposed to different combinations of: (1) sediment types, i.e. ﬁne (63 m) low (1 mm), ﬁne high (3 mm), coarse (255–350 m) low, mussel biodeposits low, and no sediments, (2) light, i.e. incoming vs dark, and (3) sediment chemistry, i.e. oxygenated, de-oxygenated, de-oxygenated with presence of sulphyde. There were 4 replicates for each combination of treatments. Slides were retrieved after 5 days. Survival and growth of embryos were measured after 5 days of recovery.

Presence of hydrogen sulphide had overriding negative impacts on both survivorship and growth of F. serratus embryos, independently of sediment type and light availability. Simple anaerobiosis did not impair survival or growth. High levels of ﬁne sediments and organic rich mussel biodeposits signiﬁcantly reduced survival of embryos. Low irradiance did not inﬂuence survival of embryos but affected growth negatively.

Treatments affected viability and growth rate of Ulva spp. Growth was fastest after freezing in natural anoxic sediments, followed by freezing in oxic sediments and freezing under dark conditions. No recovery was found after freezing in a light-dark cycle. All thalli naturally buried were able to resume growth. It was concluded that burial in natural sediments may allow Ulva spp. to survive freezing conditions during the winter.

conditions with particles of 0–2000 m or in the presence of bags. It was concluded that under turbulent conditions colliery waste acts as a physical detrimental abrading agent, while in still conditions it promotes the growth of U. lactuca.

plants were partitioned by using particles in contact with the plants or separated by bags. Criteria used to choose treatment levels were not explained. Replicated pieces of Ulva spp. were buried under 5 cm of anoxic or oxic sediments. Viability after 4 wk was tested by incubation and compared with that of unburied thalli kept in the dark or in a light-dark cycle. All treatments were frozen to mimic the natural winter conditions in the lagoon. Viability of naturally buried thalli was also tested.

Results

Experimental set-up

Chapman & Fletcher 2002

Kamermans et al. 1998*

Reference

LAURA AIROLDI

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Figure 11 Effects of variable amounts of sediment on relative survival of spores of Macrocystis pyrifera during laboratory experiments (from Devinny & Volse 1978, published with permission). Circles represent experiments in which sediments were applied before spores or were initially mixed with spore solution. Squares represent experiments in which spores were given 1 day to attach before sediment was added.

tolerant to presence of sediments. Raimondi et al. (1997) showed that while settlement of competent larvae of the red abalone Haliotis rufescens was sometimes affected weakly yet positively by the presence of drilling muds, settlement of larvae on coralline algal crusts (known to induce red abalone larvae to settle) was severely decreased. Overall, these results support the hypothesis that rocky coast organisms that persist by sexual reproduction are more vulnerable to the presence of sediments than organisms that reproduce vegetatively, apparently because the former require stable substrata for settlement and attachment, and/or their juvenile stages are more sensitive to smothering by sediments than adult individuals. In most experiments, presence of sediments exerted some degree of stress on rocky coast organisms, as indicated by the negative effects observed on growth, biomass, survival or photosynthetic activity (Table 2). The suggestion is that any hard-bottom species would suffer stress above certain levels of sedimentation. Some species, such as Zonaria farlowii, Ahnfeltiopsis linearis, or Neorhodomela larix, were, however, clearly able to tolerate those 205

LAURA AIROLDI

stresses better than others. For such “sand-tolerant” species, it has been suggested that costs directly imposed by sediments are compensated for by indirect advantages. For example, D’Antonio (1986) showed that burial of Neorhodomela larix (as Rhodomela larix) by sediments imposed stress in terms of reduced growth and loss of upright portions. This species, which dominated sediment-affected rocky shores along the northwest coast of the USA, could, however, better tolerate burial than its epiphytes or other species sometimes occurring in the same habitats. D’Antonio suggested that stresses imposed by burial were possibly compensated for by indirect positive effects, including reduced predation by herbivores, reduced cover by epiphytes, protection from desiccation during the summer months, and reduced competition by potential space occupiers. Only a few experiments tested the effects of different sediment loads and, surprisingly, some studies did not report the levels of sediment applied to treatments (Table 2). Devinny & Volse (1978) showed that attachment of spores of Macrocystis pyrifera was reduced by 90% in the presence of 8 mg cm2 of sediments, and was inhibited at levels above 10 mg cm2, which formed a thin layer enough to occlude all the surface of the culture dishes. Established germlings, however, tolerated greater amounts of sediments, survival being reduced by 90% at 108 mg cm2. Arakawa & Matsuike (1992) showed that adhesion of zoospores of Ecklonia cava and Undaria pinnatiﬁda and maturation of gametophytes were prevented in presence of levels of sediments above 3 mg cm2 and 1 mg cm2, respectively, while germination and survival of gametophytes were inhibited at levels above 10 mg cm2. These results clearly indicate that effects of sediments vary in relation to their quantity. Scarcity of information about responses of species to different levels of sediments, however, limits, at present, the possibility of identifying threshold levels of sedimentation for rocky coast organisms. Effects of sediments have sometimes been compared under still and turbulent water conditions. Experiments on Ulva lactuca (Hyslop & Davies 1998) indicated that while this species seems to tolerate well, and sometimes even be enhanced by, the presence of colliery waste particles under still water conditions, it was severely affected negatively under turbulent conditions. Devinny & Volse (1978) measured greater negative effects of sediments on survival of gametophytes of Macrocystis pyrifera in moving rather than in still water. Norton (1978) observed that spores of kelps could germinate and grow normally on still sediments but if the medium was disturbed, they drifted away. Slattery & Bockus (1997) observed that presence of silt was less problematic for the survival of the soft coral Alcyonum paessleri than scour, which produced necrotic wounds that lead to mortality. Overall, laboratory experiments suggested more severe effects of sediments on both adults and settlement stages under turbulent than still water conditions, presumably because of both abrasive scour and washing effects on individuals with an insecure attachment. It should, however, be noted that most experiments lasted for short times and although short-term burial by sediments may be less stressful in still than turbulent waters, the opposite pattern may occur during long-term burial conditions because circulation of water might enhance diffusion of gases and nutrients through the sediments. Very few attempts have been made to separate different aspects of sedimentation that affect rocky coast organisms. Chapman & Fletcher (2002) have investigated the mechanisms by which sediments negatively affect survival and growth of embryos of Fucus serratus. In particular, they separated three components related to smothering by sediments, speciﬁcally: a) physical components associated with different sediment types and grain sizes, b) light deprivation due to cover by a layer of sediment, and c) chemical components associated with 206

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

changes in oxygen and hydrogen sulphide levels that often occur under sediments in nature. Their results showed that the presence of hydrogen sulphide had overriding negative impacts on both survival and growth of F. serratus embryos, independently of type of sediment and availability of light. Interestingly, simple anaerobiosis generally did not have negative effects. Deprivation of light did not affect survival of embryos but inﬂuenced their growth, which is consistent with results by Devinny & Volse (1978) on effects of light on survival of germlings of Macrocystis pyrifera. Fine sediments and organically-rich biodeposits had more detrimental effects on embryo survival than coarse sediments, which the authors attributed to accumulation of metabolic waste products of the embryos as a consequence of constrained diffusion. Results from laboratory experiments thus give clear indications that sediment quantity as well as quality may have important effects on rocky coast organisms, and that severity of effects of sediments may be related to variable environmental and biological factors, such as the degree of water movements, or the stage of development of the organisms themselves.

Field experiments While experimental work on the effects of sedimentation has been relatively common in coral reef environments (reviewed in Rogers 1990), ﬁeld experiments to analyse the effects of sedimentation on rocky coast organisms were undertaken only recently, and such studies are still few (Table 3). Explanations for such scarcity of experimental studies in rocky coasts probably include the high labour costs required to do such experiments in temperate habitats, and the difﬁculties that are often encountered in planning and executing such experiments. Furthermore, the spatial and temporal scales of ﬁeld manipulative experiments are necessarily small, which limits the possibility of generalising results at large spatial scales relevant for predicting changes in sedimentation rates as a consequence of human activities. The results of the few available experiments (including a study by Umar et al. 1998 which was done on coral reefs but tested the effects of sediments on algae belonging to the genus Sargassum that is common on rocky coasts) are reviewed here and the methodological difﬁculties encountered discussed.

Methodological problems Manipulating sedimentation rates in the ﬁeld is difﬁcult because patterns of deposition of sediments are often inﬂuenced by many physical and biological factors (see p. 164), which can invalidate the effectiveness of the manipulation. The commonest approach used in rocky habitats has been the reduction of sedimentation rates through the removal of sediment by water motion, for example, by wafting water with one hand or ﬂushing (Neushul et al. 1976, Kendrick 1991, Umar et al. 1998, Irving & Connell in press a,b), or by using transparent panels (e.g. Fig. 12) that intercept settling particles (Duggins et al. 1990, Airoldi & Cinelli 1997, Relini et al. 1998, Maughan 2001). Alternatively, sediment load has been increased through the addition of known amounts of sediment to the experimental plots (Kendrick 1991, Airoldi 1998, Airoldi & Virgilio 1998, Umar et al. 1998), or by burying natural or artiﬁcial hard substrata into sediments (McGuinness 1987b, Gotelli 1988). In both cases, the effectiveness of the manipulation is clearly dependent on the frequency of the application or 207

Type of manipulation/experimental set-up

Effects of sedimentation on recruitment of fouling organisms were tested by comparing colonisation between replicated settlement plates from which sediment was removed manually by divers (25 times during 1 yr) or that were left unmanipulated.

Effects of disturbance by sand on existing assemblages were tested by burying replicated boulders and their associated assemblages about 1–3 cm deep into sand, or leaving them on sand as control. Effects of short (13 days) and long-term (119 days) burial were compared. Similarly, effects on developing assemblages were tested by using bare boulders that were completely buried, half-buried or left unburied. Colonisation was followed over 266 days. Both experiments were repeated at highand low-shore levels.

The effects of sand burial on recruitment and growth of L. virgulata were analysed by experiments and observations. Replicated cement patio stones were either raised on limestone, raised on sand or buried into sand. Recruitment, and survival of L. virgulata were recorded over 1 yr, and

Species-assemblage/habitat characteristics/location

Fouling assemblages on Plexiglas plates (12 m in depth). Santa Barbara, California, USA. Sedimentation rates up to 2000 g m2 d1 during the winter.

Intertidal assemblages on boulders on two rock platforms inundated by sand. Sydney, Australia.

Leptogorgia virgulata on a limestone outcropping (1.5 m in depth) with patchy cover of sand (few mm thick). Evidence of frequent sand movements. Franklin County, Florida, USA.

208

Recruitment of L. virgulata was reduced by 50% on buried stones compared with stones raised on sand. Inexplicably, recruitment was reduced also on stones raised on limestone. Observations suggested that sand reduced recruitment also on natural substrata. Mortality rates

Recruitment onto fully buried boulders was virtually zero. Disturbance by sand also reduced the number and abundance of species in existing assemblages. Effects were particularly severe and rapid for grazers, while negative effects on algae (such as Polysiphonia sp. and Ulva lactuca) or sessile invertebrates became visible after long burial. Short-burial boulders generally had more species than long-burial ones, but differences were not always signiﬁcant. Effects also varied as a function of height on the shore and size of boulder.

No differences were observed between plates from which sediments were removed and unmanipulated plates. It was assumed that sediment removal was not frequent enough to decrease signiﬁcantly sediment deposition.

Results

Gotelli 1988

McGuinness 1987b

Neushul et al. 1976

Reference

Table 3 Summary of ﬁeld experiments on the effects of sedimentation on rocky coast species or assemblages. *This experiment was done on a fringing coral reef. Nevertheless, species belonging to the genus Sargassum are common on rocky coasts.

LAURA AIROLDI

continued

Invertebrates in subtidal understorey kelp environments (7–11 m in depth) in Washington State, USA. Kelps increased sedimentation by 2.4–4.8 times with respect to no kelp areas.

Species-assemblage/habitat characteristics/location

Table 3 Reference

Rates of recruitment of the polychaetes Duggins et al. 1990 Pseudochitinopoma occidentalis and spirorbid spp., and the bryozoans Membranipora membranacea and Tubilopora spp. were inhibited in treatments facing up (with sediments) compared with treatments facing down (without sediments). It was concluded that kelps exerted an important indirect inﬂuence on recruitment of benthic invertebrates by affecting sedimentation.

and mean colony sizes were similar among treatments, although the largest colonies were collected from the buried treatment. Buried colonies grew signiﬁcantly more than unburied colonies. It was concluded that recruitment of L. virgulata was affected by the distribution of sand, and that spatial patterns established at the time of recruitment persisted in the adult population. It was also suggested that increased growth rate of buried colonies enhanced their survival.

compared with those measured on natural rock in relation to the heterogeneous distribution of sediments. Effects on growth were tested by transplanting juvenile gorgonians on top of sand or almost completely buried into it. Growth was measured after 55 days.

Experiments tested the effects of kelp on recruitment of invertebrates, and tried to separate the roles of physical factors controlled by kelps (i.e. ﬂow speed, sediment and light). Replicated recruitment plates were deployed either facing up or facing down in each of kelp and no kelp treatments. Plates facing down were either opaque or transparent. Effects of sedimentation were tested by comparing plates facing up and opaque plates facing down only in kelp treatments. There were differences in ﬂow rates between the two types of plates, but these were considered irrelevant. Possible confounding effects due to the different orientation of the substrata were not considered.

Results

Type of manipulation/experimental set-up

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

209

continued

Crustose coralline and ﬁlamentous turf algae onto lava boulders in a lagoon (3 m in depth), Galapagos Archipelago. Sedimentation rates ranged from 100 g m2 d1 to 550 g m2 d1. Small size particles (0.02 mm) accounted for up to half of total sediment.

Species-assemblage/habitat characteristics/location

Table 3

Replicated bleached and sun-dried lava boulders were subjected to sediment accretion (physical addition of sediment), sediment erosion (removal of sediment by water motion once every 2 or 7 days) and scour (physical abrasion). Effects on recruitment of turf and coralline algae were followed over 51 days.

210

Turf recruited in all treatments. Percent Kendrick 1991 cover and biomass were enhanced by sediment erosion. Accretion of sediment negatively affected cover of turf, but this did not result in negative effects on biomass. Biomass was signiﬁcantly lower in treatments simulating scour. Effects on crusts were difﬁcult to quantify as crusts were overgrown by the turf, except in treatments simulating scour. It was suggested that persistence of crusts may be favoured indirectly by sand scour that removes the overgrowing turf.

Growth of Pseudochitinopoma Eckman & Duggins 1991 occidentalis, Membranipora membranacea, the bryozoan Cheilopora praelonga, and the sponge Myxilla incrustans was reduced in treatments facing up (with sediments) compared with those facing down (without sediments). Extent of responses differed among species. Negative effects of sedimentation on growth were not consistent with the observation that most species exhibited greatest growth and survival below kelps. It was suggested that negative effects of kelps due to enhanced sedimentation were compensated by positive effects due to reduced rates of ﬂow and abundance of microalgal turfs.

An analogous experiment was done to separate the effects of the same physical factors on growth of adult invertebrates. As above, possible confounding effects due to the different orientation of the substrata were not considered.

Reference

Results

Type of manipulation/experimental set-up

LAURA AIROLDI

Assemblages dominated by turf-forming algae on rocky subtidal reef (13–17 m in depth) at Calafuria, Italy. Sedimentation rates ranged from 2 g m2 d1 to 178 g m2 d1. Turfs trapped up to 924 g m2 of sediment. This was mostly composed of ﬁne (200 m) inorganic particles. Maximal disturbance by sediments occurred in late autumn and winter, from storms and scouring by sediments, when small patches of bare rock could be produced.

Table 3 continued Species-assemblage/habitat characteristics/location

Variable size, intensity, timing and location Airoldi 1998 of disturbance by sediments signiﬁcantly inﬂuenced the success of erect algae that persisted by sexual reproduction, but did not affect spatial dominance by the turf that persisted by vegetative reproduction. Severe disturbance by sediment favoured the development of erect algae by locally removing the turf from small patches of substratum. These positive indirect effects, however, only occurred if turf was

One experiment tested whether variable size (different size of plots), intensity (chronicle smothering, or abrupt disturbance), timing (time when disturbance started) and spatial location of disturbance by sediments affected the success of turf-forming and erect algae. Sediment smothering was manipulated by weekly additions of sediments as to obtain sedimentation rates increased by 0 g m2 d1 and 100 g m2 d1. Sediment scour was

Airoldi & Cinelli 1997

Reference

Reduction of sedimentation inﬂuenced the structure of the algal assemblage. Effects were most evident on the developing assemblage, and were dependent upon the time when succession was initiated. Cover of algal turf was unaffected by sediments, while its biomass was enhanced by reduced sediment load. Conversely erect algae only grew in plots scraped in June and exposed to natural sedimentation. Overall, species diversity was lower in plots with lessened sedimentation. It was suggested that sediments affected the diversity of the assemblage both directly, by controlling the biomass of turf, and indirectly, by modulating competitive interactions between turf and erect algae.

Results

Two experiments tested whether the structure and diversity of adult or developing assemblages (plots experimentally cleared at different times of the year) differed between replicated plots unmanipulated or in which sedimentation was reduced by using transparent Plexiglas panels. Cover and biomass of dominant species were quantiﬁed over 1 yr. Potential artefacts were assessed and were found to be irrelevant. Panels effectively reduced natural sedimentation rates by 35%.

Type of manipulation/experimental set-up

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

211

continued

Species-assemblage/habitat characteristics/location

Table 3 Results

disturbed at certain times of the year, and probably did not persist over time. It was suggested that by propagating vegetatively and trapping sediment, the turf monopolises space, and relegates erect algae to the status of fugitive species. Sediments affected negatively the biomass of the algal turf at values greater than 200 g m2 d1. Effects occurred only with medium and coarse sediments, and were mainly evident on an established turf. Patchiness in the deposition of sediment affected the biomass of the turf at the scale of metres, however turf cover was not affected by sediments at any spatial scale. It was concluded that negative direct effects of sediments on turf thickness over small spatial scales (cm to m) were probably compensated for by indirect positive effects on turf cover over large spatial scales (10 s to 100 s m), possibly though limiting grazing by herbivores or inhibiting recruitment of potential competitors. Encrusting, turf-forming and erect algae responded differently to spatial and temporal variation of disturbance, depending on life histories. Turf colonised space by vegetative propagation

Type of manipulation/experimental set-up

mimicked by abrading or completely scraping off the assemblage.

Responses of turf-forming algae to spatial variations in the deposition of sediments at different spatial scales were analysed by observations and experiments. The experiment tested the hypotheses that (1) different amounts (0 g m2 d1, 100 g m2 d1 and 200 g m2 d1) and grain sizes (ﬁne, medium and coarse) of sediment affected the cover and biomass of the turf, (2) any effects of sediments were independent of the stage of development of the turf, and (3) patterns were consistent at different spatial scales, ranging from centimetres to metres.

Effects of timing and location of disturbance (as produced in the area by severe episodic perturbations from storms and sediment scour) on colonisation of encrusting, turf-forming and erect algae

Airoldi 2000b

Airoldi & Virgilio 1998

Reference

LAURA AIROLDI

212

continued

Several experiments tested the effects of three levels of sediment load (addition up to 20 mm in order to double natural sediment thickness, removal by ﬂushing plots with seawater, and unmanipulated) on recruitment, growth, survival, degeneration and vegetative propagation of S. microphyllum.

Sargassum microphyllum on an intertidal fringing coral reef usually covered by a thin deposit of ﬁne sand, Great Barrier Reef, Australia.

213

Increased amounts of sediments signiﬁcantly decreased rates of recruitment, growth, survival and vegetative regeneration, but did not affect degeneration. Sediment removal did not have relevant effects. It was suggested that the greatest abundance of Sargassum spp. often observed in habitats with greatest sediment load was related to other factors

Multivariate analyses separated panels protected or non protected by sediments. Total abundance (biomass) and diversity of species were greatest on panels protected by sediments, which were colonised by diversiﬁed assemblages of oysters, ascidians and encrusting and erect bryozoans.

and quickly regained spatial dominance in patches disturbed at all times of the year and at all locations. Encrusting and erect algae occupied space mostly by colonisation of propagules, and their recruitment was much inﬂuenced by timing and location of disturbance. A regime of disturbance by storms and sediment scour adverse to recruitment of erect algae was identiﬁed as the probable cause of their scarcity in the study area.

were investigated by clearing replicated plots at 8 different times during 1 yr. The experiment covered spatial scales ranging from about 1 to 100 s m. Colonisation of patches by different algae was followed over 1 yr.

Effects of sedimentation on recruitment of epibenthic assemblages were tested by using horizontal settlement panels protected or non protected by glass screens. Panels (two replicates for each treatment) were deployed over 24 months. Both upper and undersides of panels were sampled.

Results

Type of manipulation/experimental set-up

Subtidal epibenthic assemblages on an artiﬁcial reef (18 m in depth) at Loano, Italy. Sedimentation rates ranged from 10 g m2 d1 to 213 g m2 d1.

Species-assemblage/habitat characteristics/location

Table 3

Umar et al. 1998*

Relini et al. 1998

Reference EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

214

Effects of light and sedimentation in inﬂuencing prevalence of invertebrates on down facing surfaces were tested by exposing replicated slate settlement panels to 4 different combinations of sedimentation and light: light but no sediment, light and sediment, no light and no sediment, and down facing panels. Sediment and light were manipulated by using opaque or transparent perspex panels (or frames). The experiments did not include unmanipulated panels. Development of the assemblage was followed over 13 months. Similar sets of panels were deployed at two different times for 39 days.

Encrusting algae and invertebrates on a sheltered plateau reef (6 m in depth), Lough Hyne, Ireland. Natural sedimentation rates ranged from 2 g m2 d1 to 12 g m2 d1. Organic content varied from 15% to 30%.

Epibiotic assemblages on Effects of light and sedimentation in artiﬁcial panels (5 m depth). inﬂuencing prevalence of algae and Outer Harbour, South Australia. invertebrates on up facing and down facing surfaces, respectively, were tested by

Type of manipulation/experimental set-up

continued

Species-assemblage/habitat characteristics/location

Table 3 Reference

Alternate states of algae vs invertebrate dominated assemblages appeared to be primarily maintained by light intensity, which facilitated algal cover on up facing

Irving & Connell 2002a

Downward facing panels supported greater Maughan 2001 number of species and total per cent cover than all other treatments, followed by panels with no light and no sediment, panels with light but no sediment, and panels with light and sediment. Cover of some species, such as Lithothamnion spp. was negatively affected by the presence of sediments, whilst other species, such as Anomia ephippium were not affected. Differences among panels exposed for short times were less pronounced than differences among panels exposed for 13 months, suggesting that differences among treatments were mainly related to postsettlement mortality. It was suggested that the negative phototactic response of many larvae may be a response to sediments as well as light. Differences between down facing panels and panels with no light and no sediment could not be explained.

correlated with sediments, such as nutrients, or to indirect effects of sediments on potential competitors or predators.

Results

LAURA AIROLDI

continued

Turf-forming algae (Feldmannia spp.) on artiﬁcial panels (5 m depth). Outer Harbour, South Australia. Natural sedimentation rates ranged from 322 g m2 d1 to 2033 g m2 d1.

Species-assemblage/habitat characteristics/location

Table 3

215

Reduced sedimentation enhanced cover of turf independently of substratum microtopography. Reduced sedimentation also enhanced biomass of turf: these effects were much greater on topographically complex than simple surfaces, while no differences between smooth and roughened plates were observed at natural rates of sediment deposition. It was suggested that heavy sedimentation can obliterate differences in turf abundance related to substratum microtopography

surfaces (full light) and invertebrates on down facing surfaces (reduced light). Although sedimentation was only partially responsible for differences between habitat types, it acted as a negative disturbance on the abundance of algae and survival of invertebrates. Effects were particularly evident under shaded conditions, suggesting important interactive effects between light and sedimentation. It was observed how the ability of invertebrates to withstand sedimentation was related to their morphology, where erect forms tolerated sediments better than prostrate forms.

manipulating light and sedimentation in orthogonal combinations. Epibiota was allowed to develop on down facing surfaces. After 100 days, some plates were rotated 180° and attributed to 3 levels of light intensity (full shade by using dark roofs, procedural control with clear roofs, unmanipulated) and two levels of sedimentation (natural or reduced by removing sediment manually every 2–3 days). Effects on per cent cover of epibiota were assessed after 65 days. Potential artefacts were assessed and were found to be irrelevant.

Effects of microtopography and sediment deposition on recruitment of turf were tested by exposing replicated settlement plated to an orthogonal combination of two sediment treatments (sediment removed manually every 2–3 days and unmanipulated) and two microtopographies (smooth or roughened surfaces). Cover and biomass of turf were quantiﬁed after 50 days.

Results

Type of manipulation/experimental set-up

Irving & Connell 2002b

Reference EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

LAURA AIROLDI

Figure 12 Plexiglas roofs (34 cm 39 cm) used to manipulate rates of sedimentation on rocky subtidal reefs south of Livorno, Italy (modiﬁed from Airoldi & Cinelli 1997, published with permission, photo by the author).

removal of sediment, as well as on environmental conditions (e.g. natural sedimentation rates, water ﬂow, composition of the biological community). Neushul et al. (1976), for example, reported that, during an experiment designed to test whether sedimentation affected settlement and growth of subtidal algae, the manipulation was unsuccessful, probably because removal was not frequent enough in relation to the local depositional regime. Similarly, Airoldi & Virgilio (1998) experienced uncharacteristic storms, which probably reduced the effectiveness of their manipulation due to removal of sediments that were experimentally added to their plots. Although such problems are not unlikely to occur, the effectiveness of the manipulations of sedimentation regime and the resulting levels of sediment deposition, accumulation and movement in experimental treatments are rarely quantiﬁed or even questioned (but see Airoldi & Cinelli 1997), which represents a major shortcoming of most ﬁeld experiments. Airoldi & Cinelli (1997) also discussed possible overlooked problems related to the use of panels or similar devices to reduce sedimentation rates. These structures are valuable instruments but they require a constant maintenance because they need to be perfectly clean so as not to limit the level of incoming irradiance. Furthermore, the presence of a panel may introduce the risk of potential artefacts (e.g. on the ﬂow microenvironment, levels of irradiance, access to predators, or larval supply). Similar potential artefacts could also arise from the removal of sediments by water motion or ﬂushing. A careful assessment of possible artefacts is necessary for the interpretation of the results (Airoldi & Cinelli 1997, Irving & Connell 2002a). So far, however, the possible inﬂuence of artefacts has been largely overlooked.

Measured effects Field experiments demonstrate that sedimentation affects the composition and distribution of rocky coast organisms and the overall structure and diversity of assemblages (Table 3, p. 208). Negative effects of sediments were demonstrated for various species of invertebrates, including sponges, gorgonians, polychaetes, bryozoans and grazing gastropods (McGuinness 1987b, Gotelli 1988, Duggins et al. 1990, Eckman & Duggins 1991 but note 216

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

possible confounding effects in the last two studies, Relini et al. 1998, Irving & Connell in press a), and for several species of algae (McGuinness 1987b, Kendrick 1991, Umar et al. 1998, Irving & Connell 2002a,b). In some cases responses were complex and difﬁcult to disentangle. Airoldi & Cinelli (1997), for example, observed that a reduction of sediment inputs enhanced the biomass of turf-forming algae but did not affect their cover; conversely, erect algae only grew in patches of bare rock exposed to natural, moderately high (2 g m2 d1 to 178 g m2 d1) sedimentation rates and produced at certain times of the year. Airoldi & Virgilio (1998) showed that, whereas cover of turf-forming algae was not affected by sediments, biomass was inﬂuenced negatively at rates of sedimentation 200 g m2 d1 and with grain sizes 250 m. Maughan (2001) suggested, and Irving & Connell (2002a) clearly demonstrated, complexes patterns of recruitment of algae and invertebrates in relation to combined effects of sedimentation, light intensity and surface orientation (see also the pioneer study by Muntz et al. 1972). In particular, Irving & Connell (2002a) emphasised that attempts to separate the effects of sedimentation and light intensity, which are not independent in nature, may not reveal the true effects of sedimentation on epibiotic assemblages. In agreement with results from ﬁeld observations and laboratory experiments, susceptibility to sediments was generally more pronounced in larval and juvenile stages than in adult mature assemblages. For example, the presence of sand reduced recruitment of Leptogorgia virgulata but enhanced the growth of established juveniles (Gotelli 1988). Disturbance by sand inhibited the development of assemblages on boulders, while mortality of established assemblages became notable only after long burial (McGuinness 1987b). Sediments signiﬁcantly reduced the recruitment, growth, survival and regeneration ability of Sargassum microphyllum (Umar et al. 1998). These authors, however, observed that adult populations were never completely killed, and indicated that the major mechanisms by which enhanced sediments inhibited S. microphyllum at their study site involved preventing attachment of new recruits and smothering of young fronds. Airoldi & Cinelli (1997) observed that effects of sediments on the structure of subtidal assemblages were more evident on developing than established assemblages and were dependent upon the time of the year when succession was initiated. Most ﬁeld experiments did not attempt to elucidate the underlying mechanisms by which sediments affect rocky coast organisms. Nevertheless, there are lines of evidence which suggest that responses of species to sediments are complex, and are probably the result of both direct effects on individual species and their propagules, and indirect effects related to changes in abundance of other potential competitors or predators (see also p. 191). Umar et al. (1998), for example, demonstrated that sediments negatively affected the development and growth of Sargassum microphyllum. Results, however, contrasted with the observation that S. microphyllum was most abundant in areas with greatest sedimentation. The authors suggested that although sediments imposed a stress on S. microphyllum, the costs of living in sediment-affected habitats were probably compensated for by indirect advantages due to detrimental effects of sediments on other potential competitors or herbivores. Similarly, Kendrick (1991) demonstrated that recruitment of coralline crusts was enhanced by treatments simulating scour. These treatments were the only ones where crusts were not completely overgrown by turf at the end of the experiment because turf appeared to be sensitive to scour. Kendrick concluded that positive effects of scour on crusts were possibly related to indirect negative effects on the abundance of overgrowing turf (but see Airoldi 2000a). The complexity of the effects of sedimentation has been elucidated by studies carried out 217

LAURA AIROLDI

over six years at an exposed subtidal rocky reef in the Ligurian Sea, Italy (Fig. 13). The area was subject to sediment loads that were moderately large compared with other rocky reefs, as a consequence of enhanced runoff and erosion from extensive ﬁres of coastal vegetation (Airoldi et al. 1996). Accumulation of and disturbance from sediments were variable in space and time (Airoldi & Virgilio 1998). Sediment deposition and scour were greatest following storms, particularly intense during late autumn and winter, and the regime of sedimentation varied signiﬁcantly among nearby sites, suggesting the important role of hydrodynamic conditions in inﬂuencing transport of sediment along the coast. Superimposed on these large-scale patterns, sediment was redistributed within each site. At a scale of metres, accumulation of sediment was patchy, probably reﬂecting differences in the microtopography of the bottom and in proﬁles of ﬂow-speed at the boundary-layer. The assemblage was

Figure 13 Example of complex direct (solid arrows) and indirect (dashed arrows) effects ( positive, negative and 0 no effects) of sedimentation in rocky coast assemblages. Parentheses indicate effects that occur only under certain circumstances, while question marks indicate processes that need further experimental veriﬁcation. The scheme derives from results of experimental studies of rocky subtidal assemblages in the Ligurian sea by Airoldi & Cinelli (1997), Airoldi & Virgilio (1998) and Airoldi (1998, 2000b). Filamentous turf accumulates sediment. In turn sediment can sometimes reduce turf thickness which does not affect turf cover. Accumulation of sediment is thought to deter grazers and inhibit recruitment of erect algae that compete for space with turf. Severe scour and accumulation can locally remove turf, allowing the temporary development of erect algae. Such positive indirect effects, however, only occur at certain times depending on propagule availability. Turf always recovers quickly from eventual damage by vegetative propagation. For further explanations see text and Table 3.

218

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

characterised by a notable low diversity of species and by the dominance of ﬁlamentous turf-forming algae that entrapped large amounts of sediments (Airoldi et al. 1995, Airoldi & Virgilio 1998). Observations and experiments indicated that sediments affected the structure of this assemblage in many direct and indirect ways. Airoldi & Cinelli (1997) demonstrated that species diversity decreased in plots where sediment inputs were reduced by using transparent screens. This pattern was related to an increase in the biomass of dominating turfforming algae and a decrease in the cover of erect algae. They suggested that sediments affected the diversity of the assemblage both directly, by controlling the biomass of turf, and indirectly, by modulating competitive interactions between turf and erect algae. Subsequent experiments (Airoldi & Virgilio 1998), conﬁrmed these hypotheses but showed that effects were more complicated than initially thought and varied at different spatial scales. Local accumulation of sediments could affect negatively the vertical growth of the turf but effects varied depending on the amount and grain size of sediments, the stage of development of the turf, and the concomitant action of other factors, such as the regime of disturbance and water movement in the area (Fig. 14). Conversely, cover of turf was not affected by sediments because prostrate basal axes appeared to be resistant to smothering and scour and, if damaged, the turf quickly regained spatial dominance by vegetative propagation. Overall, small-scale direct detrimental effects of sediments on the vertical growth of the turf appeared to be compensated for by indirect advantages to its horizontal distribution at large spatial scales. In particular, experiments showed that turf inhibited the recruitment of erect

Figure 14 Variable effects of sediments on cover and biomass of subtidal turfforming algae as a function of sediment characteristics (amount and grain size) and stage of development of the assemblage (modiﬁed from Airoldi & Virgilio 1998, published with permission). Data are mean values ( 1 SE), measured 4 months after the beginning of the experiment.

219

LAURA AIROLDI

algae that propagated by sexual reproduction (Airoldi 1998). Such effects were attributed to the ability of turf to entrap sediments; however experiments designed to test this hypothesis through manipulations of the relative abundances of turfs and entrapped sediments were unsuccessful (Airoldi & Cinelli 1996b). It was suggested that entrapped sediments also limited grazing by herbivores and experiments are now in progress to test this hypothesis (L. Airoldi and S. J. Hawkins, unpubl. data). Ultimately, severe disturbance by sediments could remove the turf from small patches or decrease its thickness, thus allowing the local development of erect algae which were relegated to the status of fugitive species (Airoldi 1998). Further experiments (Airoldi 2000b), however, showed that these positive indirect effects on erect algae occurred only if turf was removed at certain times of the year; overall, the temporal regime of disturbance in the area adversely affected the recruitment of erect algae, contributing to their notable scarcity. Airoldi (1998) concluded that the ability of turf to accumulate large amounts of sediment and to quickly recover vegetatively to eventual damages were major determinants of the spatial dominance of the turf, and of the low diversity of species observed in the study area. Overall, results from ﬁeld experiments indicate that effects of sediments on rocky coast assemblages are complex because they are probably the result of both direct effects acting on individual species and indirect effects through mediation of competitive and/or predator/prey outcomes. Furthermore, the responses of species to variations in the characteristics of the depositional environment vary with changes in the scale of observation and are inﬂuenced by a number of factors acting at different spatial and temporal scales. Consideration of scale is particularly important because it emphasises the need for caution when trying to extrapolate results observed at the relatively small spatial scales of most ﬁeld experiments to larger spatial scales relevant to predict threshold levels of disturbance by sediments in coastal areas.

Modelling Limited work has been done to quantify the magnitude of the effects that different sedimentation regimes have on rocky coast organisms, and to predict threshold levels of perturbation by sediments. For example, deposits of sediment above 10 mg cm2 and above 3 mg cm2 have been shown to inhibit settlement of Macrocystis pyrifera (Devinny & Volse 1978) and Ecklonia cava and Undaria pinnatiﬁda (Arakawa & Matsuike 1992), respectively. Results of ﬁeld observations and experiments suggest that prevalence and monopolisation of space by ﬁlamentous, turf-forming algae, might be favoured by chronic, “moderately high” (i.e. up to about 15 mg cm2 d1) rates of sedimentation (Airoldi et al. 1996, Airoldi 1998, Airoldi & Virgilio 1998). Overall, in agreement with indications from coral reef habitats (Rogers 1990), available information suggests that chronic rates of sedimentation 10 mg cm2 d1 may be considered potentially stressful for rocky coast organisms, but generalisations are premature. Even fewer attempts have been made to model and predict the effects of enhanced sediment load, as a consequence of natural or human processes, on individual species or assemblages. Based on laboratory experiments and numerical simulations, Arakawa & Morinaga (1994) predicted a reduction to 10% of the extension of beds of the kelp Ecklonia cava at levels of turbidity 10 mg l1. Based on computer simulations (using a hydrodynamic model to calculate sediment concentrations) of the effects of spillage of sediments and release of nutrients as a consequence of the construction of a bridge between 220

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Denmark and Sweden, Bach et al. (1993) predicted that up to 90% of the eelgrass meadow and 50% of macroalgae in the area would be seriously affected by concentrations of suspended sediments of 5 mg l1 to 20 mg l1, with losses of biomass up to more than 30%. However, there appear to be no subsequent veriﬁcations of the predictions. Developing and testing sound quantitative models about the consequences of changes in the regime of perturbation by sediments on rocky shore assemblages is undoubtedly most urgent.

Discussion and conclusions Sedimentation has long been acknowledged as a major determinant of the composition, distribution and diversity of rocky coast organisms. Despite the early recognition of this possibility, the present review highlights the rarity with which this hypothesis has been assessed with informative quantitative observations and speciﬁcally targeted research to quantify the effects of sedimentation and identify the underlying mechanisms has been surprisingly scarce (Fig. 5). In the past few years there has been an increasing number of quantitative and experimental studies investigating the effects of sedimentation on temperate rocky coast assemblages (Tables 2 and 3). However, many questions still remain unanswered and present knowledge makes it difﬁcult to formulate predictions of the effects of sedimentation on individual species and assemblages on rocky coasts. In the following sections available information on the ecological role of sedimentation on rocky coasts is synthesised, present inability to predict responses of individual species and assemblages to the threshold sediment loads is addressed and an attempt is made to identify emerging general trends. Those factors that most hinder generalisations and predictions are highlighted, including limited knowledge of the underlying mechanisms, insufﬁcient description of the regime of perturbation by sediments, and scarce consideration of scale issues. A critical discussion of these limitations is important for the effective planning of future work.

Ecological role of sedimentation There is substantial evidence that the ecological role of sedimentation in rocky coasts is one of major signiﬁcance. Sedimentation is an important factor of stress and disturbance for hard-bottom organisms (sensu Grime 1977). As with other natural and anthropogenic perturbations, presence of sediments may deeply affect the composition, structure, dynamics and diversity of natural assemblages (e.g. Daly & Mathieson 1977, Littler et al. 1983, D’Antonio 1986, Kendrick 1991, Airoldi 1998, Irving & Connell 2002a,b,), and may play a role in the evolution of life histories (Brown 1996). Sediments that accumulate on rocky substrata can cause burial, scour and profound modiﬁcations to the characteristics of the bottom surface, and interact with other physical and biological processes, including grazing and predation, water motion, turbidity, substratum topography and pollutants (see p. 170). The degree, extent, location, frequency and duration of burial and/or scour, and the characteristics of sediment particles (e.g. grain size, shape, density, mineral and chemical composition) are all important in determining the regime of perturbation by sediments and its ecological consequences. Interactions between sediments and organisms on rocky coasts are complex (Fig. 13). 221

LAURA AIROLDI

Many species accumulate and trap sediments, thus controlling their transport, deposition and accrual rates (see pp. 168 and 187). In turn, sediments signiﬁcantly affect the abundance and distribution of hard-bottom organisms by limiting the abundance of some species and favouring the development of others (Tables 2 and 3). The underlying mechanisms of these processes are little known, and often ecologists have referred to the effects of “sedimentation” ambiguously, without explicit consideration of the different components and effects of “sedimentation” (e.g. burial, scour, turbidity). Observations and experiments suggest that sediments affect rocky coast organisms through both direct effects (generally negative, such as smothering, scour, replacement of stable with unstable substrata) on settlement, recruitment, growth or survival of individual species (see pp. 189 and 196), and indirect effects (positive and/or negative) through mediation of competitive and predator–prey outcomes (e.g. Taylor & Littler 1982, Littler et al. 1983, D’Antonio 1986, Airoldi & Cinelli 1997). Effects vary over space and time, depending on the characteristics of the depositional environment, life histories of species and the stage of development of individuals and assemblages, and in relation to variable physical factors, including hydrodynamics, light intensity and bottom topography (e.g. Airoldi & Cinelli 1997, Airoldi & Virgilio 1998, Irving & Connell 2002a,b).

Can we predict the impacts of sediments? Identifying the magnitude of the effects that different sedimentation regimes have on individual species and assemblages and the critical levels above which detrimental effects become manifest is a major requirement for predicting the impacts of sediments and for effective management of rocky shore habitats. Observations and laboratory experiments suggest that even the most tolerant hard-bottom organisms would eventually suffer inhibition and mortality above certain degrees of sedimentation (see pp. 189 and 196). Furthermore, there is evidence that “excessive” sediment load can be a threat to the diversity and functioning of rocky coast assemblages, and a prime initiator of shifts between alternate states in the composition of species (see pp. 176, 193 and 207). Paucity of quantitative data and poor understanding of the mechanisms by which sediments interact with rocky shore organisms limit our present ability to predict the effects of enhanced sediment loads on rocky coast assemblages. Nevertheless, several common patterns emerge from the body of literature that is reviewed here, and a few qualitative trends may be tentatively suggested: (1)

(2)

(3)

Rocky coast organisms that persist by sexual reproduction appear to be more vulnerable to the presence of sediments than organisms that propagate vegetatively, probably because larvae and propagules require stable substrata for settlement, and/or juvenile stages are more sensitive to smothering by sediments than adult stages; There seems to be a trend in sediment affected areas for the prevalence of species with sediment-trapping morphologies, opportunistic, vegetative propagating or migratory life histories and physiological and morphological adaptation to withstand stressful physical and chemical conditions during burial. Many of these species can probably be characterised as “sand-tolerant” species, for which negative effects due to the presence of sediments are possibly compensated for by indirect advantages, including reduced competition and predation. Low density of grazers and concomitant dominance of turf-forming and/or 222

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

(4)

(5)

opportunistic foliose algae frequently characterise rocky coasts affected by sediments, suggesting that sediments may control rocky coast vegetation through inhibition of grazing. Areas affected by sediments appear to be frequently characterised by low diversity of species, often because of the prevalence of space-monopolising forms. At the same time, however, variable patterns of sediment deposition and movement may be important sources of spatial and temporal heterogeneity in the structure and dynamics of affected assemblages, sometimes promoting diversity. There seem to be trends in areas with high human perturbations, including high sediment load, for the decline in cover of erect, canopy-forming algae and increased abundance of turf-forming algae. The latter, once established, trap sediments and seem to inhibit reinvasion of canopies and other organisms. The underlying mechanisms for the prevalence of turf-forming algae and their tendency to monopolise space in sediment impacted areas are not fully clariﬁed but recent research has suggested that the abilities to entrap and withstand sediments and to pre-empt space by propagating vegetatively are major determinants of the success of these species. It can be further speculated (but evidence either supporting or refuting this hypothesis is limited) that assemblages dominated by canopy-forming and turf-forming algae might represent alternative stable states in shallow temperate rocky reefs, and that sediments might be one of the factors triggering the shift in balance between those two states.

Veriﬁcation of any of the above scenarios requires quantitative and experimental work, including large-scale temporal and geographic comparisons among systems with different regimes of stress and disturbance by sediments. Such examination is only possible if sedimentation is quantiﬁed with comparable methods and over a range of spatial and temporal scales, that are relevant to the ecological processes being examined.

The overlooked importance of scale Although the problem of scale has been recognised as a central issue in ecological studies (e.g. Petraitis et al. 1989, Levin 1992), consideration of scale has been surprisingly limited in studies on the impacts of sedimentation on rocky coast assemblages, and analogous concern has been raised by Rogers (1990) for studies done in coral reefs. The characteristics of the depositional environment, the attributes of rocky coast habitats and assemblages, and the interactions between the two are highly variable over a range of spatial and temporal scales. Thus, the perception of coupling between sedimentation and rocky coast assemblages is inﬂuenced by the spatial and temporal extent of a study (Airoldi & Virgilio 1998). However, in many cases, effects of sedimentation have been interpreted and generalised in the absence of data on the regime of perturbation by sediments undergone by the assemblages and there is a lack of information on the spatial and temporal variability in both patterns of deposition of sediments and benthic assemblages. Very few studies have been done at more than one place or time and most observations and experiments have been undertaken at the scale of individual organisms or small habitat patches. The effects of changes in the regimes of sedimentation over large spatial scales have rarely been addressed and observations were often confounded by the concomitant variations of other physical, chemical or biological parameters. Furthermore, most studies on the effects of sedimentation on rocky 223

LAURA AIROLDI

coast assemblages were carried out over very short times. Observations and experiments indicate, however, that timing is a critical and overlooked factor in studies of the effects of sedimentation on rocky coast assemblages (e.g. Shaffer & Parks 1994, Airoldi & Cinelli 1997, Airoldi 1998). Current knowledge is limited and does not allow conclusions about the range of scales that are relevant to interactions between sediments and rocky coast assemblages. However, there are a few studies that have identiﬁed important spatial and temporal components of variability. For example, it has been demonstrated that the effects of disturbance by sediments can vary depending on the time of the year when disturbance occurs, or on the location and topography of the shore (e.g. Littler et al. 1983, Trowbridge 1996, Airoldi 2000b). The sub-lethal chronic presence of a moderate layer of sediment has been shown to have different effects on the diversity of rocky coast organisms than sporadic, severe burial or scour (e.g. Shaffer & Parks 1994, Airoldi 1998). Detrimental effects on growth of some species observed at the small spatial scales of habitat patches have been found to weaken if not be reversed at larger scales of sites and shores (Airoldi & Virgilio 1998). Similarly, there are indications that the heterogeneous and unpredictable distribution of sediments affects the small-scale patchiness of the environment (Littler et al. 1983, McQuaid & Dower 1990, Airoldi & Cinelli 1997, Airoldi 1998), but the consequences at the scale of a whole shore (e.g. in terms of control of species diversity) are less clear. These results stress the need for explicit consideration of scale issues in future studies.

Conclusions and future research needs Substantial progress has been made in the past few decades in detecting effects of sedimentation in coastal environments. However, we still know little about how individual species and assemblages on rocky coasts respond to spatial and temporal changes in the characteristics of the regime of sedimentation, and about the direct and indirect mechanisms by which sediments affect rocky coast organisms. Current limitations to knowledge can be attributed largely to a paucity of quantitative and experimental research, and especially to the scant attention devoted to quantitative measures of the regime of perturbation by sediments and responses of organisms at relevant spatial and temporal scales. Whereas limited information exists, our ability to make generalisations is restricted. Predicting the consequences of changes in sediment loads and the critical levels above which detrimental effects of sediments become manifest remains a key issue for the ecology of rocky shores and a challenge for future studies. There is a need for rigorous research, with a meaningful experimental component, to quantify the effects of sedimentation on individual species and assemblages, clarify the underlying direct and indirect causal mechanisms, and identify possible interactions with other environmental factors, including hydrodynamic conditions, substratum topography, organic and chemical pollutants, and water turbidity. There is a need for comparable estimates of the regime of perturbation by sediments across different habitats and locations, in order to interpret which are the levels of sedimentation that should be considered as “high” or “low” for rocky coasts. Such a comparison is only possible if standardised methodologies or sets of methodologies are used, and especially if the temporal and spatial context of any study are explicit. There is probably also a need to be more explicit about what is meant by “sedimentation” (i.e. sediment deposition, accumulation or movement, scour, or turbidity), 224

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

because the effects of sediments are complex, variable over space and time and nonindependent from a variety of physical and biological factors. Finally, there is much demand for studies over large spatial and temporal scales, because the scales of impact of enhanced sediment load as a consequence of human activities are much larger than those that can be covered by laboratory or ﬁeld experiments. Extrapolating results from small to larger scales is still a major problem and challenge in ecology (Gardner et al. 2001). Large-scale and long-term monitoring programmes would be very useful in helping to solve these problems, because they would provide fundamental baseline information about trends of changes in sediment loads and assemblages, identify whether the causes of change are to be attributed to natural processes or human activities, and possibly take advantage of unplanned experiments that occur as a consequence of management actions (Carpenter et al. 1995).

Acknowledgements The ideas presented in this review have been stimulated by discussions with my colleague and friend G. A Kendrick. The review was also encouraged by my attendance at two meetings, the annual conference of the Italian Ecological Society held in Pisa in 2000 and the European Marine Biological Symposium held in Minorca in 2001. I am most grateful to R. N. Gibson for inviting me to write this review, and to M. Abbiati, S. J. Hawkins, and A. J. Southward for their unfailing support. I wish to sincerely thank S. D. Connell, K. Hiscock, and A. J. Southward for inputs of ideas and insightful criticisms of my drafts. I also wish to thank A. S. Chapman, S. D. Connell, A. Pulfrich and their coauthors for putting at my disposal their manuscripts in press, the authors and publisher who agreed the reproduction of some ﬁgures from their works, F. Arenas for lending his artistry to Figure 1, and J. Anderson for help with species nomenclature. I am most grateful to the Marine Biological Association of the UK for hospitality and support while writing, and in particular to L. Noble and all the staff of the library for their invaluable assistance with library searches. Finally, I wish to thank my colleagues and students for their patient support while writing. The work was supported by a LINKECOL Exchange Grant from the European Science Foundation and from an Assegno di Ricerca of the University of Bologna.

References Airoldi, L. 1998. Roles of disturbance, sediment stress, and substratum retention on spatial dominance in algal turf. Ecology 79, 2759–2770. Airoldi, L. 2000a. Effects of disturbance, life histories, and overgrowth on coexistence of algal crusts and turfs. Ecology 81, 798–814. Airoldi, L. 2000b. Responses of algae with different life histories to temporal and spatial variability of disturbance in subtidal reefs. Marine Ecology Progress Series 195, 81–92. Airoldi, L. 2001. Distribution and morphological variation of low-shore algal turfs. Marine Biology 138, 1233–1239. Airoldi, L. & Cinelli, F. 1996a. Variability of ﬂuxes of particulate material in a submarine cave with chemolithoautotrophic inputs of organic carbon. Marine Ecology Progress Series 139, 205–217. Airoldi, L. & Cinelli, F. 1996b. Early patterns of recovery of a ﬁlamentous algal turf on a rocky

225

LAURA AIROLDI

subtidal shore (Mediterranean Sea). In Atti del VII Congresso Nazionale della Società Italiana di Ecologia, Napoli, 11–14 Settembre 1996, A. Virzo de Santo et al. (eds). Società Italiana di Ecologia, Atti 17, 341–344. Airoldi, L. & Cinelli, F. 1997. Effects of sedimentation on subtidal macroalgal assemblages: an experimental study from a Mediterranean rocky shore. Journal of Experimental Marine Biology and Ecology 215, 269–288. Airoldi, L., Fabiano, M. & Cinelli, F. 1996. Sediment deposition and movement over a turf assemblage in a shallow rocky coastal area on the Ligurian Sea. Marine Ecology Progress Series 133, 241–251. Airoldi, L., Rindi, F. & Cinelli, F. 1995. Structure, seasonal dynamics and reproductive phenology of a ﬁlamentous turf assemblage on a sediment inﬂuenced, rocky subtidal shore. Botanica Marina 38, 227–237. Airoldi, L. & Virgilio, M. 1998. Responses of turf-forming algae to spatial variations in the deposition of sediments. Marine Ecology Progress Series 165, 271–282. Albrecht, A. S. 1998. Soft bottom versus hard rock: community ecology of macroalgae on intertidal mussel beds in the Wadden Sea. Journal of Experimental Marine Biology and Ecology 229, 85–109. Aleem, A. A. 1993. The marine algae of Alexandria, Egypt. Alexandria: Privately published. Amesbury, S. S. 1982. Effects of turbidity on shallow-water reef ﬁsh assemblages in Truk, Eastern Caroline Islands. In The reef and man. Proceedings of the Fourth International Coral Reef Symposium, Manila, Philippines, 1981, E. D. Gomez et al. (eds). Quezon City: University of Philippines, 155–159. Arakawa, H. & Matsuike, K. 1990. Inﬂuence on sedimentation velocity of brown algae zoospores and their base-plate insertion exerted by suspended matters. Nippon Suisan Gakkaishi 56, 1741–1748. Arakawa, H. & Matsuike, K. 1992. Inﬂuence on insertion of zoospores, germination, survival, and maturation of gametophytes of brown algae exerted by sediments. Nippon Suisan Gakkaishi 58, 619–625. Arakawa, H. & Morinaga, T. 1994. Inﬂuence of suspended particles on dispersion of brown algal zoospores. Nippon Suisan Gakkaishi 60, 61–64. Ayling, A. M. 1981. The role of biological disturbance in temperate subtidal encrusting communities. Ecology 62, 830–847. Bach, H. K., Rasmussen, E. K. & Riber, H. H. 1993. The application of an ecological model to assess the impact of sediment spill on benthic vegetation. In Proceedings of the 12th Baltic Marine Biologists Symposium, E. Bjørnestad et al. (eds). Fredensborg: Olsen & Olsen, 7–15. Bakus, G. J. 1968. Sedimentation and benthic invertebrates of Fanning Island, Central Paciﬁc. Marine Geology 6, 45–51. Ballesteros, E. 1992. Els vegetals i la zonació litoral: especies, comunitats i factors que inﬂueixem en la seva distribució. Barcelona: Institut d’Estudis Catalans. Bally, R., McQuaid, C. D. & Brown, A. C. 1984. Shores of mixed sand and rock: an unexplored marine ecosystem. South African Journal of Science 80, 500–503. Barko, J. W., Gunnison, D. & Carpenter, S. R. 1991. Sediment interactions with submersed macrophyte growth and community dynamics. Aquatic Botany 41, 41–65. Bavestrello, G., Cattaneo-Vietti, R., Cerrano, C., Danovaro, R. & Fabiano, M. 1995. Annual sedimentation rates and role of the resuspension processes along a vertical cliff (Ligurian Sea, Italy). Journal of Coastal Research 11, 690–696. Bavestrello, G., Cattaneo-Vietti, R., Danovaro, R. & Fabiano, M. 1991. Detritus rolling down a vertical cliff of the Ligurian Sea (Italy): the ecological role in hard bottom communities. P.S.Z.N.I.: Marine Ecology 12, 281–292. Baynes, T. W. 1999. Factors structuring a subtidal encrusting community in the southern Gulf of California. Bulletin of Marine Science 64, 419–450. Benedetti-Cecchi, L., Pannacciulli, F., Bulleri, F., Moschella, P. S., Airoldi, L., Relini, G. & Cinelli, F. 2001. Predicting the consequences of anthropogenic disturbance: large-scale effects of loss of canopy algae on rocky shores. Marine Ecology Progress Series 214, 137–150.

226

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Bernier, P., Guidi, J.-B. & Böttcher, M. E. 1997. Coastal progradation and very early diagenesis of ultramaﬁc sands as a result of rubble discharge from asbestos excavations (northern Corsica, western Mediterranean). Marine Geology 144, 163–175. Bertness, M. D. 1984. Habitat and community modiﬁcation by an introduced herbivorous snail. Ecology 65, 370–381. Birje, J., Verlaque, M. & Poydenot, F. 1996. Macrophytobenthos des platiers rocheux intertidaux et semi-exposés de la région de Saﬁ-Essaouira (Maroc occidental). Oceanologica Acta 19, 561–574. Bloesch, J. & Burns, N. M. 1980. A critical review of sedimentation trap technique. Schweizerische Zeitschrift fur Hydrologie 42, 15–55. Blomqvist, S. & Håkanson, L. 1981. A review on sediment traps in aquatic environments. Archiv fur Hydrobiologie 91, 101–132. Boaventura, D., Alexander, M., Della Santina, P., Smith, N. D., Ré, P., Cancela da Fonseca, L. & Hawkins, S. J. 2002. The effects of grazing on the distribution and composition of low-shore algal communities on the central coast of Portugal and on the southern coast of Britain. Journal of Experimental Marine Biology and Ecology 267, 185–206. Boney, A. D. 1978. Marine algae as collectors of iron ore dust. Marine Pollution Bulletin 9, 175–180. Boney, A. D. 1980. Mineral deposition and stratiﬁcation in tufted growths of Audouinella purpurea (Lightf.) Woelkerling. Annals of Botany 45, 713–715. Branch, G. M., Eekhout, S. & Bosman, A. L. 1990. Short-term effects of the 1988 Orange River ﬂoods on the intertidal rocky-shore communities of the open coast. Transactions of the Royal Society of Africa 47, 331–354. Brattström, H. 1992. Marine biological investigations in the Bahamas. 22. Littoral zonation at three Bahamian beachrock localities. Sarsia 77, 81–109. Brown, A. C. 1996. Intertidal rock inundated by sand as an evolutionary corridor for benthic marine invertebrates. South African Journal of Science 92, 162 only. Burrows, E. M. 1971. Assessment of pollution effects by the use of algae. Proceedings of the Royal Society of London 177, 295–306. Burrows, E. M. & Pybus, C. 1971. Laminaria saccharina and marine pollution in north-east England. Marine Pollution Bulletin 2, 53–56. Butman, C. A., Grant, W. D. & Stolzenbach, K. D. 1986. Predictions of sediment trap biases in turbulent ﬂows: a theoretical analysis based on observations from the literature. Journal of Marine Research 44, 601–644. Capuzzo, J. M., Burt, W. V., Duedall, I. W., Park, P. K. & Kester, D. R. 1985. The impact of waste disposal in nearshore environments. In Wastes in the ocean. Volume 6. Nearshore waste disposal, B. H. Ketchum et al. (eds). New York: John Wiley, 3–38. Carballo, J. L., Naranjo, S. A. & García-Gómez, J. C. 1996. Use of marine sponges as stress indicators in marine ecosystems at Algeciras Bay (southern Iberian Peninsula). Marine Ecology Progress Series 135, 109–122. Carpenter, S. R., Chisholm, S. W., Krebs, C. J., Schindler, D. W. & Wright, R. F. 1995. Ecosystem experiments. Science 269, 324–327. Carpenter, R. C. & Williams, S. L. 1993. Effects of algal turf canopy height and microscale substratum topography on proﬁles of ﬂow speed in a coral forereef environment. Limnology and Oceanography 38, 687–694. Castilla, J. C. & Nealler, E. 1978. Marine environmental impact due to mining activities of El Salvador copper mine, Chile. Marine Pollution Bulletin 9, 67–70. Castric-Fey, A. 1988. Les facteurs limitants des peuplements sessiles sublittoraux en Baie de Concarneau (sud-Finistere). Vie Milieu 38, 1–18. Chapman, A. S. & Fletcher, R. L. 2002. Differential effects of sediments on survival and growth of Fucus serratus embryos (Fucales, Phaeophyceae). Journal of Phycology 38, 894–903. Chapman, V. J. 1943. Zonation of marine algae on the sea-shore. Proceedings of the Linnean Society of London 154, 239–253.

227

LAURA AIROLDI

Chou, L. M. 1996. Response of Singapore reefs to land reclamation. Galaxea 13, 85–92. Clarke, W. D. & Neushul, M. 1967. Subtidal ecology of the southern California coast. In Pollution and marine ecology, T. A. Olson & F. J. Burgess (eds). New York: Interscience, 29–42. Cortés, J. & Risk, M. J. 1985. A reef under siltation stress: Cahuita, Costa Rica. Bulletin of Marine Science 36, 339–356. Cotton, A. D. 1912. Clare Island Survey. Part 15. Marine Algae. Proceedings of the Royal Irish Academy 31, 1–178. Craik, G. J. S. 1980. Simple method for measuring the relative scouring of intertidal areas. Marine Biology 59, 257–260. Crothers, J. H. & Hayns, S. 1994. Rocky shore distribution patterns along the Somerset coast. Biological Journal of the Linnean Society 51, 115–121. Dahl, A. L. 1971. Development, form and environment in the brown alga Zonaria farlowii (Dictyotales). Botanica Marina 14, 76–112. Daly, M. A. & Mathieson, A. C. 1977. The effects of sand movement on intertidal seaweeds and selected invertebrates at Bound Rock, New Hampshire, USA. Marine Biology 43, 45–55. D’Antonio, C. M. 1986. Role of sand in the domination of hard substrata by the intertidal alga Rhodomela larix. Marine Ecology Progress Series 27, 263–275. Dayton, P. K. 1973. Dispersion, dispersal, and persistence of the annual intertidal alga, Postelsia palmaeformis Ruprecht. Ecology 54, 433–438. Dayton, P. K. 1975. Experimental evaluation of ecological dominance in a rocky intertidal algal community. Ecological Monographs 45, 137–159. Dethier, M. N. 1994. The ecology of intertidal algal crusts: variation within a functional group. Journal of Experimental Marine Biology and Ecology 177, 37–71. Devinny, J. S. & Volse, L. A. 1978. Effects of sediments on the development of Macrocystis pyrifera gametophytes. Marine Biology 48, 343–348. Deysher, L. E. & Dean, T. A. 1986. In situ recruitment of sporophytes of the giant kelp Macrocystis pyrifera (L.) C. A. Agardh: effects of physical factors. Journal of Experimental Marine Biology and Ecology 103, 41–63. Dodge, R. E. & Szmant-Froelich, A. 1985. Effects of drilling ﬂuids on reef corals: a review. In Wastes in the ocean. Volume 4. Energy wastes in the ocean, I. W. Duedall et al. (eds). New York: John Wiley, 341–364. Doty, M. S. 1947a. The marine algae of Oregon. Part I. Chlorophyta and Phaeophyta. Farlowia 3, 1–65. Doty, M. S. 1947b. The marine algae of Oregon. Part II. Rhodophyta. Farlowia 3, 159–215. Duggins, D. O., Eckman, J. E. & Sewell, A. T. 1990. Ecology of understory kelp environments. II. Effects of kelps on recruitment of benthic invertebrates. Journal of Experimental Marine Biology and Ecology 143, 27–45. Eagle, R. A., Hardiman, P. A., Norton, M. G., Nunny, R. S. & Rolfe, M. S. 1979. The ﬁeld assessment of effects of dumping wastes at sea: 5. The disposal of solid wastes off the north-east coast of England. Lowestoft: MAFF Fisheries Research Technical Report, Report No. 51. Ebeling, A. W., Laur, D. R. & Rowley, R. J. 1985. Severe storm disturbances and reversal of community structure in a southern California kelp forest. Marine Biology 84, 287–294. Eckman, J. E. & Duggins, D. O. 1991. Life and death beneath macrophyte canopies: effects of understory kelps on growth rates and survival of marine, benthic suspension feeders. Oecologia 87, 473–487. Eckman, J. E., Duggins, D. O. & Sewell, A. T. 1989. Ecology of understorey kelp environments. I. Effects of kelps on ﬂow and particle transport near the bottom. Journal of Experimental Marine Biology and Ecology 129, 173–187. Ellis, D. V. 1988. Case histories of coastal and marine mines. In Chemistry and biology of solid waste: dredged material and mine tailing, W. Salomons & U. Forstner (eds). Berlin: Springer, 73–100. Ellison, J. C. 1998. Impacts of sediment burial on mangroves. Marine Pollution Bulletin 37, 420–426.

228

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Emerson, S. E. & Zedler, J. B. 1978. Recolonization of intertidal algae: an experimental study. Marine Biology 44, 315–324. Engledow, H. R. & Bolton, J. J. 1994. Seaweed -diversity within the lower eulittoral zone in Namibia: the effects of wave action, sand inundation, mussels and limpets. Botanica Marina 37, 267–276. Eriksson, B. K., Johansson, G. & Snoeijs, P. 2002. Long-term changes in the macroalgal vegetation of the inner Gullmar Fjord, Swedish Skagerrak coast. Journal of Phycology 38, 284–296. Espinoza, J. & Rodriguez, H. 1987. Seasonal phenology and reciprocal transplantation of Sargassum sinicola Setchell et Gardner in the southern Gulf of California. Journal of Experimental Marine Biology and Ecology 110, 183–195. Estes, J. A. & Palmisano, J. F. 1974. Sea otters: their role in structuring nearshore communities. Science 185, 1058–1060. Evans, R. A., Gulliksen, B. & Sandnes, O. K. 1980. The effect of sedimentation on rocky bottom organisms in Balsfjord, Northern Norway. In Fjord Oceanography, H. J. Freeland et al. (eds). New York: Plenum Publishing, 603–607. Evans, S. M., Gill, M. E., Hardy, F. G. & Seku, F. O. K. 1993. Evidence of change in some rocky shore communities on the coast of Ghana. Journal of Experimental Marine Biology and Ecology 172, 129–141. Falace, A. & Bressan, G. 1995. Some observations on algal colonisation of artiﬁcial structures situated in the proximity of underwater pipes off Lignano-Grado (north Adriatic Sea). In Proceedings of the 30th European Marine Biological Symposium, Southampton, UK, September 1995, L. E. Hawkins et al. (eds). Southampton: University of Southampton, 313–318. Fariña, J. M. & Castilla, J. C. 2001. Temporal variation in the diversity and cover of sessile species in rocky intertidal communities affected by copper mine tailings in northern Chile. Marine Pollution Bulletin 42, 554–568. Farrow, G. E., Syvitski, J. P. M. & Tunnicliffe, V. 1983. Suspended particulate loading on the macrobenthos in a highly turbid fjord: Knight Inlet, British Columbia. Canadian Journal of Fisheries and Aquatic Sciences 40 (Suppl. 1), 273–288. Fonseca, M. S. & Fisher, J. S. 1986. A comparison of canopy friction and sediment movement between four species of seagrass with reference to their ecology and restoration. Marine Ecology Progress Series 29, 15–22. Fornos, J. J., Forteza, V., Jaume, C. & Martinez-Taberner, A. 1992. Present-day Halimeda carbonate sediments in temperate Mediterranean embayments: Fornels, Balearic Islands. Sedimentary Geology 75, 283–293. Foster, M. S. 1975. Algal succession in a Macrocystis pyrifera forest. Marine Biology 32, 313–329. French, P. W. 1997. Coastal and estuarine management. London: Routledge. Gabriele, M., Bellot, A., Gallotti, D. & Brunetti, R. 1999. Sublitoral hard substrate communities of the northern Adriatic Sea. Cahiers de Biologie Marine 40, 65–76. Gacia, E. & Duarte, C. M. 2001. Sediment retention by a Mediterranean Posidonia oceanica meadow: the balance between deposition and resuspension. Estuarine, Coastal and Shelf Science 52, 505–514. Gardner, R. H., Kemp, W. M., Kennedy, V. S. & Petersen, J. E. (eds) 2001. Scaling relations in experimental ecology. New York: Columbia University Press. Gardner, W. D. 1980. Field assessment of sediment traps. Journal of Marine Research 38, 41–52. Gibbons, M. J. 1988. The impact of sediment accumulations, relative habitat complexity and elevation on rocky shore meiofauna. Journal of Experimental Marine Biology and Ecology 122, 225–241. Gillanders, B. M. & Kingsford, M. J. 2002. Impact of changes in ﬂow of freshwater on estuarine and open coastal habitats and the associated organisms. Oceanography and Marine Biology: an Annual Review 40, 233–309. Gorostiaga, J. M. & Díez, I. 1996. Changes in the sublittoral benthic marine macroalgae in the polluted area of Abra de Bilbao and proximal coast (northern Spain). Marine Ecology Progress Series 130, 157–167.

229

LAURA AIROLDI

Gorostiaga, J. M., Santolaria, A., Secilla, A. & Díez, I. 1998. Sublittoral benthic vegetation of the Eastern Basque Coast (N. Spain): structure and environmental factors. Botanica Marina 41, 455–465. Gotelli, N. J. 1988. Determinants of recruitment, juvenile growth, and spatial distribution of a shallowwater gorgonian. Ecology 69, 157–166. Grahame, J. & Hanna, F. S. 1989. Factors affecting the distribution of the epiphytic fauna of Corallina ofﬁcinalis (L.) on an exposed rocky shore. Ophelia 30, 113–129. Grime, J. P. 1977. Evidence for the existence of three primary strategies in plants and its relevance to ecological and evolutionary theory. American Naturalist 111, 1169–1194. Guidetti, P. 2001. Detecting environmental impacts on the Mediterranean seagrass Posidonia oceanica (L.) Delile: the use of reconstructive methods in combination with “beyond BACI” designs. Journal of Experimental Marine Biology and Ecology 260, 27–39. Guiry, M. D. & Dhonncha, E. N. 2002. AlgaeBase. World wide web electronic publication. Online. Available HTTP: http://www.algaebase.com (accessed 30 May 2002). Gulliksen, B. 1982. Sedimentation close to a near vertical rocky wall in Balsfjorden, northern Norway. Sarsia 67, 21–27. Henley, W. F., Patterson, M. A., Neves, R. J. & Lemly, A. D. 2000. Effects of sedimentation and turbidity on lotic food webs: a concise review for natural resource managers. Reviews in Fisheries Science 8, 125–139. Herrnkind, W. F., Butler, M. J. & Tankersley, R. A. 1988. The effects of siltation on recruitment of spiny lobsters, Panulirus argus. Fishery Bulletin 86, 331–338. Hiscock, K. 1983. Water movement. In Sublittoral ecology. The ecology of the shallow subtidal benthos, R. Earll & D. G. Erwin (eds). Oxford: Clarendon Press, 58–96. Holmer, M., Lassus, P., Stewart, J. E. & Wildish, D. J. 2001. ICES symposium on environmental effects of mariculture. Introduction. ICES Journal of Marine Science 58, 363–368. Hoyt, W. D. 1920. Marine algae of Beaufort, N. C. and adjacent regions. Bulletin of the Bureau of Fisheries, Washington, D.C. 36, 367–556. Hunter, C. L. & Evans, C. W. 1995. Coral reefs in Kaneohe Bay, Hawaii: two centuries of western inﬂuence and two decades of data. Bulletin of Marine Science 57, 501–515. Hyslop, B. T. & Davies, M. S. 1998. Evidence for abrasion and enhanced growth of Ulva lactuca L. in the presence of colliery waste particles. Environmental Pollution 101, 117–121. Hyslop, B. T., Davies, M. S., Arthur, W., Gazey, N. J. & Holroyd, S. 1997. Effects of colliery waste on littoral communities in north-east England. Environmental Pollution 96, 383–400. Iannuzzi, T. J., Weinstein, M. P., Sellner, K. G. & Barrett, J. C. 1996. Habitat disturbance and marina development: an assessment of ecological effects. I. Changes in primary production due to dredging and marina construction. Estuaries 19, 257–271. Irving, A. D. & Connell, S. D. 2002a. Sedimentation and light penetration interact to maintain heterogeneity of subtidal habitats: algal versus invertebrate dominated assemblages. Marine Ecology Progress Series 245, 83–91. Irving, A. D. & Connell, S. D. 2002b. Interactive effects of sedimentation and microtopography on the abundance of subtidal turﬁng algae. Phycologia 41, 517–522. Jackson, J. B. C. 1977. Competition on marine hard substrata: the adaptive signiﬁcance of solitary and colonial strategies. American Naturalist 111, 743–767. Johannes, R. E. 1975. Pollution and degradation of coral reef communities. In Tropical marine pollution, E. J. F. Wood & R. E. Johannes (eds). Oxford: Elsevier Scientiﬁc, 1–181. Johansson, G., Eriksson, B. K., Pedersén, M. & Snoeijs, P. 1998. Long-term changes of macroalgal vegetation in the Skagerrak area. Hydrobiologia 385, 121–138. Jørgensen, L. L. & Gulliksen, B. 2001. Rocky bottom fauna in arctic Kongsfjord (Svalbard) studied by means of suction sampling and photography. Polar Biology 24, 113–121. Kamermans, P., Malta, E.-J., Verschuure, J. M., Lentz, L. F. & Schrijvers, L. 1998. Role of cold resistance and burial for winter survival and spring initiation of an Ulva spp. (Chlorophyta) bloom in a eutrophic lagoon (Veerse Meer lagoon, The Netherlands). Marine Biology 131, 45–51.

230

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Kendrick, G. A. 1991. Recruitment of coralline crusts and ﬁlamentous turf algae in the Galapagos archipelago: effect of simulated scour, erosion and accretion. Journal of Experimental Marine Biology and Ecology 147, 47–63. Kennelly, S. J. 1983. An experimental approach to the study of factors affecting algal colonization in a sublittoral kelp forest. Journal of Experimental Marine Biology and Ecology 68, 257–276. Kennelly, S. J. 1989. Effects of kelp canopies on understorey species due to shade and scour. Marine Ecology Progress Series 50, 215–224. Kennelly, S. J. 1991. Caging experiments to examine the effects of ﬁshes on understorey species in a sublittoral kelp community. Journal of Experimental Marine Biology and Ecology 147, 207–230. Kim, K. Y., Choi, T. S., Huh, S. H. & Garbary, D. J. 1998. Seasonality and community structure of subtidal benthic algae from Daedo Island, Southern Korea. Botanica Marina 41, 357–365. King, R. J. & Farrant, P. A. 1987. The phenology of the Dictyotales (Phaeophyceae) at a sheltered locality in Sydney Harbour, New South Wales, Australia. Botanica Marina 30, 341–350. Kinne, O. (ed.) 1971. Marine Ecology. Volume I. Environmental factors. Part 2. Chapter 6: Turbidity. London: Wiley Interscience. Kitching, J. A., Macan, T. T. & Gilson, H. C. 1934. Studies in sublitoral ecology. I. A submarine gully in Wembury Bay, South Devon. Journal of the Marine Biological Association of the United Kingdom 19, 677–705. Konar, B. & Roberts, C. 1996. Large scale landslide effects on two exposed rocky subtidal areas in California. Botanica Marina 39, 517–524. Larcombe, P. & Woolfe, K. J. 1999. Increased sediment supply to the Great Barrier Reef will not increase sediment accumulation at most coral reefs. Coral Reefs 18, 163–169. Lenihan, H. S. 1999. Physical-biological coupling on oyster reefs: how habitat structure inﬂuences individual performance. Ecological Monographs 69, 251–275. Levin, S. A. 1992. The problem of pattern and scale in ecology. Ecology 73, 1943–1967. Lewis, J. R. 1964. The ecology of rocky shores. London: English Universities Press. Lilly, S. J., Sloane, J. F., Bassindale, R., Ebling, F. J. & Kitching, J. A. 1953. The ecology of the Lough Ine rapids with special reference to water currents. Journal of Animal Ecology 22, 87–122. Lim, G. S. Y. & Chou, L. M. 1988. The echinoderm fauna of sediment stressed reefs in Singapore. In Proceedings of the 6th International Coral Reef Symposium, Townsville, Australia, J. H. Choat et al. (eds). Sixth International Coral Reef Symposium Executive Committee, 245–250. Little, C. & Smith, L. P. 1980. Vertical zonation on rocky shores in the Severn Estuary. Estuarine and Coastal Marine Science 11, 651–669. Littler, M. M. 1973. The population and community structure of Hawaiian fringing-reef crustose Corallinaceae (Rhodophyta, Cryptonemiales). Journal of Experimental Marine Biology and Ecology 11, 103–120. Littler, M. M., Littler, D. S. & Lapointe, B. E. 1988. A comparison of nutrient- and light-limited photosynthesis in psammophytic versus epilithic forms of Halimeda (Caulerpales, Halimedaceae) from the Bahamas. Coral Reefs 6, 219–225. Littler, M. M., Martz, D. R. & Littler, D. S. 1983. Effects of recurrent sand deposition on rocky intertidal organisms: importance of substrate heterogeneity in a ﬂuctuating environment. Marine Ecology Progress Series 11, 129–139. Littler, M. M. & Murray, S. N. 1975. Impact of sewage on the distribution, abundance and community structure of rocky intertidal macro-organisms. Marine Biology 30, 277–291. Lumb, C. M. 1990. Algal depth distributions and long-term turbidity changes in the Menai Strait, North Wales. Progress in Underwater Science 15, 85–99. Lund-Hansen, L. C., Valeur, J., Pejrup, M. & Jensen, A. 1997. Sediment ﬂuxes, re-suspension and accumulation rates at two wind-exposed sites and in a sheltered bay. Estuarine, Coastal and Shelf Science 44, 521–531. Lyngby, J. E. & Mortensen, S. M. 1996. Effects of dredging activities on growth of Laminaria saccharina. P.S.Z.N.I.: Marine Ecology 17, 345–354.

231

LAURA AIROLDI

MacDonald, L. H., Anderson, D. M. & Dietrich, W. E. 1997. Paradise threatened: land use and erosion on St. John, US Virgin Islands. Environmental Management 21, 851–863. Markham, J. W. 1968. Studies on the haptera of Laminaria sinclairii (Harvey) Farlow, Anderson et Eaton. Syesis 1, 125–131. Markham, J. W. 1972. Distribution and taxonomy of Laminaria sinclairii and L. longipes (Phaeophyceae, Laminariales). Phycologia 11, 147–157. Markham, J. W. 1973. Observations on the ecology of Laminaria sinclairii on three northern Oregon beaches. Journal of Phycology 9, 336–341. Markham, J. W. & Newroth, P. R. 1972. Observations on the ecology of Gymnogongrus linearis and related species. In Proceedings of the Seventh International Seaweed Symposium, Sapporo, Japan, S. Arasaki et al. (eds). Japan: University of Tokyo Press, 127–130. Marshall, D. J. & McQuaid, C. D. 1989. The inﬂuence of respiratory responses on the tolerance to sand inundation of the limpets Patella granularis L. (Prosobranchia) and Siphonaria capensis Q. et G. (Pulmonata). Journal of Experimental Marine Biology and Ecology 128, 191–201. Mathieson, A. C. 1967. Morphology and life history of Phaeostrophion irregulare S. et G. Nova Hedwigia 13, 293–318. Mathieson, A. C. 1979. Vertical distribution and longevity of subtidal seaweeds in northern New England, U. S. A. Botanica Marina 22, 511–520. Mathieson, A. C. 1982. Field ecology of the brown alga Phaeostrophion irregulare Setchell et Gardner. Botanica Marina 25, 67–85. Mathieson, A. C. & Fralick, R. A. 1972. Investigations of New England marine algae V. The algal vegetation of the Hapton-Seabrock Estuary and the open coast near Hampton, New Hampshire. Rhodora 74, 406–435. Matsunaga, K., Kawaguchi, T., Suzuki, Y. & Nigi, G. 1999. The role of terrestrial humic substances on the shift of kelp community to crustose coralline algae community of the southern Hokkaido Island in the Japan Sea. Journal of Experimental Marine Biology and Ecology 241, 193–205. Maughan, B.C. 2001. The effects of sedimentation and light on recruitment and development of a temperate, subtidal epifaunal community. Journal of Experimental Marine Biology and Ecology 256, 59–71. McClanahan, T. R. & Obura, D. 1997. Sedimentation effects on shallow coral communities in Kenya. Journal of Experimental Marine Biology and Ecology 209, 103–122. McGuinness, K. A. 1987a. Disturbance and organisms on boulders. I. Patterns in the environment and the community. Oecologia 71, 409–419. McGuinness, K. A. 1987b. Disturbance and organisms on boulder. II. Causes of patterns in diversity and abundance. Oecologia 71, 420–430. McQuaid, C. D. & Dower, K. M. 1990. Enhancement of habitat heterogeneity and species richness on rocky shores inundated by sand. Oecologia 84, 142–144. Meinesz, A., Lefevre, J. R. & Astier, J. M. 1991. Impact of coastal development on the infralittoral zone along the southeastern Mediterranean shore of continental France. Marine Pollution Bulletin 23, 343–347. Melville, A. J. & Connell, S. D. 2001. Experimental effects of kelp canopies on subtidal coralline algae. Austral Ecology 26, 102–108. Menge, B. A., Berlow, E. L., Blanchette, C. A., Navarrete, S. A. & Yamada, S. B. 1994. The keystone species concept: variation in interaction strength in a rocky intertidal habitat. Ecological Monographs 64, 249–286. Mettam, C. 1994. Intertidal zonation of animals and plants on rocky shores in the Bristol Channel and Severn Estuary – the northern shores. Biological Journal of the Linnean Society 51, 123–147. Miller, R. J. 1985. Succession in sea urchin and seaweed abundance in Nova Scotia, Canada. Marine Biology 84, 275–286. Moore, P. G. 1972. Particulate matter in the sublittoral zone of an exposed coast and its ecological signiﬁcance with special reference to the fauna inhabiting kelp holdfasts. Journal of Experimental Marine Biology and Ecology 10, 59–80.

232

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Moore, P. G. 1973a. The kelp fauna of northeast Britain. I. Introduction and the physical environment. Journal of Experimental Marine Biology and Ecology 13, 97–125. Moore, P. G. 1973b. The kelp fauna of northeast Britain. II. Multivariate classiﬁcation: turbidity as an ecological factor. Journal of Experimental Marine Biology and Ecology 13, 127–163. Moore, P. G. 1977. Inorganic particulate suspensions in the sea and their effects on marine animals. Oceanography and Marine Biology: an Annual Review 15, 225–363. Moran, P. J. 1991. The effects of dredging on the larval settlement and community development of fouling organisms in Port Kembla Harbour, Australia. Water Research 25, 1151–1155. Moring, J. R. 1996. Short-term changes in tidepools following two hurricanes. Hydrobiologia 328, 155–160. Moss, B., Mercer, S. & Sheader, A. 1973. Factors affecting the distribution of Himanthalia elongata (L.) S. F. Gray on the north-east coast of England. Estuarine and Coastal Marine Science 1, 233–243. Muntz, L., Norton, T. A., Ebling, F. J. & Kitching, J. A. 1972. The ecology of Lough Ine. XVIII. Factors controlling the distribution of Corynactis viridis Allman. Journal of Animal Ecology 47, 735–750. Naranjo, S. A., Carballo, J. L. & García-Gómez, J. C. 1996. Effects of environmental stress on ascidian populations in Algeciras Bay (southern Spain). Possible marine bioindicators? Marine Ecology Progress Series 144, 119–131. Neumann, A. C., Gebelein, C. D. & Scofﬁn, T. P. 1970. The composition, structure and erodability of subtidal mats, Abaco, Bahamas. Journal of Sedimentary Petrology 40, 274–297. Neushul, M., Foster, M. S., Coon, D. A., Woessner, J. W. & Harger, B. W. W. 1976. An in situ study of recruitment, growth and survival of subtidal marine algae: techniques and preliminary results. Journal of Phycology 12, 397–408. Norton, T. A. 1978. The factors inﬂuencing the distribution of Saccorhiza polyschides in the region of Lough Ine. Journal of the Marine Biological Association of the United Kingdom 58, 527–536. Norton, T. A., Hiscock, K. & Kitching, J. A. 1977. The ecology of Lough Ine. XX. The Laminaria forest at Carrigathorna. Journal of Ecology 65, 919–941. Palmer-Zwahlen, M. L. & Aseltine, D. A. 1994. Successional development of the turf community on a quarry rock artiﬁcial reef. Bulletin of Marine Science 55, 902–923. Pearson, T. H. & Rosenberg, R. 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: an Annual Review 16, 229–311. Pedersén, M. & Snoeijs, P. 2001. Patterns of macroalgal diversity, community composition and longterm changes along the Swedish west coast. Hydrobiologia 459, 83–102. Pérès, J. M. & Picard, J. 1964. Nouveau manuel de bionomie benthique de la Mer Méditerranée. Recueil des Travaux de la Station Marine d’Endoume 31, 1–137. Petraitis, P. S., Latham, R. E. & Niesenbaum, R. A. 1989. The maintenance of species diversity by disturbance. The Quarterly Review of Biology 64, 393–418. Piazzi, L., Balestri, E., Magri, M. & Cinelli, F. 1997. Expansion de l’algue tropicale Caulerpa racemosa (Forsskål) J. Agardh (Bryopsidophyceae, Chlorophyta) le long de la côte toscane (Italie). Cryptogamie, Algologie 18, 343–350. Piazzi, L. & Cinelli, F. 2001. Distribution and dominance of two introduced turf-forming macroalgae on the coast of Tuscany, Italy, northwestern Mediterranean Sea in relation to different habitats and sedimentation. Botanica Marina 44, 509–520. Pineda, J. & Escofet, A. 1989. Selective effects of disturbance on populations of sea anemones from northern Baja California, Mexico. Marine Ecology Progress Series 55, 55–62. Power, M. E. 1990. Resource enhancement by indirect effects of grazers: armored catﬁsh, algae, and sediment. Ecology 71, 897–904. Pringle, C. M. & Blake, G. A. 1994. Quantitative effects of atyid shrimp (Decapoda: Atyidae) on the depositional environment in a tropical stream: use of electricity for experimental exclusion. Canadian Journal of Fisheries and Aquatic Sciences 51, 1443–1450.

233

LAURA AIROLDI

Pulfrich, A., Parkins, C. A., Branch, G. M., Bustamante, R. H. & Velásquez, C. R. in press. The effects of sediment deposits from Namibian diamond mines on intertidal and subtidal reefs and rocklobster populations. Aquatic Conservation: Marine and Freshwater Ecosystems. Purcell, S. W. 1996. A direct method for assessing sediment load in epilithic algal communities. Coral Reefs 15, 211–213. Raimondi, P. T., Barnett, A. M. & Krause, P. R. 1997. The effects of drilling muds on marine invertebrate larvae and adults. Environmental Toxicology and Chemistry 16, 1218–1228. Reed, D. C., Laur, D. R. & Ebeling, A. W. 1988. Variation in algal dispersal and recruitment: the importance of episodic events. Ecological Monographs 58, 321–335. Rees, T. K. 1935. The marine algae of Lough Ine. Journal of Ecology 23, 69–133. Relini, G., Merello, S. E. & Torchia, G. 1998. Inﬂuenza della sedimentazione sulla colonizzazione e successione ecologica di substrati duri. Biologia Marina Mediterranea 5, 154–163. Renaud, P. E., Ambrose Jr, W. G., Riggs, S. R. & Syster, D. A. 1996. Multi-level effects of severe storms on an offshore temperate reef system: benthic sediments, macroalgae, and implications for ﬁsheries. P.S.Z.N.I.: Marine Ecology 17, 383–398. Renaud, P. E., Riggs, S. R., Ambrose Jr, W. G., Schmid, K. & Snyder, S. W. 1997. Biological-geological interactions: storm effects on macroalgal communities mediated by sediment characteristics and distribution. Continental Shelf Research 17, 37–56. Richmond, R. H. 1993. Coral reefs: present problems and future concerns resulting from anthropogenic disturbance. American Zoologist 33, 524–536. Roberts, D. E., Smith, A., Ajani, P. & Davis, A. R. 1998. Rapid changes in encrusting marine assemblages exposed to anthropogenic point-source pollution: a “Beyond BACI” approach. Marine Ecology Progress Series 163, 213–224. Robles, C. 1982. Disturbance and predation in an assemblage of herbivorous Diptera and algae on rocky shores. Oecologia 54, 23–31. Rogers, C. S. 1990. Responses of coral reefs and reef organisms to sedimentation. Marine Ecology Progress Series 62, 185–202. Round, F. E., Sloane, J. F., Ebling, F. J. & Kitching, J. A. 1961. The ecology of Lough Ine. X. The hydroid Sertularia operculata (L.) and its associated ﬂora and fauna: effects of transference to sheltered water. Journal of Ecology 49, 617–629. Ryan, P. A. 1991. Environmental effects of sediment on New Zealand streams: a review. New Zealand Journal of Marine and Freshwater Research 25, 207–221. Saiz-Salinas, J. I. & Urdangarin, I. I. 1994. Response of sublittoral hard substrate invertebrates to estuarine sedimentation in the outer harbour of Bilbao (N. Spain). P.S.Z.N.I.: Marine Ecology 15, 105–131. Saiz-Salinas, J. I. & Urkiaga-Alberdi, J. 1999. Faunal response to turbidity in a man-modiﬁed bay (Bilbao, Spain). Marine Environmental Research 47, 331–347. Santelices, B., Vásquez, J., Ohme, U. & Fonck, E. 1984. Managing wild crops of Gracilaria in central Chile. Hydrobiologia 116/117, 77–89. Santos, R. 1993. A multivariate study of biotic and abiotic relationships in a subtidal algal stand. Marine Ecology Progress Series 94, 181–190. Schroeter, S. C., Dixon, J. D., Kastendiek, J. & Smith, R. O. 1993. Detecting the ecological effects of environmental impacts: a case study of kelp forest invertebrates. Ecological Applications 3, 331–350. Scofﬁn, T. P. 1970. The trapping and binding of subtidal carbonate sediments by marine vegetation in Bimini Lagoon, Bahamas. Journal of Sedimentary Petrology 40, 249–273. Seapy, R. R. & Littler, M. M. 1982. Population and species diversity ﬂuctuations in a rocky intertidal community relative to severe aerial exposure and sediment burial. Marine Biology 71, 87–96. Seymour, R. J., Tegner, M. J., Dayton, P. K. & Parnell, P. E. 1989. Storm wave induced mortality of giant kelp, Macrocystis pyrifera in southern California. Estuarine, Coastal and Shelf Science 28, 277–292.

234

EFFECTS OF SEDIMENTATION ON ROCKY COAST ASSEMBLAGES

Sfriso, A. & Marcomini, A. 1996. Decline of Ulva growth in the lagoon of Venice. Bioresource Technology 58, 299–307. Shaffer, J. A. & Parks, D. S. 1994. Seasonal variations in and observations of landslide impacts on the algal composition of a Puget Sound nearshore kelp forest. Botanica Marina 37, 315–323. Slattery, M. & Bockus, D. 1997. Sedimentation in McMurdo Sound, Antarctica: a disturbance mechanism for benthic invertebrates. Polar Biology 18, 172–179. Smith, F. & Witman, J. D. 1999. Species diversity in subtidal landscapes: maintenance by physical processes and larval recruitment. Ecology 80, 51–69. Sousa, W. P. 2001. Natural disturbance and the dynamics of marine benthic communities. In Marine community ecology, M. D. Bertness et al. (eds). Sunderland, Massachusetts: Sinauer Associates, 85–130. Sousa, W. P., Schroeter, S. C. & Gaines, S. D. 1981. Latitudinal variation in intertidal algal community structure: the inﬂuence of grazing and vegetative propagation. Oecologia 48, 297–307. Steneck, R. S., Macintyre, I. G. & Reid, R. P. 1997. A unique algal ridge system in the Exuma Cays, Bahamas. Coral Reefs 16, 29–37. Stephenson, T. A. 1943. The causes of the vertical and horizontal distribution of organisms between tidemarks in South Africa. Proceedings of the Linnean Society of London 154, 219–233. Stewart, J. G. 1982. Anchor species and epiphytes in intertidal algal turf. Paciﬁc Science 36, 45–59. Stewart, J. G. 1983. Fluctuations in the quantity of sediments trapped among algal thalli on intertidal rock platforms in southern California. Journal of Experimental Marine Biology and Ecology 73, 205–211. Stewart, J. G. 1989. Establishment, persistence and dominance of Corallina (Rhodophyta) in algal turf. Journal of Phycology 25, 436–446. Storlazzi, C. D. & Field, M. E. 2000. Sediment distribution and transport along a rocky, embayed coast: Monterey Peninsula and Carmel Bay, California. Marine Geology 170, 289–316. Taylor, P. R. & Littler, M. M. 1982. The roles of compensatory mortality, physical disturbance, and substrate retention in the development and organisation of a sand-inﬂuenced, rocky-intertidal community. Ecology 63, 135–146. Thom, R. M. & Widdowson, T. B. 1978. A resurvey of E. Yale Dawson’s 42 intertidal algal transects on the southern California mainland after 15 years. Bulletin of the Southern California Academy of Science 77, 1–13. Towsend, C. & Lawson, G. W. 1972. Preliminary results on factors causing zonation in Enteromorpha using a tide simulating apparatus. Journal of Experimental Marine Biology and Ecology 8, 265–276. Trowbridge, C. D. 1992. Mesoherbivory: the ascoglossan sea slug Placida dendritica may contribute to the restricted distribution of its algal host. Marine Ecology Progress Series 83, 207–220. Trowbridge, C. D. 1996. Demography and phenology of the intertidal green alga Codium setchellii: the enigma of local scarcity on sand-inﬂuenced rocky shores. Marine Biology 127, 341–351. Tsuchiya, M. & Nishihira, M. 1985. Islands of Mytilus as a habitat for small intertidal animals: effects of island size on community structure. Marine Ecology Progress Series 25, 71–81. Turner, S. J., Thrush, S. F., Cummings, V. J., Hewitt, J. E., Wilkinson, M. R., Williamson, R. B. & Lee, D. J. 1997. Changes in epifaunal assemblages in response to marina operations and boating activities. Marine Environmental Research 43, 181–199. Turner, T. 1983. Complexity of early and middle successional stages in a rocky intertidal surfgrass community. Oecologia 60, 56–65. Turner, T. 1985. Stability of rocky intertidal surfgrass beds: persistence, preemption, and recovery. Ecology 66, 83–92. Umar, M. J., McCook, L. J. & Price, I. R. 1998. Effects of sediment deposition on the seaweed Sargassum on a fringing coral reef. Coral Reef 17, 169–177. Underwood, A. J. 1998. Grazing and disturbance: an experimental analysis of patchiness in recovery from a severe storm by the intertidal alga Hormosira banksii on rocky shores in New South Wales. Journal of Experimental Marine Biology and Ecology 231, 291–306.

235

LAURA AIROLDI

United Nations Environmental Programme 1995. Global biodiversity assessment. UNEP Nairobi: Cambridge University Press. Vadas, R. L., Johnson, S. & Norton T. A. 1992. Recruitment and mortality of early post-settlement stages of benthic algae. British Phycological Journal 27, 331–351. van Katwijk, M. M., Meier, N. F., van Loon, R., Van Hove, E. M., Giesen, W. B. J. T., van der Velde, G. & den Hartog, C. 1993. Sabaki River sediment load and coral stress: correlation between sediments and condition of the Malindi-Watamu reefs in Kenya (Indian Ocean). Marine Biology 117, 675–683. Vermaat, J. E., Agawin, N. S. R., Fortes, M. D., Uri, J. S., Duarte, C. M., Marbà, N., Enríquez, S. & van Vierssen, W. 1997. The capacity of seagrasses to survive increased turbidity and siltation: the signiﬁcance of growth form and light use. Ambio 26, 499–504. Vogt, H. & Schramm, W. 1991. Conspicuous decline of Fucus in Kiel Bay (Western Baltic): what are the causes? Marine Ecology Progress Series 69, 189–194. Walker, D. I. & Kendrick, G. A. 1998. Threats to macroalgal diversity: marine habitat destruction and fragmentation, pollution and introduced species. Botanica Marina 41, 105–112. White, A. T., Vogt, H. P. & Arin, T. 2000. Philippine coral reefs under threat: the economic losses caused by reef destruction. Marine Pollution Bulletin 40, 598–605. Whorff, J. S., Whorff, L. L. & Sweet, M. H. 1995. Spatial variation in an algal turf community with respect to substratum slope and wave height. Journal of the Marine Biological Association of the United Kingdom 75, 429–444. Williams, S. L., Breda, V. A., Anderson, T. W. & Nyden, B. B. 1985. Growth and sediment disturbances of Caulerpa spp. (Chlorophyta) in a submarine canyon. Marine Ecology Progress Series 21, 275–281. Wotton, R. S. 1990. The biology of particles in aquatic systems. Boca Raton, Florida: CRC Press. Yoshida, G., Terawaki, T., Yoshikawa, K. & Arima, S. 1997. Early growth and mortality of autumnalfruiting type of Sargassum horneri (Phaeophyta) in Ohno-seto strait, Hiroshima Bay. Bulletin of Nansei National Fishery Research Institute 30, 125–135.

236

Oceanography and Marine Biology: an Annual Review 2003, 41, 237–277 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN: CURRENT STATUS AND PERSPECTIVES 1

SERGE GOFAS 1 & ARGYRO ZENETOS 2 Departamento de Biologia animal, Facultad de Ciencias – E-29071 Málaga, Spain e-mail: [email protected] 2 National Centre for Marine Research, P.O. Box 712, Mavro Lithari, GR-19013 Anavissos, Greece e-mail: [email protected]

Abstract An updated synthesis is presented for the records of introduced Mollusca in the Mediterranean basin. The rationale for taking molluscan records as signiﬁcant is discussed. The Mediterranean Sea, with some 1800 native species of Mollusca, currently houses 139 exotic species, of which 85 form established populations, 52 are aliens recorded once or twice, and two are questionable. Ten species (the gastropods Cerithium scabridum, Rhinoclavis kochi, Strombus persicus and Bursatella leachi and the bivalves Pinctada radiata and Brachidontes pharaonis in the eastern Mediterranean, the gastropod Rapana venosa and the bivalves Anadara inaequivalvis, Musculista senhousia, and Xenostrobus securis in the northern Adriatic and the western Mediterranean lagoons) are locally invasive. The bulk of the introduced species (118 species, of which 70 are established, 46 aliens, and two questionable) are species of Indo–Paciﬁc origin found mainly in the eastern basin of the Mediterranean. Among these species, some which live in the Suez Canal are most likely to have spread by their own means through this waterway (these are the “lessepsian immigrants” in the most restricted sense). For other species, the intervention of transport by ship hulls or ballast water can be suspected. Only two of these Indo–Paciﬁc immigrants are found, very locally, in the western Mediterranean. The process of immigration has become unprecedented in magnitude since the 1970s and is not slowing down. The remaining introductions of marine species are connected with mariculture and/or shipping. These vectors account for the occurrence or dissemination of only 29 exotic species in the Mediterranean basin, but four of these are invasive. The data regarding the Mollusca do not support any substantial faunal change caused by an inﬂux of subtropical faunal elements through the Straits of Gibraltar. Some local species boundaries may have changed slightly in the past decades but not a single newcomer to the Mediterranean basin by this route could be detected in this survey. The open sea localities of the western Mediterranean remain virtually free of immigrant mollusc species. The areas most severely affected by the occurrence of exotic species (eastern Mediterranean, Adriatic and lagoons in the western Mediterranean) are those where the species richness of the native fauna is low. It is speculated that this low diversity is a crucial contribution to the success of the newcomers. In the affected areas, the impact on the local fauna is considerable in terms of species composition of the assemblages but so far no native Mediterranean species can be reported as endangered as an effect of a biological invasion.

237

SERGE GOFAS & ARGYRO ZENETOS

Introduction Molluscs are an important component of marine benthic fauna worldwide, and may make up to 15–25% of the benthic macrofauna both in number of species and in number of individuals. Thus, any important change in the species composition of molluscan assemblages will have a visible impact on the benthic community in general. The exotic species, which are the subject of this review, are those which do not belong to the native fauna and have arrived recently (i.e. within historical times). An organism that has colonised an area outside its native range with human assistance, intentional or unintentional, is said to be introduced. Much attention has been given to introduced species, at ﬁrst terrestrial (Elton 1958) but later also to marine species (e.g. Carlton & Geller 1993, Carlton 1996a), because it has been realised how much of a threat they are for the balance in native faunas. Other organisms extend their geographical range following natural or man-made changes in the environment (Vermeij 1996). Some species may thrive in their new environment, displace the original components of the biota, and be regarded as “invasive”. The Mediterranean Sea holds approximately 1800 native species of molluscs (Sabelli et al. 1990). It is generally believed that the Mediterranean marine fauna vanished, or at least was greatly depleted, during the Late Miocene salinity crisis between 5.96 mya and 5.33 mya (Hsü et al. 1973, McKenzie 1999). During this time, marine species survived only in nearby areas of the Atlantic Ocean. Earlier during the Miocene, c. 12 mya, the connection between the Atlanto–Mediterranean realm and the Indo–Paciﬁc was closed (Robba 1987), and was never restored until the opening of the Suez Canal in 1869. The opening of the Red Sea in the Pliocene brought the two realms very close but the threshold was impassable for the marine fauna, and allowed considerable divergence to take place. At the time of the planning of the Suez Canal, biologists were concerned about a possible dramatic merging of the Erythrean and Mediterranean faunas. Several surveys of the Bay of Suez (Fischer 1865, Vaillant 1865, Issel 1869, MacAndrew 1870) were conducted, pioneering what is today known as baseline studies. Fischer (1865) correctly concluded that the two shores of the Suez Isthmus do not share any species of their marine mollusc faunas. The only species to have crossed the isthmus by natural means was Potamides conicus, a paralic gastropod also capable of colonising landlocked waters (Plaziat 1989). The communication between the Mediterranean and the Atlantic, nevertheless, was never interrupted since the beginning of the Pliocene. Without being aware of the salinity crisis, malacologists with a good knowledge of the Mediterranean fauna considered that it was derived from the Atlantic fauna. Pallary (1907), who gathered experience from the North African coast, wrote that “the Mediterranean having received its population from the Ocean, there is no such thing as a Mediterranean fauna” (in translation). The eastern limit, towards the Mediterranean, of those species occurring in the Ibero–Moroccan area of the Atlantic had time to reach an equilibrium position. In the past decades, human activity has exposed the Mediterranean fauna to considerable factors of change. The invasion of exotic species in the Mediterranean basin is something that has been going on for some time but lately the number of reported species has increased dramatically; currently it amounts to 139 species and is certain to increase further. Threequarters of these species have been recorded in the past three decades, in part because some conditions have changed, but also as a result of greater attention being given to micromolluscs in the eastern basin of the Mediterranean. The most spectacular change in the Mediterranean fauna is the inﬂux of Indo–Paciﬁc 238

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

species through the Suez Canal, a process for which Por (1978) coined the term “lessepsian migration” (more accurately, immigration, see Vermeij (1996)). Important post-canal surveys of the Mollusca were published by Keller (1882), Tillier & Bavay (1905), Tomlin (1927) and O’Donoghue (1929). Moazzo (1939) provided a comprehensive survey of the Mollusca in the Suez Canal, based on original collections and on the reports of previous authors. His original material is currently in Goulandris Museum of Natural History, Kiﬁssia, Greece, so that some specimens could be checked by the authors. The Hebrew University – Smithsonian Institution joint programme 1967–72, was aimed at understanding some of the characteristics of the successful migrants, and produced some collections. It has been suggested in the past decade that the distribution of tropical or subtropical Atlantic species in the Mediterranean is in progress and that this may reﬂect a trend in climatic change. This viewpoint has been substantiated essentially by ﬁsh data (e.g. Francour et al. 1994, Bombace 2001, Golani et al. 2002), occasionally by invertebrate data (Morri et al. 2001) and has been termed the “tropicalisation” or “meridionalisation” of the Mediterranean Sea. This aspect is here addressed for the Mollusca, among which there is not yet any documented case of this effect. Shipping and mariculture also brought their share of exotic species. Reports are scattered in a large number of publications, usually reporting single ﬁnds, and there have been few attempts at a synthesis of this information (e.g. Zibrowius 1992). The purpose of this paper is to summarise and discuss the data gathered for the ongoing Atlas of Exotic molluscan species prepared by the International Commission for the Scientiﬁc Exploration of the Mediterranean Sea (CIESM), based in Monaco (Zenetos et al., in press, and online Internet version at http://www.ciesm.org/atlas; hereafter referred as “the CIESM Atlas”) to which we refer for details of species occurrences and distributions. Several questions require to be addressed. (1) (2) (3) (4)

Which processes are involved in the introduction of exotic species? Is the rate of inﬂux slowing down, stable, or accelerating? Does the inﬂux meet geographical limits? What are the reasons for the success of some particular species as invaders, or for the success of invaders in some particular environments? What is the impact on the native fauna? Has any native species become extinct because of invaders?

Methods Geographic scope The CIESM Atlas covers the Mediterranean basin, including the Adriatic but not the Black Sea. In addition to the Atlas species, in the present review three cases of introductions from one area to another within the Mediterranean basin are considered. These are Gibbula albida, a north Adriatic endemic introduced into western Mediterranean lagoons (Clanzig 1989), Siphonaria pectinata, an eastern Atlantic species occurring naturally in the Alboran Sea and Algeria but introduced to Greece (Nicolay 1980), and Perna picta, also eastern Atlantic and western Mediterranean, reported from the Adriatic (De Min & Vio 1998). 239

SERGE GOFAS & ARGYRO ZENETOS

Criteria for the selection of data In the process of compiling data for the CIESM Atlas, records for nearly 200 species were reviewed, of which only 136 exotics (plus Alvania dorbignyi, a cryptogenic species reported in 1982) were accepted and given a full coverage. A list of spurious records that continue to exist in twentieth century checklists or identiﬁcation guides or published in the past three decades, is presented in Table 1. For those 62 species, a statement of the reasons for which the record is not accepted is given in the Atlas. Faulty records (see below) from earlier than 1900 were omitted. The rejected records can be classiﬁed under the following headings. Table 1 Species excluded from the CIESM Atlas of Exotic Mollusca (62 entries). Abbreviations used for origin, RS: Red Sea; IP: Indo–Paciﬁc, Carib.; Caribbean. Those entries where the Mediterranean citation is considered to proceed from a misidentiﬁcation of a native species are denoted by (m). Some species with a native eastern Atlantic/Mediterranean range are listed here because they were explicitly presented as exotic in a published report. Class GASTROPODA, subclass Prosobranchia

Clelandella infucata (Gould, 1861) Gibbula cineraria (Linné, 1758) Umbonium vestiarium (Link, 1807) Littorina littorea (Linné, 1758) Littorina obtusata (Linné, 1758) Cerithium caeruleum Sowerby G.B., 1855 Cerithium echinatum Lamarck, 1822 Cerithium erythraeoense Lamarck 1822 Bittium proteum (Jousseaume, 1930) Potamides conicus (Blainville, 1824) Scaliola elata Issel, 1869 Callostracum gracile (Maltzan, 1883) Mesalia opalina (Adams & Reeve, 1850) Rissoina chesneli (Michaud, 1832) Rissoina decussata (Montagu, 1803) Natica marochiensis Gmelin, 1791 Polynices lacteus (Guilding, 1834) Cypraea pantherina Solander in Lightfoot, 1786 Erronea caurica (Linné, 1758) Monetaria moneta (Linné, 1758) Monetaria annulus (Linné, 1758) Staphylaea nucleus (Linné, 1758) Strombus lentiginosus Linné, 1758 Bursa marginata (Gmelin, 1791) Rapana rapiformis (von Born, 1778) Coralliobia madreporarum (Sowerby, 1832) Latirus polygonus (Gmelin, 1791) Mazatlania cosentini (Philippi, 1836) Vasum turbinellus (Linné, 1758) Strigatella virgata (Reeve, 1844)

Origin

IP NE Atlantic RS, IP NE Atlantic NE Atlantic RS, IP RS, IP RS, IP RS, IP RS, S. Medit. RS, IP West Africa West Africa Caribbean Caribbean W. Africa, Carib. W. Africa, Carib. RS endemic RS, IP RS, IP RS, IP RS, IP IP W. Africa RS, IP RS, IP IP Caribbean RS, IP RS, IP

240

Date collected

1969

1980

1974

1932

1976 1958 1956

Date published

Where found

1982 1979 1970 1982 1997 1989 1986 1937 1977 1824 1977 1981 1967 1886 1886 1997 1970 1970 1968 1886 1937 1993 1934 1997 1977 1970 2000 1836 1973 1970

Sicily Malta Libya Italy Alboran Sea Israel Israel Israel Egypt S. Medit. Egypt Israel Italy France France Alboran Sea Tunisia Lampedusa Greece France Israel Israel Israel Alboran Sea Israel Italy Israel Sicily Israel Italy

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Table 1

continued

Class GASTROPODA, subclass Prosobranchia

Origin

Date collected

Date published

Where found

South Arabia NW Africa RS, IP West Africa RS, IP

1984

1985 1991 1939 1862 1937

Israel Sardinia Egypt Algeria Israel

(m) E. Atlantic, Medit. Tunisia France (m)

1996 1976 1999 1992

Notarchus indicus Schweiger, 1820 Petalifera gravieri (Vayssière, 1906)

IP (m)

1970 1970

Dolabrifera holboelli Bergh, 1872

? Greenland

1972

Berthellina citrina (Rüppell & Leuckart, 1828) Sclerodoris cf. tuberculata Eliot, 1904 Chromodoris clenchi (Russell, 1935)

(m) (m) (m)

1970 1985 1983

Cyprus Italy Israel Israel (no deﬁnite data) Israel (no deﬁnite data) (no deﬁnite data) Israel Italy Israel

(m) (uncertain) (m) NW Atlantic NW Atlantic IP IP (m)

1974 1879 1974 1915 1973 1934 1998 1969

Caribbean IP Baltic Sea

1992 1937 1992

Vexillum depexum (Deshayes in Laborde, 1834) Cymbium rubiginosum (Swainson, 1822) Lophiotoma indica (Röding, 1798) Pusionella nifat (Bruguière, 1789) Conus arenatus Hwass in Bruguière, 1792 Class GASTROPODA, subclass Opisthobranchia Cylichna cf. mongii (Audouin, 1826) Aglaja taila Marcus & Marcus, 1966 Melanochlamys seurati (Vayssière, 1926) Atys blainvilliana (Recluz, 1843) Aplysia juliana Quoy & Gaimard, 1832

Class BIVALVIA Anadara notabilis (Röding, 1798) Hochstetteria munieri Bernard, 1879 Arctinula groenlandica (Sowerby, 1842) Crassostrea virginica (Gmelin, 1791) Placopecten magellanicus (Gmelin, 1791) Spondylus spectrum Reeve, 1856 Spondylus limbatus Sowerby, 1847 Galeomma polita Deshayes, 1855 Linga aurantia Deshayes, 1832 Hippopus hippopus (Linné, 1758) Parvicardium hauniense (Høpner-Petersen & Russell, 1971) Laevicardium ﬂavum (Linné, 1758) Mactrinula tryphera Melvill, 1899 Saxidomus purpuratus (Sowerby, 1852) Petricola hemprichii Issel, 1869. Penicillus vaginiferus (Lamarck, 1818)

IP IP IP RS IP

241

1980 1956

1986 2000 1999 1905 1968

Morocco France Italy Italy Italy Israel Israel (no deﬁnite data) Israel Israel France

Israel Egypt Egypt (no deﬁnite data)

SERGE GOFAS & ARGYRO ZENETOS

Unsupported records These are statements that an exotic species exists, or may exist in the Mediterranean, or even merely a name of an exotic species within a list of Mediterranean molluscs, which are not supported by data regarding locality of collecting and actual specimens.

Records based on mislabelled or misidentiﬁed material This refers to a few cases where a genuinely Mediterranean species was misidentiﬁed as an exotic species.

Records of species ﬁnally regarded as native The only case of a native Red Sea and Mediterranean range is that of the lagoon snail Potamides conicus. Otherwise, this important question regards the Atlanto–Mediterranean species which have an Atlantic type locality and recorded range, and were subsequently found in the Mediterranean. The reverse also occurs (i.e. “Mediterranean” species that were subsequently found in the Canaries or the Ibero–Moroccan area as the fauna became better known). In such cases there is no reason to suspect the authenticity of the records but the hypothesis that those species are “invaders” which have moved through the Strait of Gibraltar must be considered with a high degree of caution. Rare species may have been in both the Mediterranean and the Atlantic for a long time, and the apparent trend is only an artifact due to the timing of records. The possible inﬂux of Atlantic species into the Mediterranean as a dynamic process has been kept under scrutiny; however, the authors retain as “introduced” only those species which have obviously been brought into the Mediterranean by human interference, an example being Mya arenaria (see Pelorce 1995).

Records based on the accidental input of isolated specimens or shells, in a context which cannot, or did not, prove to be viable This situation accounts for most rejections. Unsupported records are a particularly serious problem with Mollusca because their shells may last for a long time after the animal has died, having been transported by man for food or ornament, and left or lost in places where the animals do not live. Unfortunately, reports of molluscs based on empty shells are commonplace in the malacological literature. Petit de la Saussaye (1869) was already aware of the problem and warned, in the introduction to his Catalogue of European Shells that “these records are not more meaningful for biogeography than the ﬁnding of a crocodile in the river Seine or a monkey in the forest of Fontainebleau” (in translation). In the eighteenth and nineteenth centuries, the occurrence of exotic shells on European shores was mainly a result of the discharge of ballast sand by ocean-going ships and these shells cannot be considered as introductions. Most of such exotic species in the early descriptive literature were reported without the authors being aware that the species were not native, so that several taxa were named with an erroneous European type locality. Nowadays, the meaningless records of exotic shells derive mainly from the souvenir trade and from material brought in by ﬁshermen operating in remote areas. The occurrence 242

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

of several Indo–Paciﬁc species, including the spectacular Cypraea tigris (Linné 1758) and Cymbiola vespertilio (Linné 1758), was reported by Bini (1983) from the Caprolace lagoon on the coast of Latium. There, the occurrence of the shells was correctly interpreted as purely accidental inputs, which would not result in the introduction of these species in the Mediterranean. Other authors were not so cautious and presented such ﬁndings as “new records” without any evidence that the isolated shell(s) found had been the founders of locally-sustained populations. Some records are not tenable for biological reasons, like the alleged occurrence in the Mediterranean of the tridacnid bivalve Hippopus hippopus, based on a single record by Haas (1937). This species requires a coral reef environment that does not exist in the Mediterranean (Taviani 1994). Some old records of exotic species have now fallen into oblivion (e.g. records of Monetaria annulus, a common Indo–Paciﬁc cowry that has long been used as currency). Others are perpetuated in checklists based (often not explicitly) on the same, unconﬁrmed single original citation. Such species are sought after by some collectors who are eager to possess every species but do not have the background to decide if the occurrences are meaningful. If a species is listed as Mediterranean but does not actually live there, specimens of spurious origin will sooner or later show up, as has happened for a long time with the elusive Caribbean species Mazatlania cosentini (see Bouchet & Gofas 1983). The authors believe that there is a need to remove such species from checklists and guides, because otherwise their mention will trigger more unwarranted citations and ﬁnally give the impression that the species does exist in the region. Also rejected are records of species which, though found alive, were later proved to be unable to establish populations or to spread. An example is a population of Littorina littorea reported by Barsotti & Campani (1982) but later reported as absent (Johannesson 1988). A similar case is that of a bushel of live Patella vulgata, a north Atlantic species, discarded from a nearby ﬁsh market in the harbour of Le Brusc, Mediterranean France, in 1968. Some individuals managed to crawl up the side of the pier but died after a few days (S. Gofas, pers. obs.).

Attributes of the species The complete list of exotic species recorded in the Mediterranean basin, with the associated data, is presented in Table 2. For each species that has passed the selection criteria, information has been gathered under the following headings.

Year and place of ﬁrst collection and year of ﬁrst publication This normally post-dates the actual time of introduction by several years, maybe decades, in the less-explored areas. Nevertheless, it is the only objective datum. In the case where the original publication does not state the year of collecting, the date of publication is taken by default.

Recorded Mediterranean range This range is established from the compilation of all published data, museum material and personal communications from colleagues. In the case of established species (see below) it 243

SERGE GOFAS & ARGYRO ZENETOS

Table 2 List of exotic species recorded in the Mediterranean Sea. This list includes all species treated in the CIESM Atlas except the cryptogenic Alvania dorbignyi, and three species (Gibbula albida, Siphonaria pectinata, Perna picta) not covered in the Atlas but for which intra-Mediterranean transfers have been reported. Abbreviations used: Date, Pub: Date of publication of ﬁrst report; Rec: Actual collecting date, if mentioned. Current range, Can: Suez Canal; Lev: Levantine coast (Israel, Lebanon and Syria); T/C: Turkey and/or Cyprus; Egy: Egyptian Mediterranean coast. Ion: Ionian Sea,

Pub.

Date Rec.

Class POLYPLACOPHORA Chiton hululensis (Smith E.A. in Gardiner, 1903)

1974

1934

Class GASTROPODA, Subclass Prosobranchia Cellana rota (Gmelin, 1791) Haliotis pustulata cruenta Reeve, 1846 Diodora ruppellii (Sowerby, 1834) Smaragdia souverbiana (Montrouzier, 1863) Nerita sanguinolenta Menke, 1829 Trochus erythraeus Brocchi, 1821 Gibbula albida (Gmelin, 1791)

1967 1971 1948 1994 1973 1973 1989

1961

Pseudominolia nedyma (Melvill, 1897) Stomatella impertusa (Burrow, 1815) Cerithium nesioticum Pilsbry & Vanatta, 1906 Cerithium scabridum Philippi, 1848 Cerithium egenum Gould, 1849 Rhinoclavis kochi (Philippi, 1848) Clypeomorus bifasciatus (Sowerby G.B. II, 1855) Angiola punctostriata (Smith E.A., 1872) Planaxis griseus (Brocchi, 1821) Gibborissoa virgata (Philippi, 1849) Finella pupoides Adams A., 1860 Clathrofenella ferruginea (Adams A., 1860) Diala varia Adams A., 1860 Cerithiopsis pulvis (Issel, 1869) Cerithiopsis tenthrenois (Melvill, 1896) Metaxia bacillum (Issel, 1869) Rissoina bertholleti Issel, 1869 Rissoina spirata Sowerby, 1825 Voorwindia tiberiana (Issel, 1869) Strombus persicus Swainson, 1821 Strombus mutabilis Swainson, 1821 Sabia conica (Schumacher, 1817) Crepidula aculeata (Gmelin, 1791) Crepidula fornicata (Linné, 1758) Erosaria turdus (Lamarck, 1810) Purpuradusta gracilis notata (Gill, 1858) Palmadusta lentiginosa (Gray, 1825) Natica gualteriana Récluz, 1844 Cycloscala hyalina (Sowerby, 1844)

244

1972 2000 1977 1882 2001 1973 1985 1977 1905 1989 1977 1977 1976 1983 1989 1985 1973 1984 1980 1983 2001 1986 1992 1982 1986 1983 1990 1986 1994

Can.

Lev.

T/C

Range Egy.

*

* *

* * *

1968 1968

* *

*

*

* *

1966 1999 1970

*

*

* * * *

*

* 1971 1963 1983 1950

* *

1970 1958 1970 1935 1978 1982 1978 1965 1974

* * *

* 1978 1991 1980 1973 1957 1980 1981 1989 1966 1992

*

* * * * * * * * * * * * * * * * * * * *

* * * *

* * * *

* * * * * * * * * * * * *

* * * *

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

including Malta but not the North African shores. Gab: Gulf of Gabès. Aeg.: Aegean Sea. Geographic origin, RS: Red Sea; IO: Indian Ocean; IP: Indo–Paciﬁc. Status, E: Established; A: Alien; Q: Questionable. Mode of Introduction, Less: Lessepsian; L?: Suspected lessepsian; Ship.: Shipping; Aq.: Mariculture; ???: Unknown. Some species are scored for more than one mode. There are in total 139 exotic species, of which 85 are established, 52 alien and 2 questionable. The ten invasive species appear in boldface. Origin Ion.

Gab.

Aeg.

*

* * W Med. lagoons

*

*

* * *

Status

Others

Spain **

Less.

RS, IP

A

RS, IP RS, IP RS, IP RS, IP RS endemic RS endemic Adriatic

E A E E A E E

*

RS, IP RS, IP RS, IP RS, IO RS, IP RS, IP RS, IP RS RS endemic RS, IP RS, IP RS, IP RS, IP RS, IP IP RS, IP RS, IP RS, IP RS, IP IO, Persian Gulf RS, IP RS, IP S. Atlantic N. Atlantic RS, IP RS, IP RS, IP RS, IP RS, IP

E A A E A E A A A E E E E E E E E A A E A A E E A E A E E

*

245

Mode of introduction L? Ship. Aq.

???

*

* * * * * *

* * *

* * *

* * * * * * * * * * * * * * * * * *

*

* * * * *

SERGE GOFAS & ARGYRO ZENETOS

Table 2

continued

Pub.

Date Rec.

Sticteulima cf. lentiginosa (Adams A., 1861) Ergalatax obscura Houart, 1996 Thais lacera (Born, 1778) Thais sacellum (Gmelin, 1791) Rapana venosa (Valenciennes, 1846) Murex forskoehlii Röding, 1798 Nassarius arcularius plicatus (Röding, 1798) Fusinus verrucosus (Gmelin, 1791) Zafra savignyi (Moazzo, 1939) Zafra selasphora (Melvill & Standen, 1901) Conus fumigatus Hwass in Bruguière, 1792

1994 1995 1939 2002 1974 1905 1977 1905 1963 1993 1986

Class GASTROPODA, Subclass Heterobranchia Murchisonella columna (Hedley, 1907) Chrysallida ﬁscheri (Hornung & Mermod, 1925) Chrysallida maiae (Hornung & Mermod, 1924) Chrysallida pirintella Melvill, 1910 Adalactaeon fulvus (Adams A., 1851) Adalactaeon amoenus (Adams A., 1851) Styloptygma beatrix Melvill, 1911 Cingulina isseli (Tryon, 1886) Turbonilla edgari (Melvill, 1896) Syrnola fasciata (Jickeli, 1882) Syrnola cinctella Adams A., 1860 Odostomia lorioli (Hornung & Mermod, 1924) Oscilla jocosa Melvill, 1904 Iolaea neofelixoides (Nomura, 1936) Hinemoa cylindrica (de Folin, 1879) Leucotina cf. eva Thiele, 1925

1995 1979 1963 1989 1981 1985 1992 1983 1989 1987 1998 1987 1989 1998 2001 2001

Class GASTROPODA, Subclass Opisthobranchia Acteocina mucronata (Philippi, 1849) Cylichnina girardi (Audouin, 1826) Pyrunculus fourierii (Audouin, 1826) Bulla ampulla Linné, 1758 Haminaea callidegenita Gibson & Chia, 1989

1990 1976 1987 1982 1993

1986

2002

2001

1961 1940 1977 1965 1991 1982 1977

1959

Haminaea cyanomarginata Heller & Thompson, 1983 Chelidonura fulvipunctata Baba, 1938 Bursatella leachi de Blainville, 1817 Pleurobranchus forskalii Rüppell & Leuckart, 1828 Polycerella emertoni Verrill, 1881 Polycera hedgpethi Marcus Er., 1964 Plocamopherus ocellatus Rüppell & Leuckart, 1828 Discodoris lilacina (Gould, 1852)

246

1939 2000 1974

Can.

*

* *

*

* * * * *

1968 * 1954 1980

1974 1935 1982 1967 1978 1989 1980 1980 1958 1994 1974 1984 1994 1992 1995

*

* * * * * * * * *

T/C

Range Egy.

* * *

*

* * * * *

* * * * * * * * * * *

*

* * * * *

* * 1978 1992

1975 1964 1986 1977 1974

Lev.

*

* * * *

* * * *

* * *

* *

* *

*

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Origin Ion.

Gab.

Aeg.

*

**

Libya

*

* * * Venice lagoon

* * * *

* *

Status

Others

* ** Italy

Less. IP RS, IP IP RS, IP NW Paciﬁc RS endemic RS, IP RS endemic RS RS, IP RS, IP

A E E E E E A E E E A

RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP IP IP IP

A E E E E E E E E E A A A A A A

RS, IP RS, IP RS, IP RS, IP N Atlantic

E E E E E

RS, IP

A

IP RS, IP RS, IP N Atlantic S Atlantic RS, IP RS, IP

E E A E A E E

247

Mode of introduction L? Ship. Aq.

??? * *

* * * * * * * * *

* * * * * * * * * * * * * * * *

* * * * * * * * * * * * *

SERGE GOFAS & ARGYRO ZENETOS

Table 2

continued

Hypselodoris infucata (Rüppell & Leuckart, 1830) Chromodoris quadricolor (Rüppell & Leuckart, 1830) Dendrodoris fumata (Rüppell & Leuckart, 1830) Melibe ﬁmbriata Alder & Hancock, 1864 Cuthona perca (Marcus Er., 1958) Flabellina rubrolineata (O’Donoghue, 1929) Caloria indica (Bergh, 1896) Aeolidiella indica (Bergh, 1888) Class GASTROPODA, Subclass Pulmonata Siphonaria crenata Blainville, 1827 Siphonaria pectinata (Linné, 1758)

Class BIVALVIA Acar plicata (Dillwyn, 1817) Anadara demiri (Piani, 1981) Anadara inaequivalvis (Bruguiere, 1789) Anadara natalensis (Krauss, 1848) Glycymeris arabica (Adams H., 1871) Limopsis multistriata (Forsskål, 1775) Perna picta (Born, 1780) Musculista perfragilis (Dunker, 1857) Musculista senhousia (Benson in Cantor, 1842) Modiolus auriculatus (Krauss, 1848) Xenostrobus securis (Lamarck, 1819) Brachidontes pharaonis (Fischer P., 1870) Septifer forskali Dunker, 1855 Crassostrea gigas (Thunberg, 1793) Saccostrea commercialis (Iredale & Roughley, 1933) Saccostrea cucullata (Born, 1778) Dendrostrea frons (Linné, 1758) Pinctada margaritifera (Linné, 1758) Pinctada radiata (Leach, 1814) Malvufundus regulus (Forsskål, 1775) Chlamys lischkei (Dunker, 1850) Spondylus spinosus Schreibers, 1793 Spondylus groschi Lamprell & Kilburn, 1995 Spondylus cf. multisetosus Reeve, 1856 Divalinga arabica Dekker & Goud, 1994 Trapezium oblongum (Linné, 1758) Diplodonta cf. subrotunda Issel, 1869 Chama paciﬁca Broderip, 1834 Pseudochama corbieri (Jonas, 1846)

248

Pub.

Date Rec.

1977 1986

Can.

Lev.

T/C

1965 1982

*

*

1986 1984 1995 1993 1993 1968

1980 1982 1977 1988 1986

*

1972 1980

1965

1986 1977 1973 1937 1977 1977

* *

*

*

*

*

1978 1972 1969 1935 1966 1965

*

*

*

* * *

1971 1971 1937 1994 1877 2001 1964 1985

1960 1964

*

2001 2001 1974 1878 1931 1985 1993 1998 2001 1979 1980 1999 1905 1946

1999 1998

*

1874 1931 1985 1988

* *

* *

* *

*

* *

*

*

* * * * *

*

*

*

* * *

Range Egy.

*

* *

1992 1999

* *

*

* *

*

* *

*

1984

1992 1956 1980

* * *

*

* * *

*

*

*

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Origin Ion.

Gab.

Aeg.

Status

Others

Less.

RS, IP Ligurian Sea RS, IP

*

*

* Adriatic

*

*

Mode of introduction L? Ship. Aq.

E A

*

RS, IP IP S. Atlantic RS, IP RS, IP RS, IP

A E A A A A

*

RS, IP Atlantic, W.Med.

E E

???

*

* * * * *

* *

E

*

Adriatic **

Adriatic

** ** *

* *

*

*

**

* * **

*

RS, IP SE Asia NW Paciﬁc RS, IP RS, IP RS, IP Atlantic, W.Med. RS, IP SE Asia RS, IP SE Asia RS, IP RS NW Paciﬁc S Australia

A E E E A A A

RS, IP RS, IP RS, IP RS, IP RS, IP S. Atlantic RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP

A E E E E A E Q Q E A A E A

249

E E E E E A E A

* * *

*

* * * * * *

*

* * *

* * * * * * * *

* *

* * * * * * * * *

SERGE GOFAS & ARGYRO ZENETOS

Table 2

continued

Pub. Fulvia australis (Sowerby G.B., 1834) Fulvia fragilis (Forsskål, 1775) Afrocardium richardi (Audouin, 1826) Tellina valtonis Hanley, 1844 Psammotreta praerupta (Salisbury, 1934) Hiatula rueppelliana (Reeve, 1857) Mactra olorina Philippi, 1846 Mactra lilacea Lamarck, 1818 Atactodea glabrata (Gmelin, 1791) Gafrarium pectinatum (Linné, 1758) Circenita callipyga (Born, 1778) Clementia papyracea (Gray, 1825) Paphia textile (Gmelin, 1791) Mercenaria mercenaria (Linné, 1758) Ruditapes philippinarum (Adams & Reeve, 1850) Antigona lamellaris Schumacher, 1817 Dosinia erythraea Römer, 1860 Timoclea roemeriana (Issel, 1869) Petricola pholadiformis Lamarck, 1818 Sphenia rueppelli Adams A., 1850 Mya arenaria Linné, 1758 Gastrochaena cymbium (Spengler, 1783) Laternula anatina (Linné, 1758) Totals (questionable not included)

1948 1973 1999 1977 1999 1905 1889 2001 1977 1905 1986 1948 1939 1965 1981 1999 1905 1999 1994 1986 1976 1973 1905

Date Rec.

1955 1999 1970 1992

1972 1973 1972

1981 1992

Can.

* * * * * * * * * * *

Lev.

T/C

* * *

* * * * *

* * * * * * *

Range Egy.

* * *

*

*

* *

*

1997

* * *

* *

1978

*

*

1960

* *

* *

* *

*

53

92

66

34

*

generally represents the current range at the time of publication, and is likely to expand. Conversely, for the species recorded as “alien”, records more than a few years old may represent occasional ﬁndings or failed introductions, and records do not imply that the species currently lives there.

Relation to current Mediterranean range The geographical ranges of living organisms are subject to ﬂuctuate both on a historical timescale and on a broader geological timescale. The “current” range is actually the summary of records over many years in historical times (some records dating back from the late eighteenth century) and ignores small short-term variations. The concept of an exotic species refers to new occurrences which clearly depart from this “current” range. The following categories can be considered. Exotic (or non-indigenous) All species, in the broader sense, that are not native. This category encompasses species that have arrived within historical times with the help of human intervention or by their own means. Broadly, this grouping includes lessepsian immigration, 250

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Origin Ion.

Gab.

Aeg.

*

*

** **

* * *

10

7

Status

Others

**

Less. RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP RS, IP N. Atlantic NW Paciﬁc RS, IP RS, IP RS, IP N. Atlantic RS, IP N. Atlantic RS, IP RS, IP

21

E E E E A E E E A E A E E E E A A E A A E E E

Mode of introduction L? Ship. Aq.

???

* * * *

*

* * * * * * * * * * * * * * * * * * * 47

43

17

12

25

the possible immigration of Atlantic species into the Mediterranean through the Straits of Gibraltar, and introduced species. Introduced A term used here only for species which have been transported by man or his artifacts. These are clearly identiﬁed as not belonging to the native fauna, and have a known range from which a source area may be hypothesised. In most cases, there are reliable surveys indicating their absence at some particular time prior to introduction. Cryptogenic Carlton (1996b) coined this term for species which cannot be readily determined as members of the native fauna, or as exotic ones. This usually includes species which, if introduced, have been brought in a long time ago (typically, before AD 1800) so that the history of introduction has not been witnessed. Reasons for suspecting introductions include historical texts, discrepancies with usual biogeographical patterns, and observations that the Mediterranean populations form a disjunct part of the total range. The cryptogenic species (except one, Alvania dorbignyi – see below) have been excluded from the CIESM Atlas for the reason of being long-established, but these will be discussed herein. 251

SERGE GOFAS & ARGYRO ZENETOS

Native (or indigenous) Species, not considered in this survey, with a “current” Mediterranean range stable at least since historical times. Evidence for this status comes from early surveys (for the Mediterranean, prior to 1869) and museum material, and from fossil occurrences in Pleistocene or Holocene deposits.

Native range, as given by the literature The worldwide native range, which includes the source area, was taken from literature data, referring as much as possible to ﬁrst-hand data in revisions or reports with an actual material examined, not to general statements in identiﬁcation guides.

Introduced range outside the Mediterranean Reports of introductions in other parts of the world were also taken into account, since there are many cases of species prone to introduction worldwide.

Establishment status This entry considers the success of invasion and perenniality of introduced populations. Because the recorded species are not usually monitored continuously through time, the following standards have been used. Established when documented by at least three reliable records, or when the available information shows that there are perennial populations in the Mediterranean. Several records or observations at the same site but at least one year apart are counted as separate records (e.g. Crepidula aculeata). Alien based on one or two records. Among these, most records more than a few years old do not imply current occurrence of the species at the reported sites (otherwise, they would probably qualify as “established”). Questionable This has been used for a few taxa for which there are unsettled taxonomic issues, and which may in the future be placed in synonymy. The species with this status will not be considered further in the species counts and evaluation of trends.

Mode of arrival The mode of arrival is seldom known with certainty. Usually, it is surmised from the distribution data and the circumstances of the earliest record. The following categories were considered in this review. Lessepsian The term was proposed by Por (1978) for species which have penetrated from the Red Sea into the Mediterranean, via the Suez Canal using their natural means of dispersal. This pathway is veriﬁed for species that have become established in the Suez Canal prior to their settlement in the Mediterranean, and have penetrated progressively along the 252

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

shores of the Suez Canal and the eastern Mediterranean. Species that are recorded in the Red Sea, in the Suez Canal, and in Egypt or the Levantine coast are scored as lessepsian. The species that occur in the Red Sea and ﬁrst showed up in Egypt or along the Levantine coast but are not recorded from the Canal were scored as probable lessepsian. Conversely, species of Indo–Paciﬁc origin that ﬁrst showed up in more distant Mediterranean localities (e.g. Turkey, Ionian Sea) were scored with unknown mode of introduction. Shipping Species that were ﬁrst noticed as isolated populations in a major harbour are scored as introduced by shipping. Mariculture Species that were voluntarily introduced to be cultivated or species that were ﬁrst observed in areas dedicated to mariculture were scored as introduced with mariculture. The species maintained exclusively in artiﬁcial conditions, or in aquaria, are not considered. For species suspected to have arrived in the Mediterranean after previous introductions to intermediate areas, and not directly from their native range (e.g. Mya arenaria, Rapana venosa), it is the ﬁnal step that is taken into account. In a few cases more than one mode is considered for the same species, regarding different subranges within the Mediterranean.

Type of larval development The larval stage is known to be crucial for dispersal in the Mollusca. Species that have a long planktonic larval stage are likely to be transported to remote sites by currents but also in ballast water of ocean-going ships. In shelled gastropods, the type of larval development can be inferred from shell morphology (Jablonski & Lutz 1980, Levin & Bridges 1995). The type of larval development has been scored for all the prosobranch species involved, because in this group the interpretation of protoconch features is quite straightforward. All vetigastropods have been scored as “acteplanic”, because they release short-lived larval stages that stay a few days in the plankton but do not feed prior to metamorphosis.

Sources of the exotic species Cryptogenic species The authors view as cryptogenic ﬁve molluscan species of the Mediterranean, which are discussed hereafter but are not considered in the species counts. Only one of them was treated in the CIESM Atlas, complying with the editorial policy of not considering very old records. Littorina saxatilis (Olivi, 1792) was originally described from Venice, which is part of a limited subrange of the species in the northern Adriatic. It also occurs in small subranges in the Gulf of Gabès, Tunisia, in the Straits of Gibraltar, and currently has its main range in the North Atlantic from Canada, Greenland and northern Norway south to Portugal. Two conﬂicting hypotheses can be considered for the Mediterranean occurrences (Reid 1996: 331): that it is a relict from a former range that included the Mediterranean at the time of 253

SERGE GOFAS & ARGYRO ZENETOS

Pleistocene cold spells, or an old introduction through navigation. The relict hypothesis is tenable because the localities involved are those in the Mediterranean with a substantial tidal range, and at least one reliable fossil occurrence is known in Pleistocene deposits off Marseille, France (one shell in Muséum National d’Histoire Naturelle, Paris, unpubl. datum). A genetic analysis attempted to investigate the latter hypothesis but remained inconclusive (Janson 1985). A tenable alternative is a natural airborne introduction by wading birds (see Rees 1965), which could also explain the occurrences in Tunisia and southern Morocco. There are some uncertainties regarding the small rissoid Alvania dorbignyi Audouin, 1826 (see Mienis 1985). In favour of accepting it as a native Mediterranean species, it belongs to a species group including several well-known Mediterranean natives, for example, A. discors (Allan 1818), A. lanciae (Calcara 1841) and A. consociella di Monterosato, 1884. All these species live among brown algae in clean, shallow open sea locations. Conversely, all undisputed native Indo-Paciﬁc Rissoidae belong to different genera or, at least, different species groups. The origin of Audouin’s specimen is uncertain, because that material included Mediterranean as well as Red Sea species, and the subsequent records of A. dorbignyi in the Gulf of Suez (Moazzo 1939) are well after the opening of the Suez Canal. The point is that there are no documented introductions of Mediterranean molluscs into the Red Sea. There is no reason to believe that this species, rather rare on the Levantine coast and not particularly hardy, would be one of the few exceptions. A further support for holding this species as Indo-Paciﬁc is that a very similar, if not conspeciﬁc, form has been found living on the coast of New Caledonia (S. Gofas, pers. obs. 1993). Barash & Danin (1973, 1992) suggested Aspella anceps (Lamarck 1822) as an IndoPaciﬁc species immigrating into the Mediterranean. The same species is also found in Pleistocene deposits bordering the Red Sea at Hurghada, Egypt (Houart & Vokes 1995), indicating that it is deﬁnitely native in the Indo–Paciﬁc realm. However, the Mediterranean populations were recorded very early by di Monterosato (1880) who described it as a new species Epidromus gladiolus from specimens collected in sponges of undetermined origin, and from two specimens dredged in the harbour of Alexandria, Egypt (Gaglini 1987). Pallary (1912) recorded as two separate entries di Monterosato’s species and a “Ranella sp.” which, judging from the description given and source, is the same. Modern records (Houart & Vokes 1995) do not extend beyond the eastern basin of the Mediterranean. It is very unlikely that this rare, sublittoral species, not recorded from the Suez Canal, could be one of the pioneer migrant species, whereas all the other early migrants (see below) are euryhaline species well installed in the canal. One possibility is that it has been brought into the Mediterranean in historical times (independently from the opening of the Suez Canal) with the trade of commercial sponges from the Red Sea, or that all Mediterranean records are based on discards from the sponge trade. The alternative interpretation, also unproved, is that the Mediterranean population is native, and that there is a morphologically indistinguishable sibling species in the Indo–Paciﬁc. Mytilus galloprovincialis Lamarck, 1819 has its type locality in the French Mediterranean, but is prone to introductions worldwide and is recorded as one of the “top 100 invasive species”. There are reasons to suspect that some Mediterranean populations are introduced, particularly where the species is cultured. For the populations found in the western Mediterranean, particularly in the Alboran Sea, there is a continuity of the range with the Atlantic Ibero–Moroccan area, and there is little doubt that in this area the species is native. Teredo navalis (Linné 1758) ﬁrst appeared around 1730 in the North Sea where it severely affected the Dutch maritime constructions (Selli 1733, Reise et al. 1999, Hoppe 254

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

2002). It was probably brought from SE Asia by wooden ships (indiamen). The extent and origin of its occurrence in the Mediterranean is not easy to determine, and there are few positive records (e.g. the Venice lagoon, see Olivi 1792: 197).

Immigrants from the Indo–Paciﬁc The bulk of the exotic species recorded from the Mediterranean is of Indo–Paciﬁc origin (marked IP, RS or Persian Gulf on Table 2). There are 47 species classiﬁed as lessepsian, 43 more as probable lessepsians, which makes a total of 65% of the exotic species. To this list may be added the 17 tropical Indo–Paciﬁc species which are not found near the Mediterranean entrance of the Suez Canal but occur on the coasts of Turkey, one species (Pinctada margaritifera) introduced deliberately, two species (Thais lacera and T. sacellum) introduced by shipping, scattered occurrences of unknown pathway (Rissoina spirata, Conus fumigatus, Melibe ﬁmbriata, Haminaea cyanomarginata, Chromodoris quadricolor), and the two questionable species of Spondylus, bringing the share of tropical Indo–Paciﬁc exotics to a total of 118 (85%). The data in Table 3, comparing current results with Moazzo’s (1939) survey, demonstrate that having established populations in the Suez Canal is an efﬁcient launching pad for subsequent settling in the Mediterranean. Of the 57 species recorded by Moazzo as established in the Suez Canal (scored as “common” or “quite common” at least in one section of the Canal), about half are now present in the Mediterranean. The other half never made the move. Conversely, of the 152 species recorded from the Gulf of Suez but not in the Suez Canal, only 15 (10%) are now found in the Mediterranean. The percentage is only slightly higher (23%) if one considers species that do occur in the Suez Canal but are scarce. The inﬂux of Indo–Paciﬁc species into the Mediterranean is a dynamic process. The early settlers have expanded their ﬁrst ranges, and both the number of species and the affected areas have been increasing since 1869. There are only four Indo–Paciﬁc molluscan species which were proved to be established in the Mediterranean prior to 1900. These are Cerithium scabridum, Brachidontes pharaonis, Pinctada radiata, and Mactra olorina. All four were recorded in the Suez Canal by Tillier & Bavay (1905) and later authors, and three of them are now at the forefront of the penetration of lessepsians in the eastern basin of the Mediterranean. Cerithium scabridum was mentioned from Port Said, Egypt by Keller (1882). By the end of the nineteenth century, it had reached the Levantine coast. There are specimens from Table 3 Status of molluscan species in the Suez Canal according to Moazzo (1939) and previous records cited therein, compared with current status in the Mediterranean Sea according to CIESM Atlas data. Suez Canal data (Moazzo 1939)

Current status in the Mediterranean Established Alien Not recorded or questionable

Established in the Canal Others present in Canal Recorded in Gulf of Suez but not in Canal

22 10 11

57 79 152

7 8 4

255

28 61 137

SERGE GOFAS & ARGYRO ZENETOS

Jaffa (now Yafo, near Tel Aviv) in the Vignal collection, Muséum National d’Histoire Naturelle, Paris, collected in 1899. Pallary (1938) later mentions it as well established on the coasts of Syria and Lebanon. The species was found on the east coast of Sicily in the late 1970s (Di Natale 1978), in the harbour of Porto Megarese. An extensive survey of molluscan faunas around Djerba, Tunisia carried out in 1982 by P. Bouchet and colleagues (unpublished material in Muséum National d’Histoire Naturelle, Paris) did not detect C. scabridum, but our own collecting in October 1999 yielded large numbers at several sites around this island. This means that the Sicilian population was introduced at one or more point sources, possibly by ships calling at Porto Megarese, and not by progressive spreading along the North African coast and across the Sicily straits. Pinctada radiata was ﬁrst recorded by di Monterosato (1878, 1884) from Alexandria (collected in 1874) and the Mediterranean populations were later named as a new species Meleagrina savignyi di Monterosato 1884. It was recorded in the Gulf of Gabès, Tunisia, by 1890 (Seurat 1929) but had not been found there in a previous survey in 1882. It was also established in Cyprus in the late nineteenth century (di Monterosato 1899). Brachidontes pharaonis was recorded very early at Port Said but only later along the Mediterranean coast. It had reached the Levantine coast by 1931, and now extends to Cyprus, southern Turkey and a few Aegean localities. It was reported from Sicily in 1969 (Di Geronimo 1971) and is now well established there. As with Cerithium scabridum, the most likely cause for this latter occurrence is transportation by ships from the Middle East. The species was not found in the Gulf of Gabès in any of the previously mentioned samplings. The fourth early settler, Mactra olorina, is extremely common in the Suez Canal but, possibly due to the lack of suitable habitats, the established Mediterranean populations remain mainly conﬁned to the Egyptian coast. The progress of the Indo–Paciﬁc immigrants in the ﬁrst half of the twentieth century is difﬁcult to assess, because there were no data for the coastline of Turkey and Cyprus. Tillier & Bavay (1905) conﬁrmed the occurrence of the above four species, and reported seven additional immigrants in the harbour of Port Said and one more at Alexandria. It is noteworthy that the very thorough survey of Pallary (1912) for the Mediterranean coast of Egypt did not reveal any novel immigrant. This ﬁnding means that the ﬁgures published at that time are realistic, and that no more than 12 Indo–Paciﬁc species had a foothold in the Mediterranean in the early twentieth century (see Fig. 1). At the same time, the molluscan fauna of the Gulf of Gabès was quite well known (Pallary 1904–6) and by then Pinctada radiata was the only Indo–Paciﬁc species to occur there. The reports of Haas (1937, 1948) and Pallary (1938) brought new data for the Levantine coast, and at the same time Moazzo (1939) provided an update for the Egyptian coast. Moazzo conﬁrmed all previous records except Planaxis griseus, and added Paphia undulata and Thais lacera. Malvufundus regulus, now well established on the Levantine coast, was ﬁrst recorded there by Gruvel & Moazzo (1931). Haas reported from Palestine Anadara natalensis, Modiolus auriculatus, and O’Donoghue & White (1940) added the opisthobranch Bursatella leachi. A few more species (Chiton hululensis, Diala varia and Chrysallida maiae) were mentioned in later literature from specimens collected in the 1930s. Several records by Aharoni (1934), Haas (1937) and Steuer (1939) have been rejected in the present review because they are based on shells only; their records are for species which were never found again and require environments that are not found in the Mediterranean. These are Cerithium erythraeoense, Strombus lentiginosus, Lophiotoma indica, Conus arenatus and Hippopus hippopus. The detailed survey of the Levantine fauna by Pallary (1938) 256

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Figure 1 Numbers, per decade, of newly recorded immigrant molluscan species in the Mediterranean basin, and of publications regarding the subject (based on the reference list of the CIESM Atlas, Zenetos et al. in press). For each decade, the right column shows the years when the species was ﬁrst reported in a publication, the left column the year of actual collecting (by default, same as year of publication). The questionable and cryptogenic species are not counted. Dark grey: species of tropical Indo–Paciﬁc origin, light grey: others.

did not involve any previously unrecorded species. The number of Indo–Paciﬁc molluscs in the eastern Mediterranean at that time, when pruned of unwarranted occurrences, amounted to 21 species. Of these, only four (Cerithium scabridum, Pinctada radiata, Brachidontes pharaonis, Malleus regulus) had signiﬁcantly spread out and became common, whereas 10 (Fusinus verrucosus, Thais lacera, Planaxis griseus, Chama paciﬁca, Hiatula rueppelliana, Mactra olorina, Dosinia erythraea, Gafrarium pectinatum, Paphia textile, Laternula anatina) were still conﬁned to the Egyptian coast. Data are totally wanting for the coastlines of Turkey and Cyprus in the same period but a realistic number of exotics in that region could have comprised between four (the four species whose range is expanding) and 12 (the total number of species then mentioned for the Levantine coast). It is after 1950 that many species were reported from the Levantine coast. The number of new reports increased dramatically in the 1970s (Fig. 1), but if dates of collecting are considered rather than dates of ﬁrst publication, the histogram is less skewed to the right. Two major events have been invoked to explain the acceleration in the rate of immigration: the construction of the Aswan high dam in 1964 (Oren 1969, El Sayed & van Dijken 1995), which dramatically reduced the freshwater outﬂow from the Nile, and the closure of the Suez Canal from 1967 to 1975, which allowed marine life to develop in the canal without disturbance. The data indicate that some important immigrants (Finella pupoides, Rhinoclavis kochi, Zafra savignyi, Syrnola fasciata, Chrysallida maiae, Fulvia fragilis) were already in the Mediterranean prior to those events. Nevertheless, it cannot be disputed that the bulk of the immigrants post-date them. 257

SERGE GOFAS & ARGYRO ZENETOS

The species which now have the widest distribution are those that have been in the Mediterranean for a long time. The seven species present in Gulf of Gabès and/or Sicily had, by 2001, a mean of 74 yr in the Mediterranean, compared with the 39 yr mean residence of the 90 species present on the Levantine and/or Egyptian coast but not in the outposts of Gabès or Sicily. Figures are not given for Turkey as these would be of little value because of the lack of data before 1980. Figure 2 shows the progress in numbers of Indo–Paciﬁc immigrants in the last quarter of the twentieth century. The data for Figure 2A are essentially the same as those considered by Por (1978: 138). The counts are slightly at odds with Por’s because in the present review the authors (a) counted as Mediterranean the species established at Port Said, whereas Por classiﬁed them as Suez Canal species, (b) rejected several of the records considered in Por (1978), here listed in Table 1, and (c) added species which had been collected before 1975 but published later. The very low number of immigrants (the bivalves Pinctada radiata, Malvufundus regulus and the opisthobranch Chelidonura fulvipunctata) registered for Turkey and Cyprus by 1975 may be an artifact, since this coastline was virtually unexplored until the 1980s. Otherwise, the ﬁgures show that the number of immigrants has approximately doubled in 25 yr and that the incidence of immigrants has extended to other areas of the eastern Mediterranean. The Gulf of Gabès, which for a long time housed only Pinctada radiata, today features also Cerithium scabridum, Bursatella leachi, Melibe ﬁmbriata, Acteocina mucronata and Fulvia fragilis. Among the Indo–Paciﬁc immigrants reported from the Mediterranean in the 1980s and 1990s, there is a pool of species recorded from the southern coast of Turkey and from Cyprus, but not from the Levantine coast nor from Egypt. The most noteworthy is Strombus persicus, ﬁrst sighted in the Mediterranean in 1978 (Nicolay & Romagna-Manoja 1983) from SW Turkey, and today invasive in several areas. It rapidly expanded its range to include Israel (Mienis 1984), Rhodes (Verhecken 1984), Cyprus (Bazzocchi 1985) and Lebanon (Bogi & Khairallah 1987). It is now invasive in most of these places and was sold on the Yafo ﬁsh market (Mienis 1999). This species, which is native to the Persian Gulf and southern Arabia and does not occur in the Red Sea, has been presented as an example of “non lessepsian” immigration (Oliverio 1995). It was postulated that S. persicus may have arrived in Iskenderun in discharged ballast water from oil tankers coming from the Persian Gulf. However, the gastropod has been found far from oil terminals, and because tankers at the time lacked segregated ballast tanks, oil-laden vessels would not carry ballast water. Other species which present a distributional gap between the Suez Canal and their Mediterranean subrange are the prosobranchs Smaragdia souverbiana, Stomatella impertusa, Cycloscala hyalina, Sticteulima cf. lentiginosa, Ergalatax obscura, ﬁve species of heterobranchs, and the bivalves Saccostrea cucullata, Dendrostrea frons, Septifer forskali, Psammotreta praerupta and Antigona lamellaris. These species have not been seen on the Levantine coast despite appropriate attention in the same period, or have been recorded there later than in Turkey. Thus, their pattern of spreading is unlikely to be a progression along the shore starting from the Suez Canal. Two conﬂicting hypotheses may be considered: (1)

they have been introduced by other means, probably shipping (the tanker terminal at Iksenderun has been proposed as a possible source, but Mersin is also a major port in this area);

258

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Figure 2 (A) Situation by year 1975 of the exotic species of tropical Indo–Paciﬁc origin. (B) Situation by year 2001 of the exotic species of tropical Indo–Paciﬁc origin. (C) Same, considering only the established species. For each area, the numbers in parenthesis indicate species occurring also in the Red Sea (RS) and within the Suez Canal (SC). The questionable, cryptogenic and cultivated species are not included in the counts.

259

SERGE GOFAS & ARGYRO ZENETOS

(2)

these species came through the Suez Canal as larvae, which could ﬁnd appropriate environments to settle in southern Turkey but not in Egypt nor on the Levantine coast. As a support for the latter hypothesis, ﬁve of the gastropods cited above have multispiral larval shells indicating planktotrophic development, and the other species have planktonic, albeit short-lived, larvae.

Other records may remain unexplained. The isolated live specimen of the Indo–Paciﬁc nudibranch Chromodoris quadricolor collected in 1982 in the Ligurian sea (Cattaneo-Vietti 1986) may have escaped from a tropical aquarium – a pathway that contributed to the Mediterranean ﬂora, the well-known invasive alga Caulerpa taxifolia.

Introductions for mariculture The introduction of molluscs as a consequence of mariculture is considered when it is the cultured species which has become established or when the species ﬁrst became established on sites used for mariculture. It does not appear that any pre-1950 introductions have resulted in establishing viable populations in the Mediterranean. An Italian company imported live Meleagrina margaritifera (the pearl oyster) from the Dahalak Islands, Red Sea around 1860 and kept them in an aquarium, where they grew, reproduced and even produced pearls (Bellet 1899). There were plans to introduce the species on the Calabrian coast between Bova and Torre di Riacci, but there are no ascertained records. The importation process gained in importance in the 1960s, and reports often appear in unpublished reports (“grey literature”) which makes them difﬁcult to trace. One such record refers to the occurrence of the pearl oyster in Abou-Kir Bay, Alexandria, Egypt (Hasan 1974). The oyster Crassostrea gigas was one of the earliest introductions but it ﬁrst became established on the Atlantic coast of Europe (see Lambert et al. 1929, for details) and was brought much later to the Mediterranean (Raimbault 1964). Lamarck (1819) described Gryphaea angulata as originating from the Tagus estuary, Portugal, and did not relate it to the Japanese species. Populations introduced from Portugal to the Atlantic coast of France in 1869 were long known as “Portuguese” oysters and believed to be a native European species. In the 1970s, when oyster farms were depleted by a disease and Japanese spat of Crassostrea gigas was imported to Europe, it was realised that both stocks freely interbred and could be considered conspeciﬁc (Menzel 1974). Two pieces of evidence support the introduction hypothesis. First, the early nineteenth century range of the “Portuguese” oyster does not match any pattern observed in native species; molluscs occurring in Portugal are found as a rule elsewhere on the European and/or Moroccan Atlantic coast. Second, such similar gene pools could not have been maintained since the Miocene (supposing a Tethyan relict) without divergence. The history of Lisbon as a cosmopolitan port makes introduction a very likely explanation. Recent evidence from sequences of the mitochondrial genome suggests Taiwan as the likely source for the Portuguese populations (Ó Foighil et al. 1998), and also indicates that sufﬁcient divergence has taken place for both lineages to be traced (Fabioux et al. 2002). Japan is the main source for the recent imports of spat. Transfers in the context of mariculture may also result in the introduction of native Mediterranean species to parts of the basin where they did not occur naturally. This is veri260

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

ﬁed for the endemic Adriatic trochid Gibbula albida, which has been successfully introduced to the French Mediterranean lagoons (Clanzig 1989) and to the Ebro delta (S. Gofas, pers. obs.). Large amounts of the naticid Neverita josephina (Risso 1826), which does not occur in the Alboran Sea, are brought to southern Spain with Adriatic catches of the commercial venerid Chamelea gallina (Linné 1758). They are occasionally discarded in harbours, and shells are found, but to date this has not resulted in an introduction of the former species. An important addition was Ruditapes philippinarum. In 1983, 200 000 juvenile individuals originating from an English hatchery and, in 1984, another 500 000 individuals of Atlantic origin were imported in the Venice lagoon for marine farming. The species was subsequently imported in various Italian localities where it was acclimatised, successfully farmed and spread in the wild, locally displacing the native R. decussatus. The species has become of high commercial importance. It is now farmed in lagoonal areas where it reaches a density of 1000 ind. m2 (Cesari & Pellizzato 1985, Breber 2002). In French lagoons the species was also imported in the 1980s, where it became well established, acclimatised and spread rapidly forming natural populations. A number of accompanying exotic species have been attributed to the import of R. philippinarum. Among oysters, the farming of Crassostrea gigas has been successful in the French Mediterranean lagoons and sustains a prosperous commercial activity (Gangnery et al. 2001). Conversely, the introduction of Saccostrea commercialis in the Venice area was not, and the species has not been seen in recent years (Mizzan 1999). Accidental introductions may accompany the voluntary transfers. A typical example of an accidental invader with mariculture is that of Musculista senhousia, a fouling organism which is an aggressive invader, with adaptations to a variety of habitats. The species originates from the West Paciﬁc whence it was introduced to northwest America in 1941 (Carlton 1992) and to Australia in 1983 (Willan 1987). In the Mediterranean it is now common in lagoons of the Adriatic and France. On the Paciﬁc coast of America the dramatic increase of the mussel’s population is presumably as a result of transfer in the ballast waters of ships. French populations of this species were probably imported with oysters from Japan, around 1978 (Hoenselaar & Hoenselaar 1989). Similarly, the Adriatic populations were possibly introduced with the clam Ruditapes philippinarum, which was massively imported for mariculture in 1986 (Lazzari & Rinaldi 1994). Very similar in behaviour, and invasive in character, are the bivalves Xenostrobus securis and Anadara inaequivalvis whose massive presence in the Adriatic lagoons is related to intense shellﬁsh farming in the area (Occhipinti-Ambrogi 2000).

Introduction via shipping The introduction via shipping routes results from transport on ships’ hulls or in ballast water. An occurrence ﬁrst reported in a harbour is, of course, a clue to this pathway of introduction. Some of the Indo–Paciﬁc species in the eastern Mediterranean have been discussed above and may enter this category. Other records are haphazard and sometimes cannot be clearly demarcated from those resulting from mariculture activities. The occurrence of Crepidula aculeata in the harbour of Alicante, Spain, registered in 1980 and still observed in 1997, is a representative example. In several cases, transport via shipping routes may have occurred in combination with 261

SERGE GOFAS & ARGYRO ZENETOS

other pathways of arrival. This situation can apply to exotic species already established in one part of the Mediterranean as, for example, the introduction to Sicily of the lessepsian migrants Cerithium scabridum and Brachidontes pharaonis. Preliminary results of molecular studies have shown that there are some Red Sea genotypes of the latter species in the Mediterranean Sea, but the non-Red Sea genotypes are fairly common and their frequency increases with distance from the Suez Canal (A. Abelson, pers. comm.). This suggests that ship transport from elsewhere may have occurred for this species rather that natural migration. Like mariculture, shipping can also affect the range of species originally native in a separate part of the Mediterranean, such as the pulmonate Siphonaria pectinata, originally restricted to the Alboran Sea and western Algeria (Morrison 1972) and now thriving in the Saronikos Gulf, Greece (Nicolay 1980, and S. Gofas, pers. obs.). Figure 3 summarises occurrences of molluscs of other than tropical Indo–Paciﬁc origin in the Mediterranean, in total 21 species (15 established and 6 alien). Among these, 10 originate from temperate areas of the Paciﬁc Ocean, eight from the American Atlantic, and three were native in a distinct part of the Mediterranean. It is noteworthy that among the American species, four (Crepidula fornicata, Mercenaria mercenaria, Petricola pholadiformis, Mya arenaria) were well established as introduced species on the Atlantic European coast before proceeding to the Mediterranean.

Are any species entering from the Atlantic? The current data regarding the Mollusca do not support any hypothesis of dynamic faunal change caused by an inﬂux of subtropical faunal elements through the Straits of Gibraltar. In Table 4, a selection of shelled Mollusca (mainly prosobranchs and bivalves) is listed that have the main part of their range in the tropical or subtropical Atlantic, and which extend northwards at least to southern Morocco. This list is necessarily a choice, because the biotic

Figure 3 Situation by year 2001 of the introduced species of other than tropical Indo–Paciﬁc origin. The introduced species on the Atlantic coastline, and some isolated occurrences of these in the Mediterranean, are not shown.

262

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

units in this area are not straightforward (see Ekman 1953: 81). Many of the species present on the coastline of northwest Africa extend northwards to temperate Europe (including the British Islands and Scandinavia) and the Mediterranean, and belong to the Lusitanian province. The list of Table 4 is based on the data used by Gofas (1999) for latitudinal ranges in northwest Europe and Morocco, from which have been deleted all species with a northern boundary north of 45ºN, all species with a southern boundary within Morocco or further north, and some widespread species with teleplanic larvae (e.g. tonnoideans). The remaining taxa represent a pool of tropical/subtropical species which are liable to enter the Mediterranean. Most tropical/subtropical species which occur in Atlantic Morocco, north of Agadir have had a limited range within the Mediterranean for a long time (Fig. 4), occurring on the coasts of Morocco (Pallary 1920), Algeria (Pallary 1900) and southern Spain (Hidalgo 1917). There are only a few exceptions to this: two (Osilinus sauciatus and Nassarius pfeifferi) are intertidal, have a sharp boundary at the Strait of Gibraltar and are barred from the Mediterranean by the lack of a suitable habitat, and a few more species (e.g. Solatia piscatoria or Marginella glabella) are established in Atlantic Morocco but do not enter the Mediterranean. The eastern boundary of the species which penetrate into the Alboran Sea may yet experience some variations. One example is Patella nigra, a large, prevalently West African limpet which also lives along the coasts of North Africa and southern Spain. On European shores, prior to 1998, this species was found only to the west of the city of Málaga, whereas it now thrives at Rincón de la Victoria, some 20 km further east. This is, however, an anthropophilous species, which rapidly builds up populations in perturbated sites, especially where new artiﬁcial rock piers are built. Another possible “natural” change in Mediterranean distributions regards the small prosobranch Tricolia miniata, originally described from the North

Figure 4 The occurrence, inside and near the Mediterranean, of Mollusca with a main range in tropical or subtropical West Africa. Most of the species reach a natural, stable limit near Gibraltar.

263

264 * * *

*

* * * * * * * * * * * * * * * * * * * * * *

* * * *

Sahara

* *

* *

* * *

*

* * * * * * * * * *

* *

* *

* *

* *

*

*

* * * * * * *

* * * *

*

* * * * *

*

*

* * * *

Alboran

* * * * *

*

N. Spain

* * *

Algarve

* *

Morocco

* * * *

* *

*

*

* *

* *

Algeria

*

*

*

*

*

Other W Med.

Distribution of tropical or subtropical species of Mollusca, occurring on the Moroccan coast and liable to occur in the Mediterranean Sea.

Class GASTROPODA, subclass Prosobranchia Patella nigra da Costa, 1771 Fissurella nubecula (Linné, 1758) Clanculus kraussi (Philippi, 1846) Osilinus sauciatus (Koch, 1845) Tricolia miniata (Monterosato, 1884) Littorina punctata (Gmelin, 1791) Mesalia brevialis (Lamarck, 1822) Dendropoma petraeum (Monterosato, 1884) Natica fanel Récluz, 1844 Natica vittata (Gmelin, 1791) Tectonatica ﬁlosa (Philippi, 1845) Sinum bifasciatum (Récluz, 1851) Zonaria pyrum (Gmelin, 1791) Erosaria spurca (Linné, 1758) Trivia bitou Pallary, 1912 Epitonium jolyi (Monterosato, 1878) Opalia crenata (Linné, 1758) Stramonita haemastoma (Linné, 1766) Orania fusulus (Brocchi, 1814) Nassarius goreensis (von Maltzan, 1884) Nassarius elatus (Gould, 1845) Nassarius heynemanni (von Maltzan, 1884) Nassarius denticulatus (Adams A., 1851) Nassarius pfeifferi (Philippi, 1844) Nassarius vaucheri (Pallary, 1906) Bullia miran (Bruguière, 1792) Demoulia obtusata (Link, 1807) Mitrella bruggeni van Aartsen, Menkhorst & Gittenberger, 1984 Mitrella broderipi (Sowerby G.B. I, 1844) Marginella glabella (Linné, 1758) Gibberula oryza (Lamarck, 1822) Gibberula secreta Monterosato, 1889

Table 4

SERGE GOFAS & ARGYRO ZENETOS

*

* * * * * * * * * * * * * * * * * * * * 56

Class GASTROPODA, subclass Pulmonata Siphonaria pectinata (Linné, 1758)

Class BIVALVIA Nuculana bicuspidata (Gould, 1845) Anadara corbuloides (Monterosato, 1878) Perna perna (Linné, 1758) Myoforceps aristatus (Dillwyn, 1817) Modiolus stultorum Jousseaume, 1891 Modiolus lulat Dautzenberg, 1891 Amygdalum agglutinans (Cantraine, 1835) Pinna rudis Linné, 1758 Ungulina cuneata (Spengler, 1782) Diplodonta brocchii (Deshayes, 1852) Scacchia zorni van Aartsen & Fehr de Wal, 1985 Digitaria digitaria (Linné, 1758) Cuna gambiensis Nicklès, 1955 Sinupharus combieri (Ficher-Piette & Nicklès, 1946) Tellina compressa Brocchi, 1814 Gari intermedia (Deshayes, 1855) Gari pseudoweinkaufﬁ Cosel, 1989 Eastonia rugosa (Helbling, 1779) Panopea glycimeris (von Born, 1778) Clavagella aperta Sowerby G.B. I, 1823

Totals

Sahara

* * * * *

continued

Gibberula caelata (Monterosato, 1877) Cymbium olla (Linné, 1758) Cymbium tritonis (Broderip, 1830) Cymbium cucumis Röding, 1798 Cancellaria cancellata (Linné, 1767) Solatia piscatoria (Gmelin, 1791) Fusiturris similis (Bivona And., 1838)

Table 4

265 * * * * * *

* * * * * * 47

* *

30

*

*

*

*

*

*

* *

Algarve

* * * * * * *

* * *

*

* * *

* *

Morocco

6

*

*

N. Spain

26

* * *

* * * * 38

*

*

* *

* * *

*

*

*

*

Algeria

*

*

* * * * *

* * *

*

*

*

* *

Alboran

5

Other W Med.

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

SERGE GOFAS & ARGYRO ZENETOS

African coast, but never mentioned from France in the nineteenth century literature. In 1988, the authors recorded large numbers of this species in Les Embiez, southern France. These isolated case histories can be balanced by others in the opposite direction. There are nineteenth century records regarding tropical Atlantic species which are nowadays established in the southern Iberian Peninsula but are not found further inside the Mediterranean. Examples are Cancellaria cancellata and Siphonaria pectinata, reported for the coast of France by Locard (1886). The prosobranch Sinum bifasciatum is well established on the coast of Málaga, but there are scanty nineteenth century records from the Balearics and southern Italy (see Rueda & Gofas 2000, and references therein). Several recent records of tropical West African species in Mediterranean localities have been disregarded since they were based on discards from ﬁshing boats. Examples are the records of West African Mesalia opalina in the Ionian Sea (Garavelli & Melone 1967) and of Cymbium cucumis in Sardinia (Giuseppetti et al. 1991). To summarise, the authors do not know of any tropical or subtropical Atlantic molluscan species recently found in the Mediterranean that could be suspected of being a newcomer, which has arrived from the Atlantic in a dynamic process comparable with lessepsian immigration. The possible variation in boundaries of the species listed in Table 4 should be the focus of ongoing studies, because some may expand their range inside the Mediterranean and this may be in response to long-term climatic change.

Summary of invasion pathways The processes inferred for the arrival of exotic species (without considering the questionable species) in the Mediterranean are summarised in Figure 5. The main process, termed “Erythrean invasion” by Galil & Zenetos (2002), is clearly the inﬂux from the Red Sea. These are not “introduced” species in the strict sense (i.e. man is responsible for the construction of the new waterway, but not the transport of the molluscs themselves). Shipping is the next most important vector and is extremely unpredictable. Among the 137 established or alien exotic molluscs of the Mediterranean, only 17 appear to have arrived or have been transported within the Mediterranean via shipping. Transport of live molluscs for mariculture was a vector for 12 species, of which six are deliberate introductions (two rock oysters: Crassostrea gigas and Saccostrea commercialis; two pearl oysters: Pinctada margaritifera and Pinctada radiata; and two clams: Mercenaria mercenaria and Ruditapes philippinarum); the others have been accidental introductions.

Possible reasons for success Much has already been written on invasive species, both plant and animal. Investigations have focused on the genetic background of the successful invaders, the breadth of trophic niche and the life-history strategy (Golani 1998). The molluscan data in the Mediterranean suggest that some important factors are associated with the characteristics of the recipient environment, and with the intensity and time of exposure to the immigrants. In the Mediterranean, 10 species are locally invasive. These are the gastropods Cerithium scabridum, Rhinoclavis kochi, Strombus persicus and Bursatella leachi, and the bivalves 266

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Figure 5 The inferred pathways for exotic molluscan species in the Mediterranean basin. Seven species (see Table 2) are considered twice because two different modes of introduction are assumed. The questionable and cryptogenic species are not considered.

Pinctada radiata and Brachidontes pharaonis in the eastern Mediterranean, the gastropod Rapana venosa, and the bivalves Anadara inaequivalvis, Musculista senhousia and Xenostrobus securis in the northern Adriatic and/or the western Mediterranean lagoons. There is a rule-of-thumb that for every 100 species brought alive to an exotic locality, 10 will settle and one may become invasive (Williamson & Fitter 1996). If this is applicable in the Mediterranean, some 1000 species may have had an opportunity to settle.

Ecological characters related to successful establishment There has been much debate on the life-history characteristics that make a successful invader (review in Morton 1996). Regarding Mediterranean exotics, Ritte & Pashtan (1982) and Lavie & Nevo (1986) tried to investigate a possible relationship between the proﬁle of successful colonisers and the amount of genetic variation evidenced by allozyme polymorphism but a clear picture did not emerge. Species which have been introduced or have invaded tend to occur in a similar manner elsewhere in the world. Among the species introduced in the Mediterranean, examples are Rapana venosa (see Mann et al. 2002), Musculista senhousia (see Willan 1987). Table 5 lists the Mediterranean exotic or cryptogenic species which are known to have been introduced (or are regarded as cryptogenic) in other areas, namely the Black Sea (Zaitsev & Ozturk 2001), North Sea (Reise et al. 1999), European Atlantic coast (Goulletquer et al. 2002), North American Atlantic coast (Carlton 1992, Mann et al. 2002), Paciﬁc coast of North America (Carlton 1992), and Australasia (Willan 1987). Some studies ﬁnd that exotics are “typical” opportunistic species, while others state that there are as many exceptions to this generalisation as there are examples supporting it. A good example of an opportunist that ﬁts the classical concept is that of Musculista senhousia. It is a short-lived mussel that suffers high mortality, can experience very high, but often variable, population size, is small in body size, grows quickly and has a long planktonic dispersal stage (Crooks 1996). It is depicted as an “opportunist which is likely to pre-empt food and space” (Willan 1987). Brachidontes pharaonis, another very successful mytilid, is described as an euryhaline species by Morton (1988). Pinctada radiata, a very 267

NW Atlantic NW Paciﬁc NW Paciﬁc SE Asia Mediterranean NW Paciﬁc NW Atlantic NW Paciﬁc NW Atlantic NW Atlantic SE Asia

Origin 1957 1974 1969 1964 native 1964 1965 1981 1994 1976 1800

Medit.

native 1819 1861 1973–4 1800 1800

1966 1800

1949 1998

NE Atlantic

native 1900s

1946 1982

Black Sea

Records of introduction of some non-lessepsian immigrants worldwide.

Crepidula fornicata Rapana venosa Anadara inaequivalvis Musculista senhousia Mytilus galloprovincialis Crassostrea gigas Mercenaria mercenaria Ruditapes philippinarum Petricola pholadiformis Mya arenaria Teredo navalis

Table 5

268

native native native 1800

1890 1800 1730

1998

NW Atlantic

1964 1864

1887

North Sea

1946 1913

1941 1880? 1932 1967 1936

1937

NE Paciﬁc

1980s

Australia N.Z.

SERGE GOFAS & ARGYRO ZENETOS

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

successful and early settler in the Mediterranean, is described as a sturdy, euryhaline and eurythermal species, which can stay alive after several days out of the water (Seurat 1929). The superfamily Cerithioidea, which includes many euryoecious shallow marine species, is also illustrative of the success of opportunistic species. Nevertheless, the family with the highest number of recorded immigrant species is the Pyramidellidae (Fig. 6), which are specialised animals, obligate ectoparasites of other invertebrates, mainly polychaetes and molluscs. There is a loose speciﬁcity of the parasite species to the host species, but a pyramidellid species will parasitise hosts only within a given phylum. The success of the pyramidellids as immigrants relies on their ability to adapt to new host species of the appropriate higher taxon. The pyramidellid Adalactaeon amoenus has been observed (Oliverio 1994) on Acanthocardia tuberculata, a native Mediterranean species which cannot be its original host in the Red Sea. Nevertheless, another factor in this case may be the high number of species present in the Red Sea (47 species according to Dekker & Orlin 2000, certainly underestimated). The exotic species experience changes in their niche. The newcomers may ﬁnd some companion species but will usually ﬁnd drastic changes in their biotoic environment compared with their place of origin. The small columbellid Anachis selasphora certainly was not an inhabitant of Posidonia meadows, but now thrives in the cavities of the rhizomes of Posidonia in southern Turkey (Tringali & Villa 1995). In the long run, such shifts may lead to evolutionary change and this will make the Mediterranean an extraordinary natural laboratory for the study of adaptive evolution.

Pattern of larval development The pattern of larval dispersal should have some bearing on colonisation success. Marine invertebrates, including molluscs, can be roughly divided into species with a pelagic larval

Figure 6 Ranking of molluscan families according to the number of immigrant species registered in the Mediterranean.

269

SERGE GOFAS & ARGYRO ZENETOS

Table 6 Numbers of species, among the prosobranchs listed in Table 2, according to their type of larval development and their establishment status.

Total 25

Planktotrophic Established 17

Invasive

Total

4

17

Non-planktotrophic Established Invasive 11

–

stage, and those with so-called direct development. The former remain several days to several weeks in the plankton, on which they feed before they settle down for adult life, and this makes them prone to a broad dispersal. The direct developers hatch directly as small benthic organisms and normally settle at the same place as their progenitors. However, there is no straightforward picture as to which are the more aggressive invaders. Among prosobranchs (Table 6), where the type of larval development is best assessed because it can be deduced from the morphology of the protoconch, more than half of the recorded species are planktotrophic, and this pattern holds true even when only the established species are considered. This ﬁgure is similar to the natural occurrence of planktotrophic species in temperate latitudes (Thorson 1950) and does not indicate that larval planktotrophy is a key factor for the success of immigration. It is nevertheless noteworthy that the invasive gastropods Cerithium scabridum, Rhinoclavis kochi, Strombus decorus, Rapana venosa and also the invasive bivalve Musculista senhousia have a planktotrophic development.

Factors related to the recipient sites Vermeij (1996) posed the question of how the assemblages rich in immigrants could differ from those in which few or no foreign species have become established. The data collected in the Mediterranean show that the occurrence of an immigrant component in the fauna is signiﬁcant in the eastern Mediterranean as a whole, in the Adriatic and in the western Mediterranean lagoons. In the eastern Mediterranean, the present day fauna is possibly not as diverse as its environment could support, because the area is not readily accessible to tropical Atlantic species and thus the niche for tropical/subtropical elements may be only partly occupied. In the western Mediterranean, there are only two reported lessepsians, and they did not arrive before the 1990s. These are the bivalve Fulvia fragilis found at Cullera, Valencia, Spain (collected 1991 by J. Trigo; original observation communicated by E. Rolán, Vigo) and the opisthobranch Bursatella leachi, reported in various points of the Tyrrhenian Sea and in Cagliari, Sardinia (original observation, communicated by A. Olita). The ﬁnding of the Indo–Paciﬁc nudibranch Chromodoris quadricolor in 1982 was an isolated occurrence. The occurrence of the other introduced species is limited to lagoons, to the coastal northern Adriatic, which is a secluded area and in this respect is comparable to lagoons, and to harbours. Any randomly taken sample in the open sea is currently free of exotic molluscs. The authors speculate that the rarity of exotic species in the western Mediterranean can be explained because that area has a higher species richness and is closer to the potential source in the Atlantic. According to this view, the western Mediterranean open-sea is saturated for the sustainable species richness compatible with the heterogeneity and the 270

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

resources of the environment. Thus, it is a difﬁcult place for the penetration of alien species. The same conclusion was reached by Carlton (1992: 500) regarding the open-sea communities of the American Paciﬁc coast. Conversely, lagoons and other marginal marine sites have low species richness compared with the surrounding open sea. Their environment is constrained by demanding physical factors, which make them unsuitable for most of the locally available open-sea species. Thus, lagoons are oligospeciﬁc because they have a limited pool of eligible species locally available. In this context, any species belonging to marginal marine biota from a remote area makes a very likely successful coloniser. Success in the new environments may be enhanced by the lack of natural competitors and enemies. The native species in general do not have means to compete with the aliens because they had not co-evolved with them.

Impact on the native fauna and prospects In those areas with high impact of immigrants, and where the native fauna was formerly species poor, the newcomers imply a major reorganisation of the faunal assemblages. The impact of the Indo–Paciﬁc fauna is restricted to the eastern basin of the Mediterranean, but there it is of major signiﬁcance. A comparison of the results of Tom & Galil (1991) with those of Gilat-Gottlieb (1959) revealed massive penetration of four Indo–Paciﬁc species, among which are the gastropods Rhinoclavis kochi and Minolia nedyma. Rhinoclavis kochi was ﬁrst reported from Haifa Bay in 1963, and in the 1970s large numbers were collected (Barash & Danin 1973). In a benthic sample collected off Haifa (leg. B. Galil, 9.1997), 15 species out of 119 (i.e. 12%) were lessepsian, and these comprised 270 individuals out of 894 (i.e. 30%). The most abundant species was Rhinoclavis kochi, and the second most abundant was also an immigrant, Finella pupoides. The Israeli coast holds 90 exotic species of Erythrean origin (25%) in a total of 372 species cited by Barash & Danin (1992). At some 300 km away from the Levantine coast, in a sample from Kyrenia, northern Cyprus (leg. Zibrowius 11.1998), six species out of 76 (7.9%) were lessepsian but they represent only 13 specimens out of 527 (2.5%) of the specimens. More globally, 32 out of 627 (5%) of the species of molluscs cited from Cyprus (Cecalupo & Quadri 1996, Buzzurro & Greppi 1997) are exotics. In Greece ﬁgures are still lower, with only 23 exotic species (2.3%) out of a total of 1095 (Delamotte & Vardala-Theodorou 2001). Data illustrated in Figures 1 and 2 (pp. 257, 259) indicate that the rate of inﬂux for migrants from the Red Sea remains steady with no sign of saturation. Both the total number of Red Sea species and the extension of the area colonised by at least one Red Sea species are increasing. There is still a pool of about 90 species (see Table 3, p. 255) reported from the Suez Canal which are likely to be found in the Mediterranean in the near future but none of these has been seen there yet. The total number of Red Sea species, amounting to a conservative ﬁgure of 1765, cited by Dekker & Orlin (2000), shows that the source area has as least as many species as the Mediterranean itself. Only two lessepsian immigrants have settled in the western Mediterranean, and they did not arrive before the 1990s. The number of such immigrants is likely to increase in the future. The non-lessepsian introductions, and especially those connected with mariculture, seem to have reached a peak in the 1960s (Crassostrea gigas and Mercenaria mercenaria, 271

SERGE GOFAS & ARGYRO ZENETOS

introduced deliberately; Rapana venosa, Polycerella emertoni, Anadara inaequivalvis, and Musculista senhousia, introduced inadvertently). Their later movements have been mainly inside the basin, namely, from the Black Sea to the Adriatic and thence to the western Mediterranean coastal lagoons. Nothing remains of the nineteenth century attempts to introduce certain molluscs for farming (the American oyster Crassostrea virginica shipped to Sicily, see di Monterosato (1915), and the pearl oyster Pinctada margaritifera, which was not mentioned by Seurat (1929)). In the 1980s, there were several non-lessepsian newcomers (Polycera hedgpethi, Chromodoris quadricolor, Melibe ﬁmbriata, Chlamys lischkei, Saccostrea commercialis, Ruditapes philippinarum) but only the last two are related to mariculture and only Melibe ﬁmbriata and Ruditapes philippinarum have become established. Displacements may occur, as can be observed on the Adriatic coast for R. decussatus driven out by R. philippinarum. In the 1990s, apart from the Red Sea migrants, Haminaea callidegenita and Xenostrobus securis were added to the Mediterranean fauna. The slow-down is real, and can be explained because the commercial routes for circulation of spat repeatedly carry the same species. Thus, those species from the current donor areas which were brought to the Mediterranean, and were likely to settle, became established exotics, whereas others that are regularly presented but cannot survive (e.g. Littorina littorea, several species of Gibbula from the European Atlantic coast) remain barred. A new wave of immigrants is to be expected only if new sources of species are made available. The most unpredictable source of non-native species remains transportation by ships, both through ballast waters, on hulls, and as discards from ﬁshing operations. For these introductions, the limiting factors (i.e. space where physical factors and/or the native fauna will not bar the immigrants) are the same as for mariculture but the potential source areas are virtually unlimited. However alarming this may look, the situation is still such that there is no recorded extinction of a marine species in the Mediterranean by anthropogenic means. Furthermore, among the factors of extinction in the marine realm (see Carlton et al. 1999, Roberts & Hawkins 1999), the impact of introduced species comes far behind those of overexploitation and of urban development.

References Aharoni, J. 1934. From Asquelon to Rubin. Nature and Country 2, 473–476 (cited in Barash & Danin 1973). Barash, A. & Danin, Z. 1973. The Indo-Paciﬁc species of mollusca in the Mediterranean and notes on a collection from the Suez Canal. Israel Journal of Zoology 21, 301–374. Barash, A. & Danin, Z. 1992. Fauna Palaestina: Mollusca I. Annotated list of Mediterranean molluscs of Israel and Sinai. Jerusalem: The Israel Academy of Sciences and Humanities. Barsotti, G. & Campani, E. 1982. Il promontorio di Castiglioncello (LI): III. Rinvenimento di una popolazione di Littorina littorea (L.) – Moll. Gastropoda Prosobranchia. Quaderni del Museo di Storia Naturale di Livorno 3, 65–71. Bazzocchi, P. 1985. Prima segnalazione di Strombus (Conomurex) decorus raybaudii Nicolay and Romagna-Manoja, 1983 per l’Isola di Cipro. Bollettino Malacologico 21, 64 only. B[ellet], D. 1899. La culture des huîtres perlières en Italie. La Nature (Paris: Masson), 27(1355), 375 only. Bini, G. 1983. Immissione antropica di molluschi esotici nel Mediterraneo: I – Il lago di Caprolace. Studi per l’Ecologia del Quaternario, Firenze 5, 113–125.

272

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Bogi, C. & Khairallah, N. H. 1987. Nota su alcuni molluschi de provenienza Indo-Paciﬁca raccolti nella baia di Jounieh (Libano) – Contributo I-. Notiziario del CISMA 10, 54–60. Bombace, G. 2001. Inﬂuence of climatic changes on stocks, ﬁsh species and marine ecosystems in the Mediterranean Sea. Archivio di Oceanograﬁa e Limnologia 22, 67–72. Bouchet, P. & Gofas, S. 1983. “Terebra” cosentini Philippi, 1836 an American columbellid species. The Nautilus 97, 26–28. Breber, P. 2002. Introduction and acclimatisation of the Paciﬁc carpet clam, Tapes philippinarum, to Italian Waters. In Invasive aquatic species of Europe: distributions, impacts and management, E. Leppäkoski et al. (eds). Dordrecht: Kluwer Scientiﬁc, 120–126. Buzzurro, G. & Greppi, E. 1997. Notes on the molluscs of Cyprus, with special attention to the alloctone species. La Conchiglia 283, 21–31, 61–62. Carlton, J. T. 1992. Introduced and estuarine mollusks of North America: an end-of-the-20th-century perspective. Journal of Shellﬁsh Research 11, 489–505. Carlton, J. T. 1996a. Marine bioinvasions: the alteration of marine ecosystems by nonindigenous species. Oceanography 9, 36–43. Carlton, J. T. 1996b. Biological invasions and cryptogenic species. Ecology 77, 1653–1655. Carlton, J. T. & Geller, J. B. 1993. Ecological roulette: the global transport of nonindigenous marine organisms. Science 261, 78–82. Carlton, J. T., Geller, J. B., Reaka-Kudla, M. L. & Norse, E. A. 1999. Historical extinctions at sea. Annual Review of Ecology and Systematics 30, 515–538. Cattaneo-Vietti, R. 1986. On the probable presence of Chromodoris quadricolor quadricolor (Mollusca: Nudibranchia) in the Mediterranean Sea. Bollettino Malacologico 22, 167–168. Cecalupo, A. & Quadri, P. 1996. Contributo alla conoscenza malacologica per il nord dell’ isola di Cipro. Bollettino Malacologico 31, 95–118. Cesari, P. & Pellizzato, M. 1985. Molluschi pervenuti in Laguna di Venezia per apporti volontari o casuali. Acclimazione di Saccostrea commercialis (Iredale and Roughley, 1933) e di Tapes philippinarum (Adams and Reeve, 1850). Bollettino Malacologico 21, 237–274. Clanzig, S. 1989. Invertébrés d’introduction récente dans les lagunes Mediterranéennes du LanguedocRoussilon (France). Bulletin de la Societé Zoologique de France 114, 151–152. Crooks, J. A. 1996. The population ecology of an exotic mussel Musculista senhousia, in a southern California Bay. Estuaries 19, 42–50. Dekker, H. & Orlin, Z. 2000. Check list of Red Sea Mollusca. Spirula 47 (Suppl.), 1–46. Delamotte, M. & Vardala-Theodorou, E. 2001. Shells from the Greek seas (2nd edn). Athens: Goulandris Natural History Museum. De Min, R. & Vio, E. 1998. Exotic molluscs in the northern Adriatic. Annals for Istran and Mediterranean Studies, Annales 13, Serie historia naturalis 3, 43–54. Di Geronimo, I. 1971. Prima segnalazione sulle costa italiane di Brachiodontes variabilis (Krauss). Bollettino delle sedute dell’ Accademia Gioenia di Scienze Naturali in Catania 10, 847–852. di Monterosato, T. A. 1878. Enumerazione e sinonima delle conchiglie mediterranee. Giornale di Scienze Naturali ed Economiche di Palermo 13, 61–115. di Monterosato, T. A. 1880. Notizie intorno ad alcune conchiglie delle coste d’Africa. Bullettino della Società Malacologica Italiana 5, 213–233. di Monterosato, T. A. 1884. Nomenclatura generica e speciﬁca di alcune conchiglie mediterranee, Palermo: Virzi. di Monterosato, T. A. 1899. Coquilles marines de Chypre. Journal de Conchyliologie 47, 392–401. di Monterosato, T. A. 1915. Ostreae ed Anomiae del Mediterraneo. Annali del Museo Civico di Storia Naturale di Genova (3a)7, 7–16, pl. 1–4. Di Natale, A. 1978. Note sur la présence et la répartition de Cerithium scabridum Philippi, 1849 (Mollusca, Gastropoda) sur les côtes siciliennes. Bulletin de l’Ofﬁce National des Pêches de Tunisie 2, 193–198. Ekman, S. 1953. Zoogeography of the sea. London: Sidgwick & Jackson.

273

SERGE GOFAS & ARGYRO ZENETOS

Elton, C. S. 1958. The ecology of invasions by animals and plants. London: Methuen. El Sayed, S. & van Dijken, G. L. 1995. The southeastern Mediterranean ecosystem revisited: thirty years after the construction of the Aswan High Dam. Quarterdeck (Department of Oceanography, Texas A&M University) 3, 1–4. Fabioux, C., Huvet, A., Lapègue, S., Heurtebise, S. & Boudry, P. 2002. Past and present geographical distribution of populations of Portuguese (Crassostrea angulata) and Paciﬁc (C. gigas) oysters along the European and North African Atlantic coasts. Haliotis 31, 33–44. Fischer, P. 1865. Note sur les faunes conchyliologiques des deux rivages de l’Isthme de Suez. Journal de Conchyliologie 13, 241–248. Francour, P., Boudouresque, C. F., Harmelin, J. G., Harmelin-Vivien, M. L. & Quignard, J. P. 1994. Are the Mediterranean waters becoming warmer? Information from biological indicators. Marine Pollution Bulletin 28, 523–526. Gaglini, A. 1987. Nota su Epidromus gladiolus Monterosato (Aspella anceps Lamarck). Notiziario del CISMA 10, 1–2. Galil, B. & Zenetos, A. 2002. A sea change – exotics in the eastern Mediterranean. In Invasive aquatic species of Europe: distributions, impacts and management, E. Leppäkoski et al. (eds.). Dordrecht: Kluwer Scientiﬁc, 325–336. Gangnery, A., Bacher, C. & Buestel, D. 2001. Assessing the production and the impact of cultivated oysters in the Thau lagoon (Mediterranée, France) with a population dynamics model. Canadian Journal of Fisheries and Aquatic Science 58, 1012–1020. Garavelli, C.L. & Melone, N. 1967. Ritrovamenti malacologici nel Mediterraneo. Conchiglie 3, 132–148. Gilat-Gottlieb, E. 1959. Study on the benthos in Haifa Bay. Ecology and zoogeography of invertebrates. Haifa: Special Publication of Sea and Fisheries Research Station. Giuseppetti, G., Silesu, M., Manunza, B. & Doneddu, M. 1991. Finding of Cymbium cucumis Roeding, 1798 in the Gulf of Cagliari. La Conchiglia 260, 46–47. Gofas, S. 1999. Marine molluscs with a very small range in the Strait of Gibraltar. Diversity and Distributions 4, 255–266. Golani, D. 1998. Impact of Red Sea ﬁsh migrants through the Suez Canal on the aquatic environment of the Eastern Mediterranean. Bulletin series, Yale School of Forestry and Environmental Studies 103, 375–387. Golani, D., Orsi-Relini, L., Massuti, E. & Quignard, J. P. 2002. CIESM Atlas of exotic species in the Mediterranean. Volume 1. Fishes, F. Briand (ed.). Monaco: CIESM. Goulletquer, P., Bachelet, G., Sauriau, P. G. & Noel, P. 2002. Open Atlantic coast of Europe – a century of introduced species into French waters. In Invasive aquatic species of Europe: distributions, impacts and management, E. Leppäkoski et al. (eds). Dordrecht: Kluwer Scientiﬁc, 276–290. Gruvel, A. & Moazzo, G. 1931. Contribution à la faune malacologique marine des côtes LibanoSyriennes. In Les etats de Syrie. Richesses marines et ﬂuviales, A. Gruvel (ed.). Paris: Société des Editions Géographiques, Maritimes et Coloniales, 437–453. Haas, G. 1937. Mollusca marina. In Prodromus faunae Palestinae, F. S. Bodenheimer (ed.). Cairo: Mémoires de l’Institut d’Egypte, 275–280. Haas, G. 1948. Sur l’immigration de mollusques de l’Indo-Paciﬁque dans les eaux côtières de la Palestine. Journal de Conchyliologie 88, 141–144. Hasan, A. K. 1974. Studies on bottom Mollusca (gastropods and bivalves) in Abou-Kir Bay. MSc thesis, Faculty of Science, University of Alexandria, Egypt. Hidalgo, J. G. 1917. Fauna malacologica de España, Portugal y las Baleares – Moluscos testaceos marinos. Trabajos del Museo Nacional de Ciencias Naturales (Série Zoológica) 30, 1–752. Hoenselaar, H. J. & Hoenselaar, J. 1989. Musculista senhousia (Benson in Cantor, 1842) in the western Mediterranean (Bivalvia, Mytilidae). Basteria 53, 73–76. Hoppe, K. 2002. Teredo navalis – the cryptogenic shipworm. In Invasive aquatic species of Europe: distributions, impacts and management, E. Leppäkoski et al. (eds.). Dordrecht: Kluwer Scientiﬁc, 116–119.

274

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Houart, R. & Vokes, E. H. 1995. On the identity of Aspella anceps (Lamarck, 1822) (Gastropoda, Muricidae). Bollettino Malacologico 31, 13–16. Hsü, K. J., Ryan, W. B. F. & Cita, M. B. 1973. Late Miocene desiccation of the Mediterranean. Nature 242, 240–244. Issel, A. 1869. Malacologia del Mar Rosso. Pisa: Biblioteca Malacologica. Jablonski, D. & Lutz, R. A. 1980. Molluscan larval shell morphology. Ecological and paleontological applications. In Skeletal growth of aquatic organisms, D. C. Rhoads & R. A. Lutz (eds.). New York: Plenum Press, 323–377. Janson, K. 1985. A morphologic and genetic analysis of Littorina saxatilis (Prosobranchia) from Venice, and on the problem of saxatilis-rudis nomenclature. Biological Journal of the Linnean Society 24, 51–59. Johannesson, K. 1988. The paradox of Rockall: why is a brooding gastropod (Littorina saxatilis) more widespread than one having a planktonic larval dispersal stage (L. littorea)? Marine Biology 99, 507–513. Keller, C. 1882. Die Fauna im Suez Kanal und die Diffusion der Mediterranen und Erythräischen Thierwelt. Eine thiergeographische Untersuchung. Neue Denkschriften der allgemeinen Schweizerischen Gesellschaft fur die gesammten Naturwissenschaften, Basel, 28, 1–39. Lamarck, [J. B. de] 1819. Histoire naturelle des animaux sans vertèbres, Vol. 6(1). Paris: published by the Author. Lambert, L., Faideau, F. & Bluteau, R. 1929. Ostréiculture et Mytiliculture sur le littoral centre-ouest. La Rochelle: Imprimerie Jean Foucher. Lavie, B. & Nevo, E. 1986. Genetic diversity of marine gastropods: contrasting strategies of Cerithium rupestre and Cerithium scabridum in the Mediterranean Sea. Marine Ecology Progress Series 28, 99–103. Lazzari, G. & Rinaldi, E. 1994. Alcune considerazioni sulla presenza di specie extra Mediterranee nelle lagune salmastre di Ravenna. Bollettino Malacologico 30, 195–202. Levin, L. A. & Bridges, T. S. 1995. Pattern and diversity in reproduction and development. In Ecology of marine invertebrate larvae, L. McEdward (ed.). Boca Raton: CRC Press, 1–48. Locard, A. 1886. Prodrome de malacologie française. Catalogue général des Mollusques vivants de France. Mollusque Marins. Paris: Baillière. MacAndrew, R. 1870. Report on the testaceous Mollusca obtained during a dredging-excursion in the Gulf of Suez in the months of February and March 1869. Annals and Magazine of Natural History 6, 429–450. Mann, R., Occhipinti, A. & Harding, J. M. 2002. ICES special advisory report on the current status of invasions by the marine gastropod Rapana venosa. In ICES Advisory Committee on the Marine Environment. Report of the Working Group on Introductions and Transfers of Marine Organisms, Gothenburg, Sweden 20–22 March 2002, Copenhagen: International Council for the Exploration of the Sea, 117–134. McKenzie, J. A. 1999. From desert to deluge in the Mediterranean. Nature 400, 613–614. Menzel, R. W. 1974. Portuguese and Japanese oysters are the same species. Journal of the Fisheries Research Board of Canada 31, 453–456. Mienis, H. K. 1984. Strombus decorus persicus found in Israel. Hawaiian Shell News 32, 4 only. Mienis, H. K. 1985. Is Alvania dorbignyi (Audouin, 1826) a lessepsian migrant? Levantina 59, 652 only. Mienis, H. K. 1999. Strombus persicus on the ﬁshmarket of Yafo, Israel. De Kreukel 35, 112 only. Mizzan, L.1999. Le species alloctone del macrozoobenthos della Laguna di Venezia: il punto della situazione. Bollettino del Museo Civico di Storia Naturale di Venezia 49, 145–177. Moazzo, P. G. 1939. Mollusques testacés marins du Canal de Suez. Mémoires de l’Institut d’Egypte 38, 1–283. Morri, C., Peirano, A. & Bianchi, C. N. 2001. Is the Mediterranean coral Cladocora caespitosa an indicator of climatic change? Archivio di Oceanograﬁa e Limnologia 22, 139–144.

275

SERGE GOFAS & ARGYRO ZENETOS

Morrison, J. P. E. 1972. Mediterranean Siphonaria: West and East – old and new. Argamon 3, 51–62. Morton, B. 1988. The population dynamics and reproductive cycle of Brachidontes variabilis in a Hong Kong mangrove. Malacological Review 21, 109–117. Morton, B. 1996. The aquatic nuisance species problem: a global perspective and review. In Zebra mussels and other aquatic nuisance species, F. D’itri (ed.). Ann Arbor: Ann Arbor Press, 1–54. Nicolay, K. 1980. Siphonaria pectinata L., 1758 (Pulmonata, Siphonariidae). Prima segnalazione per le acque della Grecia. La Conchiglia 130–131, 11 only. Nicolay, K. & Romagna-Manoja, E. 1983. Strombus (Conomurex) decorus raybaudii n. ssp. La Conchiglia 174–175, 20–21. Occhipinti-Ambrogi, A. 2000. Biotic invasions in a Mediterranean Lagoon. Biological invasions 2, 165–176. O’Donoghue, C. H. 1929. Report on the Opisthobranchiata. Zoological Results of the Cambridge Expedition to the Suez Canal, 1924. Transactions of the Zoological Society of London 22, 713–841. O’Donoghue, C. H. & White, K. M. 1940. A collection of marine molluscs, mainly opisthobranchs, from Palestina. Proceedings of the Malacological Society of London 24, 92–96. Ó Foighil, D., Gaffney, P. M. & Hilbish, T. J. 1998. Mitochondrial gene sequences support an Asian origin for the Portuguese oyster, Crassostrea angulata (Lamarck, 1819). Marine Biology 131, 497–503. Oliverio, M. 1994. On the record of a living lessepsian Pyramidellid from Turkey (Heterobranchia, Heterostropha). Notiziario del CISMA 15, 79–82. Oliverio, M. 1995. The status of the living Mediterranean Strombus or: What is a lessepsian migrant? Notiziario del CISMA 16, 35–40. Olivi, G. 1792. Zoologia Adriatica. Venice: Bassano. Oren, O. H. 1969. Oceanographic and biological inﬂuence of the Suez Canal, the Nile and the Aswan Dam on the Levant Basin. Progress in Oceanography 5, 161–167. Pallary, P. 1900. Coquilles marines du littoral du Départment d’Oran. Journal de Conchyliologie 48, 211–422. Pallary, P. 1904–6. Addition à la faune malacologique du Golfe de Gabès. Journal de Conchyliologie 52, 212–248; 54, 77–124. Pallary, P. 1907. Sur l’extension de la faune équatoriale au nord-ouest de l’Afrique, et réﬂexions sur la faune conchyliologique de la Méditerranée. Bulletin Scientiﬁque de la France et de la Belgique 41, 421–427. Pallary, P. 1912. Catalogue des mollusques du littoral méditerranéen de l’Egypte. Mémoires de l’Institut d’Egypte 7, 1–139. Pallary, [P.] 1920. Exploration scientiﬁque du Maroc organisée par la Société de Géographie de Paris et continuée par la Société des Sciences Naturelles du Maroc. Deuxième fascicule. Malacologie (1912). Rabat and Paris: Larose. Pallary, P. 1938. Les mollusques marins de la Syrie. Journal de Conchyliologie 82, 5–57. Pelorce, J. 1995. Un débarquement paciﬁque sur les cotes du Golfe du Lion! Xenophora 71, 5 only. Petit de la Saussaye, S. 1869. Catalogue des Mollusques Testacés des Mers d’Europe. Paris: F. Savy. Plaziat, J. C. 1989. Signiﬁcation écologique et paléogéographique des peuplements oligotypiques de Potamides Gastéropodes thalassiques. In Atti 3° simposio di Ecologia e Paleoecologia delle communità bentoniche, Catania-Taormina 12–16 ottobre 1985, I. Di Geronimo (ed.). Catania: Università di Catania, 25–52. Por, F. D. 1978. Lessepsian migration: the inﬂux of Red Sea biota into the Mediterranean by way of the Suez Canal. Ecological Studies 23, Berlin: Springer, 1–228. Raimbault, R. 1964. Croissance des huîtres atlantiques élevées dans les eaux méditerranéennes françaises. Science et Pêche 126, 1–10. Rees, W. J. 1965. The aerial dispersal of Mollusca. Proceedings of the Malacological Society of London 36, 269–282. Reid, D. G. 1996. Systematics and evolution of Littorina. London: The Ray Society.

276

EXOTIC MOLLUSCS IN THE MEDITERRANEAN BASIN

Reise, K., Gollash, S. & Wolff, W. J. 1999. Introduced marine species of the North Sea coasts. Helgoländer Meeresuntersuchungen 52, 219–234. Ritte, U. & Pashtan, A. 1982. Extreme levels of genetic variability in two Red Sea Cerithium species (Gastropoda, Cerithiidae). Evolution 36, 403–407. Robba, E. 1987. The ﬁnal occlusion of Tethys: its bearing on Mediterranean benthic mollusks. In Shallow Tethys, K. G. McKenzie (ed.). Rotterdam: Balkema, 405–426. Roberts, C. M. & Hawkins, J. P. 1999. Extinction risk in the sea. Trends in Ecology and Evolution 14, 241–246. Rueda, J. L. & Gofas, S. 2000. Sinum bifasciatum (Récluz, 1851) (Gastropoda: Naticidae) conﬁrmed in Mediterranean fauna. Journal of Conchology 36, 81–82. Sabelli, B., Giannuzzi-Savelli, R. & Bedulli, D. 1990. Catalogo annotato dei molluschi marini del Mediterraneo. Vol. I. Bologna: Libreria Naturalistica Bolognese. Selli, G. 1733. Historia naturalis Teredinis, seu Xylophagi marini tubuloconchoidis speciatim Belgici. Arnhem: Christof Nebe. Seurat, L. G. 1929. La petite Pintadine du Golfe de Gabès. Bulletin des travaux publiés par la Station d’Aquaculture et de Pêche de Castiglione 1, 9–28. Steuer, A. 1939. Die Fischereigründe vor Alexandrien. XIX. Mollusca. Thalassia 3, 3–143. Taviani, M. 1994. Tridacne nel Pliocene Italiano. Bollettino Malacologico 29, 322–326. Thorson, G. 1950. Reproductive and larval ecology of marine bottom invertebrates. Biological Reviews 25, 1–45. Tillier, L. & Bavay, A. 1905. Les mollusques testacés du Canal de Suez. Bulletin de la Société Zoologique de France 30, 170–181. Tom, M. & Galil, B. 1991. The macrobenthic associations of Haifa Bay, Mediterranean coast of Israel. Marine Ecology 12, 75–86. Tomlin, J. R. le B. 1927. Report on the Mollusca (Amphineura, Gastropoda, Scaphopoda, Pelecypoda). Zoological Results of the Cambridge Expedition to the Suez Canal, 1924. Transactions of the Zoological Society of London 22, 291–320. Tringali, L. & Villa, R. 1995. On the identity of the Lessepsian species of the genus Anachis H. and A. Adams, 1853, subgenus Zafra A. Adams, 1860 found in the Levant Sea. La Conchiglia 274, 15–19. Vaillant, E. 1865. Recherches sur la faune malacologique de la Baie de Suez. Journal de Conchyliologie 13, 97–127. Verhecken, A. 1984. Strombus decorus raybaudii in de Middellandse Zee. Gloria Maris 23, 79–88. Vermeij, G. J. 1996. An agenda for invasion biology. Biological Conservation 78, 3–9. Willan, R. C. 1987. The mussel Musculista senhousia in Australasia; another aggressive alien highlights the need for quarantine at ports. Bulletin of Marine Science 41, 475–489. Williamson, M. & Fitter, A. 1996. The varying success of invaders. Ecology 77, 1661–1666. Zaitsev, Y. & Ozturk, B. 2001. Exotic species in the Aegean, Marmara, Black, Azov and Caspian Seas. Turkish Marine Research Foundation Publication 8, 1–267. Zenetos, A., Gofas, S., Russo, G. & Templado, J. in press. CIESM Atlas of exotic species in the Mediterranean. Vol. 3: Molluscs, F. Briand (ed.). Monaco: CIESM. Zibrowius, H. 1992. Ongoing modiﬁcations of the Mediterranean marine fauna and ﬂora by the establishment of exotic species. Mesogée 51, 83–107.

277

Oceanography and Marine Biology: an Annual Review 2003, 41, 279–309 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES SHAUN K. WILSON, 1,3 DAVID R. BELLWOOD, 1 J. HOWARD CHOAT, 1 & MILES J. FURNAS 2 1 Department of Marine Biology and Aquaculture, James Cook University, Townsville, Qld 4811, Australia e-mail: swilson@ﬁeldstudies.org (corresponding author) 2 Australian Institute of Marine Science, PMB 3 Townsville MC, Qld 4810, Australia 3 Current address: School for Field Studies, South Caicos, Turks & Caicos Islands, British West Indies

Abstract The epilithic algal matrix (EAM) is a ubiquitous component of coral reefs and is the primary grazing surface for many reef ﬁshes. Detritus accounts for at least 10% to 78% of all the organic matter present in the EAM, variation being attributed to hydrodynamic forces such as wave energy and biological elements such as algal morphology. When compared with ﬁlamentous algae, the other major source of organic matter in the EAM, protein : energy ratios, C : N ratios and total hydrolysable amino acids all suggest that detritus is of higher nutritional value than the algae. Lipid biomarkers indicate that more than 70% of the detritus is derived from the ﬁlamentous algae but the addition of bacteria and microalgae add essential nutrients and improve the nutritional value of the detritus. The detritus is typically of an amorphic form with protein : energy ratios which indicate that it is capable of sustaining ﬁsh growth. Detritus within the EAM may be derived from dissolved organic matter, which reduces refractory material, enhancing the palatability and digestibility of detritus relative to ﬁlamentous algae. Detritus in the EAM may also come from settling material and ﬁsh faeces. Studies that quantiﬁed the amount of detritus ingested by ﬁshes have identiﬁed at least 24 species from ﬁve families that predominantly ingest detritus. These species represent some of the most widespread and abundant EAM feeding ﬁshes on coral reefs. It is estimated that detritivorous ﬁshes account for at least 20% of individuals and 40% of the biomass of an EAMfeeding ﬁsh assemblage on the Great Barrier Reef. Comparisons of ingested material with the EAM indicate that many of these species selectively feed on detritus, particularly the small, organic rich particles 125 m. Furthermore, analysis of lipids in body tissues of blennies and assimilation of nutrients from the alimentary canal of scarids and acanthurids provide strong evidence that detritus is assimilated by coral reef ﬁshes. Consequently, a large percentage of EAM-feeding ﬁshes on coral reefs can unequivocally be classiﬁed as detritivores. The ingestion and assimilation of detritus by these ﬁshes represents a signiﬁcant pathway for transferring energy from within the EAM to secondary consumers, making detritivorous ﬁshes a critically important component of coral reef trophodynamics.

279

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Introduction Understanding the pattern of energy ﬂow between trophic levels is a fundamental goal of ecology. In aquatic ecosystems energy ﬁxed by primary producers is transferred to other trophic levels via the grazing of herbivores, or the detrital pathway (Begon et al. 1990). For coral reef ecosystems, ﬁshes that feed on epilithic algal communities (EACs) are an important component of food webs, providing a major trophic link between primary production by algae and secondary consumers. It is often assumed that turf-forming ﬁlamentous algae are the primary source of nutrition for these ﬁshes (Horn 1989, Polunin 1996, Hatcher 1997). However, it is well recognised that the EAC is a complex structure, in which invertebrates, detritus, microbes and microalgae co-occur and all represent potential sources of nutrition (Hatcher 1983, Choat 1991). Consequently, this conglomeration of living and non-living resources may be better referred to as an epilithic algal matrix (EAM) (Wilson & Bellwood 1997). Fishes feeding on the EAM can not necessarily be assigned the trophic status of herbivore, as they may ingest, digest and assimilate dietary items other than algae. The relative importance of the different EAM components to a ﬁsh’s diet has important implications for our understanding of coral reef trophodynamics. A major part of the energy ﬁxed by algae on coral reefs passes through detrital food webs (Hatcher 1983, Alongi 1989). Indeed, a trophodynamic model of a coral reef at Moorea, French Polynesia, estimated that 59–69% of net primary production enters detrital pools (Arias-Gonzalez et al. 1997). This high estimate is supported by observations recording large inputs of organic carbon to a coral reef lagoon from detritus relative to primary productivity (Hansen et al. 1992), and a higher standing biomass of detritus relative to ﬁlamentous algae in pomacentrid territories (Wilson & Bellwood 1997). Detrital production and standing biomass are therefore substantial on coral reefs and detritus represents a potentially valuable dietary resource. The detritus produced on reefs may contribute to further algal production, become incorporated into sediments and buried, be transported off reefs, or consumed by micro-organisms and detritivores (Hatcher 1983). Micro-organisms, in particular bacteria, transform or consume detritus, making it more digestible for metazoan consumers (Alongi 1989). Similarly, benthic and pelagic invertebrates have been identiﬁed as consumers of detritus (Alongi 1989) and, in conjunction with microbial organisms, provide a link between detritus and coral reef ﬁshes. It has also been recognised that some coral reef ﬁshes ingest detritus but the quantitative implications of this feeding mode are not well resolved. This is despite ﬁndings that detritivorous ﬁshes are a major component of ﬁsh assemblages in fresh water ecosystems (Gerking 1994) and that some of these ﬁshes are capable of assimilating detritus, without bacterial or invertebrate intermediates (Bowen 1981). Previous reviews by Hatcher (1983) and Alongi (1989) have examined detrital pathways in coral reef ecosystems. However, as most of the published work at the time concentrated on detritus in the water column or loose sediments, there is little reference to detritus in the EAM and detritivory by coral reef ﬁshes. This review redresses the balance and provides an overview of recent ﬁndings. It reveals the extent to which detritus is ingested and assimilated by coral reef ﬁshes and demonstrates how these ﬁndings may alter our understanding of coral reef food webs. The review will concentrate on the detritus within the EAM, as this is the primary substratum on which herbivorous/detritivorous ﬁshes feed (Hatcher 1983, Choat 1991). The ﬁrst part of the review speciﬁcally looks at detritus, discussing sources, nutritional quality, distribution and processes that may lead to its formation. The second part of 280

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

the review looks at the coral reef ﬁshes that feed on detritus, some of the feeding characteristics of these species and their contribution to coral reef ﬁsh assemblages.

What is detritus? In the strictest sense, detritus is deﬁned as non-living organic matter derived from non-fossil living sources (Begon et al. 1990). In reality, this deﬁnition is rarely used because “detritus” often contains numerous microbes, such as bacteria, protozoans, algae and fungi. Most working deﬁnitions of detritus therefore, incorporate these heterotrophic and autotrophic microbes (e.g. Bowen 1979, Wilson 2000, Crossman et al. 2001). In aquatic ecosystems, ecologists have also made a distinction between morphic and amorphic forms of detritus. Morphic detritus, originates from decaying organic matter, contains cellular remnants of its organic origin and is typically composed of particles 100 m. In contrast, amorphic detritus is believed to originate principally from dissolved organic matter (DOM) and organic colloids, is of irregular shape, lacks cellular structures and is typically composed of particles 100 m (Bowen 1984). Amorphic detritus derived from DOM typically contains less refractory organic matter and is likely to be more digestible than morphic detritus (Bowen 1984). Consequently, it is often suggested that amorphic detritus has a higher nutritional value than morphic detritus. In a direct comparison, amorphic detritus from ﬁve different species of macroalgae had a higher protein content and lower C : N ratios than morphic detritus (Alber & Valiela 1994). Furthermore, sheepshead minnows, Cyprinodon variegatus, fed amorphic detritus derived from marine algae, assimilated approximately 4% of total body nitrogen per day from the detritus (D’Avanzo et al. 1991). This was 10 to 40 times the nitrogen assimilation rate from morphic detritus derived from the same algal species (D’Avanzo & Valiela 1990). Amorphic detritus, therefore, appears to be of high nutritional quality and better assimilated than morphic detritus. On coral reefs, detrital aggregates collected from the EAM within the territories of three pomacentrid and one blenny species had no distinct shape or cellular structure and were predominantly composed of particles 125 m (Wilson & Bellwood 1997, Wilson 2000). The appearance and size of these aggregates suggests they are amorphic detritus and that detritus in the EAM of these ﬁsh territories represents a valuable dietary resource.

How much detritus is there? Previous research has concentrated on detritus in soft sediments and the water column, however, recent data show hard substrata to be a major source of detritus. The EAM covers large areas of hard reef substrata (McClanahan et al. 2001a,b, Klumpp & McKinnon 1992) and has high concentrations per unit area of detritus (Table 1). Consequently, detritus in the EAM represents a potentially massive reservoir of organic matter on coral reefs. The contribution of detritus to the organic content of the EAM is substantial, although the relative amount of detritus in the EAM may vary considerably among and within reef zones of differing wave exposure. Studies at various locations on the Great Barrier Reef (GBR) 281

Lagoon Lagoon

Lizard Island Lizard Island

Lagoon Windward base Windward crest Windward fore ﬂat Windward mid ﬂat Exposed crest Exposed crest Exposed crest Exposed crest Sheltered crest

Lagoon

Lizard Island

Lizard Island Lizard Island Lizard Island Lizard Island Lizard Island Outer shelf GBR Outer shelf GBR Mid shelf GBR Mid shelf GBR Mid shelf GBR

Reef zone

Dischistodus perspicillatus Stegastes nigricans Hemiglyphidodon plagiometopon Salarias patzneri Open EAM Open EAM Open EAM Open EAM Open EAM Acanthurus lineatus Open EAM Acanthurus lineatus Open EAM

Fish territory or open EAM

125 20 20 20 20 63 63 63 63 63

10–125 10–125

10–125

Particle size sampled (m)

21 4 14 4 4 15.7 19.8 21.5 19.3 17.5

46 34

27

Organic content of particulates (% of total weight)

Contribution of detritus to the epilithic algal matrix (EAM). OM Organic matter.

Location

Table 1

53.7 50–78 12–22 27–45 42–50 10.2 11.1 27.4 33.1 33.4

60 54

46

% detritus in total EAM OM

Wilson 2000 Purcell & Bellwood 2001 Purcell & Bellwood 2001 Purcell & Bellwood 2001 Purcell & Bellwood 2001 Crossman et al. 2001 Crossman et al. 2001 Crossman et al. 2001 Crossman et al. 2001 Crossman et al. 2001

Wilson & Bellwood 1997 Wilson & Bellwood 1997

Wilson & Bellwood 1997

Reference

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

282

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

indicate that detritus comprises 10% to 78% of the total organic content of the EAM (Table 1). Unfortunately, a variety of factors make it difﬁcult to compare studies and assess the proportion of amorphic detritus present. Most studies to date have under-sampled the very small particles, which are likely to be amorphic and have a higher nutritional value. These ﬁne particles may also constitute a large proportion of the available organic content of detritus and the EAM. It is therefore likely that the contribution of detritus to the organic content of the EAM has consistently been underestimated. Spatial variation in stocks of detritus within the EAM has recently come under closer investigation. Purcell & Bellwood (2001) demonstrated that the amount of detritus on an exposed reef crest was signiﬁcantly lower than on the reef base or ﬂat. On a broader scale, the detritus content of EAMs from outer and mid-shelf reefs of the GBR suggest a direct relationship between wave energy and detritus load (Crossman et al. 2001). It was estimated that 53% of the variation in the dry weight of detritus within the EAM could be explained by impinging wave energy (Crossman et al. 2001). Consequently, a lower wave energy in sheltered coral reef lagoons may partially explain the relatively high detritus content of EAM habitats in these areas (Table 1). Apart from hydrodynamic forces, there are a variety of biological factors that can inﬂuence the distribution of detritus. Purcell (2000) found inorganic sediment loads in the EAC were directly related to canopy height. Because the areal standing stock of detritus is positively related to the inorganic sediment load (Purcell & Bellwood 2001), algal canopy height can strongly inﬂuence the abundance of detritus in the EAM. Purcell (2000) also found that algal functional groups differed in their ability to capture sediment. Filamentous and small branching algae have high sediment loads, whilst crustose algae have lower sediment loads. Physiological features of algae may also inﬂuence the amount of detritus caught in the EAM. Many algae are known to exude polymer compounds (Clayton & King 1990) and organic detritus in the water column may adhere to this mucus material. Variability in the capability of algae to produce mucus, may therefore inﬂuence the ability of the algae to trap detritus.

Nutritional value of detritus Although detritus may represent a large percentage of the EAM organic matter, the full nutritional status of this resource needs to be established. The nutritional value of detritus can be characterised by its protein and energy content (Bowen 1987). Protein is important because growth rates of ﬁshes can be positively related to protein consumption (Fris & Horn 1993). Energy is also considered an important dietary component, as it is required by all living organisms for metabolism. The relative protein and energy content of detritus can be summarised by the protein : energy ratio, which is often approximated with C : N ratios. Several recent studies on coral reefs have assessed the potential nutritional value of detritus within the EAM. These studies have typically compared the nominal nutritional value of detritus to that of ﬁlamentous algae collected from the same EAM. Filamentous algae typically represent a substantial proportion of the organic matter in the EAM and have traditionally been viewed as the major source of nutrition for ﬁsh feeding on the EAM. Wet chemical analyses of protein in detritus and ﬁlamentous algae indicate that the overall protein content of these two resources is similar, although algae tend to have slightly 283

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Table 2 Nutritional value of detritus and ﬁlamentous algae collected from the same EAM. Figures in parentheses are standard errors. OM Organic matter. Index of nutritional value

Detritus

Filamentous algae

Protein (mg g1) Protein (mg g1 OM) Protein : energy (mg Kj1) C:N C:N C:N C:N Amino acids (mg g1)

1.8 (0.1) 2.1(0.1) 9.5–19.6 16.2–25.0 10–11 8–9 17.2 (0.8) 20.0 (1.8) 6.3 (0.5) 6.8 (0.8) 6.8 (0.8) 9.1 (0.6) 6–8 8–10 20.9 (2.0) 11.6 (1.0)

Location

Reference

Lizard Island Lizard Island Lizard Island Lizard Island Lizard Island Lizard Island Lizard Island Northern GBR

Wilson 2000 Wilson 2002 Wilson 2002 Wilson 2000 Wilson & Bellwood 1997 Wilson & Bellwood 1997 Purcell & Bellwood 2001 Crossman et al. 2001

higher protein levels (Table 2). However, when protein : energy ratios are compared, detritus typically has higher ratios than ﬁlamentous algae. The C : N ratios of detritus are also consistently lower than algae collected from the same EAM, again indicating that detritus is a relatively more nutritious source of dietary nitrogen than algae. Furthermore, protein : energy ratios of coral reef detritus are higher than those found by Bowen (1984) to be suitable for growth of Tilapia mossambicus. This suggests that detritus in the EAM is of a nutritional value which is more than capable of sustaining growth in coral reef ﬁshes. Using total hydrolysable amino acids, Crossman et al. (2001), compared the protein content of detritus and algae in EAM samples. At a number of sites, the concentration of amino acids in detritus samples was typically two times greater than in algae (Table 2). These higher concentrations may be attributed to non-protein amino acids, which make up a substantial proportion of the total amino acid content of detritus in freshwater lakes (Bowen 1980). Non-protein amino acids are readily assimilated by detritivorous Tilapia mossambicus (Bowen 1980), indicating they are a viable and potentially important source of nutrition. Higher concentrations of amino acids in detritus samples from the EAM provide further evidence that detritus has a nutritional value equal to or better than that of ﬁlamentous algae. The organic content of detrital aggregates can also be used as an index of energetic content available to detritivores and as a measure of nutritional value (Bowen 1987). Purcell & Bellwood (2001) found that sediments in EAM collected on the crest of a windward reef had a higher organic content than sediment from the reef base or ﬂat. The total amount of detritus in the EAM was low on the crest relative to the other reef zones but C : N ratios did not vary signiﬁcantly. Consequently, although the amount of organic detritus present on the crest is relatively small, a low percentage of inorganic sediments indicate that the detritus in this zone is of a relatively higher nutritional value. This may, in part, explain why detritivorous ﬁshes congregate on windward-facing crests (Wilson 2001). Microbes are often perceived to be the main source of dietary nutrients in detritus (Newell 1965) and microbial biomass has been used to assess the potential nutritional value of detritus (Bowen 1987). Bacteria are an important component of detritus on coral reefs, in terms of both nutrition for ﬁshes (Choat & Clements 1998) and decomposition of detritus (Ducklow 1990). Wilson et al. (2001a) estimated that bacteria accounted for 4–10% of the organic matter in detritus collected from the EAM in blenny territories. Other estimates of bacterial contributions to organic matter in coral reef sediments have ranged between 2% and 8% (Sorokin 1974, Moriarty 1982). These estimates are greater than those from other aquatic habitats, where bacteria typically accounts for 1% of the organic detritus (see 284

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Bowen 1987). The relatively large contribution of bacteria to coral reef detritus suggests bacteria may have a more important role in determining the nutritional quality of coral reef detritus than in other aquatic ecosystems. Another group of microbes that contribute to detrital nutrients are the microalgae. Microalgal productivity and biomass in shallow tropical waters (0–5 m) are higher than in other coastal ecosystems (Cahoon 1999). Diatoms are one of the most abundant microalgae encountered in detritus. Based on concentrations of the fatty acid 20 : 53, Wilson et al. (2001a) and visual analyses (Wilson 2002), it is estimated that diatoms accounted for 1–14% of the organic matter in detrital aggregates in the EAM. Diatoms are also a potentially important source of essential dietary compounds such as fatty acids (Volkman et al. 1989, Viso & Marty 1993) and protein (Lourenco et al. 1998). Therefore, although diatoms and other microbes like bacteria are minor components of the detritus, they are potentially valuable contributors to the pool of detrital nutrients (see Fig. 1 and Table 3).

Figure 1 Sources of detritus in the epilithic algal matrix (EAM) collected from Salarias patzneri territories. Based on data from Wilson et al. 2001a.

285

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Table 3 Potential sources of protein in detrital samples collected from the EAM within Salarias patzneri territories. % contribution of each source calculated from biomarkers (Wilson et al. 2001a)# or visual examination (Wilson 2002)*. Protein input is calculated by multiplying the protein content of each source by the estimated contribution of each source to the detritus. Source

Protein content % dry weight

% of detritus

Protein input (mg g1 detritus)

Reference

Bacteria Blue green algae Copepods Diatoms Diatoms Filamentous algae

63 37–52 55–58 16–18

4–10# 1.9–3.2* 0.2–0.4* 1.5–1.9* 4–18# 72–79#

25–63 7–16 1–2 2–3 6–32 20–104

Simon & Azam 1989 Vargas et al. 1998 Alonzo et al. 2001 Lourenco et al. 1998

7.6–10.4*

2–14 37–217

Filamentous algae TOTAL

2.8–13.2

Montgomery & Gerking 1980

Sources of detritus Hatcher (1983) estimated that 10–80% of net algal production on coral reefs ultimately enters detrital food-webs. Most of the algal productivity on coral reefs can be attributed to turf or epilithic algae (Larkum 1983) and it is these algae that are most likely to contribute to detrital pools (Hansen et al. 1992). The similarity of amino acid and lipid proﬁles of detritus and ﬁlamentous algae, and presence of algal biomarkers in detritus samples lends support to this hypothesis (Crossman et al. 2001, Wilson et al. 2001a). In particular, the fatty acid 18 : 26, which is used as a biomarker for plant and algal-derived detritus (Napolitano 1999) is abundant in detritus collected from the EAM (Wilson et al. 2001a). The percentage of 18 : 26 in detritus and ﬁlamentous algae collected from territories of the blenny Salarias patzneri has been used estimate of the contribution of ﬁlamentous algae to the detritus. Assuming 18 : 26 in detritus is solely from ﬁlamentous algae, it was calculated that ﬁlamentous algae account for 72 9% (SE) and 79 15% (SE) of the organic detritus collected from the EAM in the summer and winter, respectively. These values indicate that ﬁlamentous algae are the major source of detritus in S. patzneri territories but it is clear that a signiﬁcant portion of the detritus comes from other sources. Using a suite of lipid biomarkers, Wilson et al. (2001a) quantitatively estimated the contribution of dinoﬂagellates, diatoms and bacteria to detrital samples collected during the summer and winter (Fig. 1). Regardless of season, microalgae and bacteria make up a substantial proportion of the detritus, particularly in the summer, when they account for approximately 25% of the organic matter. This pattern is consistent with the theory that detritus from senescent algae accumulates in the winter and is metabolised by a more active microbial community in the summer (Wilkinson 1987). This suggestion is supported by the observations of Moriarty et al. (1985), who found higher bacterial biomass and productivity in water collected from the Lizard Island lagoon during the summer. In contrast, Johnstone et al. (1990) found bacterial biomass in sediments from the One Tree Island lagoon to be highest in the winter, although it is unclear if there was any seasonal variation in bacterial production. Inconsistencies between these two studies highlight the differences in microbial 286

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

contributions with season as well as habitat type. For this reason subsequent studies should focus on the contribution of bacteria and microbes to detritus within the EAM where most ﬁshes feed. The nutritional potential of different detritus sources was assessed by comparing estimates of the protein contributions from quantiﬁed detrital sources (Table 3). This comparison suggests that ﬁlamentous algae are the most likely source of protein in detritus. However, there are also substantial contributions from bacteria and diatoms, which have a higher concentration of protein than ﬁlamentous algae. Contributions from microbes will therefore increase protein : energy ratios and amino acids content in detrital samples, improving its nutritional value relative to ﬁlamentous algae (Table 2). Other organisms, such as fungi and corals may also contribute to the total protein content of detritus, as Table 3 does not contain all of the potential sources of protein. In particular, the addition of coral mucus, which is abundant on coral reefs, may further improve the nutritional value of detritus, because coral mucus is a rich source of protein (Ducklow & Mitchell 1979). The composition of detritus is also likely to vary over spatial scales. The local ﬂora and fauna, along with hydrodynamic forces, are two factors that inﬂuence the composition and nutritional value of detritus (Alongi 1989, Hatcher 1983). For example, in the Low Isles, a coastal system in the Northern GBR, lipid biomarkers suggest that higher plants, in particular mangroves, are an important source of detritus (Shaw & Johns 1985). It is believed that some of this detrital material originated from local stands of mangroves, although other markers suggested that a substantial amount of the material was transported to the Low Isles reef from the mainland. It has also been hypothesised that within a coral reef, highly productive areas like the windward crest and ﬂat act as sources of detritus, whereas sheltered areas like the lagoon act as sinks (Hatcher 1983). However, this hypothesis is based primarily on detritus collected from the water column using sediment traps and not from within the EAM. As the composition of algal assemblages is likely to vary between reef zones (Larkum 1983, Klumpp & McKinnon 1992) and given the similarity of lipids found in ﬁlamentous algae and detrital samples collected from EAMs in the lagoon (Wilson et al. 2001a), it is quite possible that a large portion of the reef-based detritus originates within the EAM.

Processes of detrital production Understanding the processes of detritus production in the EAM are crucial to understanding the trophodynamics of coral reefs and yet it remains one of the most challenging aspects of the subject. Although approximately 70% of the EAM detritus comes from ﬁlamentous algae, the mechanisms linking algae and detritus remain elusive. Here, evidence is presented of three processes that may be important to detrital production. The processes do not necessarily operate independently of each other and, for simplicity, they are discussed separately.

287

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Detritus from dissolved organic matter The absence of any obvious cellular or morphological structures in EAM detritus suggests that non-cellular organic matter is a key component of the detritus and that aggregation of dissolved organic matter (DOM) may be an important process in the formation of EAM detritus (Wilson & Bellwood 1997, Wilson 2000). Both living and dead organisms are potential sources of dissolved nutrients on coral reefs and measures of dissolved nutrients in reef waters suggest that natural levels are much higher than has previously been assumed (Szmant 1997). The DOM from these sources may be exported from the reef, or converted to particulate organic matter (POM) via biotic and abiotic processes. Biotically, bacteria use DOM and are an important component of the microbial loop (Azam et al. 1983). Bacterial productivity, however, cannot account for all the DOM produced on coral reefs (Ducklow 1990). Furthermore, the microbial loop is energetically inefﬁcient, because energy must pass through many levels before it is consumed by apex predators (Mann 1988). The microbial loop alone is therefore unable to sustain a high abundance of detritivorous ﬁshes. Abiotically, organic particulates can form from DOM when they are adsorbed on surfaces, such as carbonates (Otsuki & Wetzel 1973), which may be especially important on coral reefs. Bubbling through nutrient rich water will also produce organic particulates (e.g. Alber & Valiela 1994), as can laminar shearing of water (Passow 2000) and it is possible that DOM can spontaneously form particulates in the form of polymer gels (Chin et al. 1998). Most of the research on the production of organic particulates from DOM has concentrated on the DOM released by phytoplankton (e.g. Passow & Wassman 1994, Hong et al. 1997, Passow 2000) and the same principles of detritus production may be applicable to the EAM. There are many potential sources of DOM, although the most likely source on coral reefs is benthic algae (Ducklow 1990), which release DOM equivalent to 5–10% of their net primary production (Schramm et al. 1984). Consequently, the concentration of DOM is likely to be greater within the EAM than surrounding waters, particularly in the boundary layer surrounding ﬁlamentous algae. A large portion of the detritus in the EAM may therefore be produced in situ. Indeed, comparisons of detritus collected from the EAM (Wilson et al. 2001a) and water column (Currie & Johns 1988) in the Lizard Island lagoon suggest the source and composition of detritus in these two media is different. Furthermore, the biochemical properties of detritus collected in sediment traps, which sample the water column and resuspended material, may be quite different to detritus in the EAM (Dommisse 2001), lending credence to the suggestion that a large percentage of the detritus within the EAM is produced in situ.

Detritus from the water column The vast quantities of water which ﬂow over, around and through reefs carry large amounts of living and non-living particulate organic matter. Suspended particulate matter in reef waters comes from a range of sources and is structurally and chemically diverse. Classes of particulate matter in reef waters include plankton (bacteria, phytoplankton, zooplankton), coral mucus (Johannes 1967), ﬁltering structures and discarded houses of gelatinous zooplankton (Hamner et al. 1988) and uncharacterised organic aggregates in a range of sizes (Johannes 1967, Alldredge & Silver 1988). Water motions in the hydrodynamic boundary zone of reefs are often very energetic. A signiﬁcant proportion of the suspended particulate matter within the boundary zone is likely to be torn or re-suspended from the reef surface. This reef-sourced, suspended particulate matter includes inorganic sediment, algal fragments and detritus washed from EAMs. 288

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Suspended detritus is formed by shredding or disaggregation of larger pre-existing organic structures such as algal fragments and discarded appendicularian houses, or through progressive aggregation of organic particles, colloids and polymers in algal exudates (Passow & Wassman 1994, Engel & Schartau 1999), coral mucus (Qasim & Sankaranarayanan 1970, Sorokin 1974, Ducklow & Mitchell 1979) and bacterial bioﬁlms (Biddanda 1985). Bacteria play an important role as nucleation sites for aggregation and as producers of bioﬁlm polymers (Biddanda 1985). Over time, coral mucus and marine snow particles collect and grow populations of bacteria, microalgae and protozoa (Sorokin 1974, Ducklow & Mitchell 1979, Alldredge & Silver 1988) that enrich and diversify the biochemical composition of the particles. However, regardless of source, most of the particulate organic matter in reef waters is non-living. Concentrations of carbon, nitrogen and phosphorus in suspended particulate matter are usually poorly correlated with chlorophyll concentrations. Estimates of microalgal biomass in particles derived from chlorophyll concentrations indicate that phytoplankton and living pieces of benthic microalgae constitute a relatively minor portion of the suspended particulate matter near reefs (e.g. Charpy 1985). Detritus from pelagic sources can land directly onto the reef surface where it may be trapped by the EAM, or it may arrive as faeces after being collected and eaten by benthic ﬁlter feeders, zooplankton (Gerber & Marshall 1974, Gottfried & Roman 1983) or mobile planktivores (Johannes 1967). Aggregations of planktivorous and particle-feeding ﬁshes congregate along the front (upcurrent) side of reefs to collect zooplankton and other particulate matter carried into the reef by currents (“the wall of mouths”; Hamner et al. 1988). These ﬁshes feed on both plankton and organic aggregates (Johannes 1967). Fast-sinking faecal matter from these ﬁshes delivers undigested remnants of plankton and detrital matter to the benthos. Deposition ﬂuxes of suspended organic matter in reef habitats vary with the local degree of wave energy, which resuspends loose material, and to a lesser extent on the amount and type of suspended particles in the surrounding water column (Dommisse 2001). Sedimentation ﬂuxes of organic matter can vary several-fold spatially and temporally (Koop & Larkum 1987, Clavier et al. 1995) but are generally correlated with wind speed, the principal driver of wave activity (Dommisse 2001). Concentrations of suspended organic aggregates and particles are higher in productive coastal (Fabricius et al., in press) and offshore waters (M. J. Furnas, pers. obs.) than in lower energy, low productivity locations. Particulate sedimentation ﬂuxes on reef front habitats in the GBR and other locations are inversely related to cross-shelf location and water depth (Dommisse 2001). Most of the difference is due to suspended inorganic material. Detritus collected at nearshore reefs is often heavily diluted with resuspended terrigenous sediment. The smaller amounts of material collected by sediment traps at clearer offshore sites had a higher carbon, protein and carbohydrate content than sediment trapped at nearshore sites. Sedimentation ﬂuxes of labile organic material (protein, carbohydrates) at the offshore sites were similar to ﬂuxes at nearshore sites. Overall, coral reef waters tend to be enriched in particulate matter compared with surrounding oceanic and open shelf waters (Marshall 1965, Charpy 1985). This particulate matter is produced locally within reefs (Johannes 1967) or imported from surrounding waters. In shelf waters of the Great Barrier Reef, suspended organic matter may be readily transported between reefs by strong currents. On individual reefs or sections of reefs, this largely non-living organic matter may be trapped on reef surfaces by sedimentation, or pass through one or more trophic levels as it is consumed by detritus and zooplankton feeding ﬁshes and invertebrates. 289

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

The quantitative contribution of pelagic sourced detritus to reef systems is difﬁcult to estimate because it varies with local reef hydrodynamics, detritus levels in surrounding waters and day-to-day ﬂuctuations in wave energy. Quantitatively, pelagic detritus is likely less important to overall reef food webs than locally produced algal material. However, a number of species or guilds of species take advantage of, or depend on, suspended particulate matter. For this reason, suspended detritus is an important contributor to reef community structure.

Detritus from ﬁsh faeces A potentially important source of detritus in the EAM on coral reefs is the deposition of faecal material from ﬁshes. Fishes at all trophic levels excrete organic material on a regular basis and although other ﬁshes ingest a substantial amount of the faeces via coprophagy (Robertson 1982), some of the faecal material ultimately settles on the reef substratum. Faeces from truly herbivorous ﬁshes will be especially important in this process. Species such as Acanthurus lineatus, are abundant on reefs, predominantly ingest algae (Choat et al. 2002) and have high gut through-put rates (Polunin & Klumpp 1992a). At Davies Reef (Central GBR) grazing of epilithic algae by ﬁshes and invertebrates is estimated to account for 42% to 72% of net algae production on the reef ﬂat (Klumpp & Polunin 1990) and 43% to 65% of net production on the windward reef crest (Polunin & Klumpp 1992b). However, assimilation of algae by ﬁshes is often inefﬁcient (Horn 1989) and much of the ingested material is excreted as faeces. Faecal material from herbivorous species is less likely to be consumed (Roberston 1982) and may represent a substantial input to detritus pools in the EAM. The faeces from herbivores are likely to be of low nutritional value (Baily & Robertson 1982) although they may be enriched by microbes attracted to the mucus sheath surrounding many faeces. Therefore, at reef sites, where ﬁshes display a high degree of ﬁdelity (Meyer et al. 1983, Marnane 2000), repeatedly defaecate (Bellwood 1995a), or maintain territories (Klumpp & Polunin 1989) faecal material may represent a substantial and reliable source of detritus.

Detrital consumption: What is a detritivorous ﬁsh? A detritivorous ﬁsh can be loosely deﬁned as one that predominantly ingests detritus (Gerking 1994). Unfortunately, many of the earlier analyses of ﬁsh stomach contents did not consider detritus as a distinct dietary category, often combining detritus with sediment or algae (see Table 4). Furthermore, whereas some studies identiﬁed the number of specimens from each species that had ingested detritus, they did not quantify the amount or volume of detritus within the gut (e.g. Hiatt & Strasburg 1960, Kuo & Shao 1991). These types of analyses make it difﬁcult to assess the contribution of detritus to the diet, and to differentiate between detritivorous and herbivorous ﬁshes. As a consequence, the value of ingested detritus in the diet of ﬁshes was often overlooked in early studies and it was assumed that few species of coral reef ﬁshes were detritivores (Hiatt & Strasburg 1960). The recent use of techniques that quantify the amount of detritus relative to other items in the alimentary tract has, however, provided some enlightening results (Table 5). Although 290

Acanthurus bahianus Acanthurus chirurgus Acanthurus coeruleus Acanthurus nigricauda Acanthurus olivaceus Acanthurus tennenti Ctenochaetus striatus Ctenochaetus striatus Ctenochaetus strigosus Atrosalarias fuscus Blennius cristatus Blennius marmoreus Entomacrodus nigricans Ophioblennius atlanticus Petroscirtes breviceps Petroscirtes mitratus Salarias fasciatus Salarias luctuosus Coryphopterus glaucofraenum Gnatholepis thompsoni Centropyge argi Centropyge ferrugatus Microspathodon chrysurus Pomacentrus fuscus Pomacentrus variabilis Scarus altipinnis Scarus ghobban Scarus sordidus Scarus schlegeli

Acanthuridae

291

Scaridae

Pomacentridae

Pomacanthidae

Gobiidae

Blenniidae

Species Detritus & algae Detritus & algae Detritus & algae Detritus & sediment Detritus & calcareous powder Detritus & sediment Detritus & sediment Detritus & calcareous powder Detritus & sediment Detritus & calcareous powder Detritus & algae Detritus & algae Detritus & algae Detritus & algae Detritus & calcareous powder Detritus & calcareous powder Detritus & calcareous powder Detritus & calcareous powder Detritus & algae Detritus & algae Detritus & algae Detritus & algae Detritus & algae Detritus & algae Detritus & algae Detritus & calcareous powder Detritus & calcareous powder Detritus & calcareous powder Detritus & calcareous powder

Gut contents

Coral reef ﬁshes that ingest detritus and associated algae and/or sediment.

Family

Table 4

92 94 93 98 50 93 99 50 99 69 99 79 96 99 60 60 61 67 50 74 100 70 89 56 51 85 81 82 66

% Total Caribbean Caribbean Caribbean Aldabra Okinawa Is. Aldabra Aldabra Okinawa Is. Aldabra Okinawa Is. Caribbean Caribbean Caribbean Caribbean Okinawa Is. Okinawa Is. Okinawa Is. Okinawa Is. Caribbean Caribbean Caribbean Okinawa Is. Caribbean Caribbean Caribbean Okinawa Is. Okinawa Is. Okinawa Is. Okinawa Is.

Location

Randall 1967 Randall 1967 Randall 1967 Robertson & Gaines 1986 Sano et al. 1984 Robertson & Gaines 1986 Robertson & Gaines 1986 Sano et al. 1984 Robertson & Gaines 1986 Sano et al. 1984 Randall 1967 Randall 1967 Randall 1967 Randall 1967 Sano et al. 1984 Sano et al. 1984 Sano et al. 1984 Sano et al. 1984 Randall 1967 Randall 1967 Randall 1967 Sakai & Kohda 1995 Randall 1967 Randall 1967 Randall 1967 Sano et al. 1984 Sano et al. 1984 Sano et al. 1984 Sano et al. 1984

Reference

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Acanthurus nigricauda Acanthurus olivaceus Ctenochaetus striatus Atrosalarias fuscus Cirripectes chelomatus Ecsenius bicolor Ecsenius mandibularis Ecsenius mandibularis Ecsenius stictus Glyptoparus delicatulus Salarias fasciatus Salarias guttatus Salarias patzneri Amblygobius nocturnis Asterropteryx semipunctatus Asterropteryx semipunctatus Bathygobius fuscus Istigobius decoratus Istigobius goldmanni Valenciennea muralis Dischistodus perspicillatus Hemiglyphidodon plagiometopon Stegastes nigricans Chlorurus microrhinos Chlorurus sordidus Scarus schlegeli

Acanthuridae

292

Scaridae

Pomacentridae

Gobiidae

Blenniidae

Species

38 60 72 62.8 82.9 79.5 74.7 85.5 68 43 80.6 76.3 62.3 73.2 47 71.5 63.3 56.9 45 50.4 35 93.2 55 80 87 87

Detritus in gut % Total 87.7 88.9 88.6 98.9 92.2 93.6 89.5 99.6 95.1 59.7 94.0 97.6 87.5 95.3 55.3 90.7 76.3 66.8 66.6 64.7 78.9 98.2 56.8 91.4 91.6 94.5

% of OM* Northern GBR Northern GBR Northern GBR One Tree Is. Orpheus Is. Lizard Is. Orpheus Is. Orpheus Is. Lizard Is. Lizard Is. One Tree Is. Orpheus Is. Lizard Is. Orpheus Is. Okinawa Is. Orpheus Is. Orpheus Is. Orpheus Is. Orpheus Is. Orpheus Is. Lizard Is. Lizard Is. Lizard Is. Northern GBR Northern GBR Northern GBR

Location

Choat et al. 2002 Choat et al. 2002 Choat et al. 2002 Wilson 2000 Wilson 2000 Wilson 2000 Wilson 2000 Depczynski & Bellwood in press Wilson 2000 Wilson 2000 Wilson 2000 Wilson 2000 Wilson 2000 Depczynski & Bellwood in press Sano et al. 1984 Depczynski & Bellwood in press Depczynski & Bellwood in press Depczynski & Bellwood in press Depczynski & Bellwood in press Depczynski & Bellwood in press Wilson & Bellwood 1997 Wilson & Bellwood 1997 Wilson & Bellwood 1997 Choat et al. 2002 Choat et al. 2002 Choat et al. 2002

Reference

Coral reef ﬁshes identiﬁed as detritivores. *% of material in guts when inorganic sediment is excluded. OM organic material.

Family

Table 5

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

limited to a relatively small number of studies, quantitative analyses of detritus in the alimentary tracts of ﬁshes have identiﬁed 24 species, from 5 families, as detritivores. Representatives of the small cryptic blennies and gobies, as well as larger, conspicuous members of the families Pomacentridae, Acanthuridae and Scaridae have all been shown predominantly to ingest detritus. The results of these studies are surprising; members of these families have traditionally been classiﬁed as herbivores and the presence of detritivorous species in these coral reef ﬁsh families suggests detritivory is more prevalent than previously thought. Species from these families also represent some of the smallest ﬁshes that feed on the EAM (Fig. 2) and detritivorous species are at the lower end of the size distribution described by Choat (1991) for “herbivorous” reef ﬁshes. Limitations on ﬁsh body size may reﬂect the nutritional value of detritus relative to ﬁlamentous algae. Small ﬁshes, with high mass speciﬁc metabolic rates require relatively more energy than large ﬁshes, and easily digested amorphic detritus may be more capable of satisfying the energetic requirements of these small ﬁshes than ﬁlamentous algae. However, generalisations regarding the relative body size of herbivores and detritivores must be made cautiously, as a relatively small number of species have been studied. For most of the ﬁshes in Table 5, detritus accounts for over 85% of the organic matter ingested. For these species detritus is undoubtedly the primary source of nutrition and they are correctly classiﬁed as detritivores. In other species, such as Stegastes nigricans and Glyptoparus delicatulus, the amount of detritus and ﬁlamentous algae in the alimentary tract is very similar and nutrients from both resources are likely to contribute to the ﬁshes’ diet.

Figure 2 Mean and range of maximum standard length of detritivorous coral reef ﬁshes. Sizes taken from Myers 1999.

293

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

The nutritional value of detritus relative to ﬁlamentous algae (Table 2) however, suggests even in these cases detritus is the primary source of nutrients for these ﬁshes and that these species should also be classiﬁed as detritivores. Some gobies, however, ingest substantial amounts of crustaceans (Depczynski & Bellwood in press), that are likely to be of a relatively high nutritional value compared with detritus (Bowen et al. 1995). For these species an assessment of copepod and detrital nutrients is required before trophic status can be assigned. Some researchers also argue that even those ﬁsh that ingest 85% or more detritus are still herbivorous, because microalgae within the sediment and detritus are their primary source of nutrition. For example, Robertson (1984) found the prominent items in the guts of three species of Caribbean territorial pomacentrids were diatoms. In contrast, examination of gut contents at 400 magniﬁcation from two scarid, two acanthurid and one blenny species, collected from the GBR have determined that diatoms account for a very small percentage of ingested material (Choat et al. 2002, Wilson 2002). Microalgae represented 1% of the gut contents in scarids and blennies and 10% of acanthurid gut contents. Fishes feeding on the EAM will undoubtedly ingest diatoms and other microalgae but, when compared with the amount of detritus ingested, microalgae typically represent a small percentage of gut contents. It is therefore important that future studies not only identify micro-organisms in ﬁsh guts, but also quantify their contribution relative to other items, particularly detritus. Examination of ingested material under high magniﬁcation will help to identify many of the smaller components of detritus. However, because many of the ﬁshes that feed on the EAM triturate ingested material, ingested items may be unidentiﬁable. Furthermore, it is possible that detritus in the gut is an artifact of these digestive processes. In such cases biochemical analysis of ingested material is a useful method of assessing diet. An examination of short chain fatty acids (SCFA) in the guts of nominal herbivorous ﬁshes found that those species that grazed on detritus and sediment had high isovalerate and low acetate content relative to other SCFA in their guts (Choat & Clements 1998, Choat et al. 2002). For ﬁshes feeding on the EAM, high amounts of isovalerate in the gut is believed to originate principally from the digestion of bacteria, which contain relatively high concentrations of leucine, the precursor of isovalerate. As bacteria and other microbes are an important component of detritus on coral reefs the presence of isovalerate and amorphic organic matter in the guts of ﬁshes provides strong evidence of detritivory by these species. Combined with biochemical analyses that use stable isotopes and lipid biomarkers, SCFA offer a means of investigating the true nature of ambiguous material in the diet. One of the major limitations of gut content analyses is that, at best, it can only identify what the ﬁsh ingests, not whether the item is digested and ultimately assimilated. Wilson et al. (2001b) examined assimilation of detritus by a coral reef blenny by comparing the fatty acid composition of Salarias patzneri tissues with detritus and algae collected from ﬁsh territories. They found the relative concentration of two dietary biomakers in S. patzneri tissues and detritus was very similar, yet signiﬁcantly different to samples of ﬁlamentous algae collected from the same territory. This ﬁnding provides direct evidence that ingested detritus is assimilated by this species and implies that many of the other ﬁshes that ingest detritus are capable of digesting and assimilating nutrients from detritus. Studies on assimilation of nutrients by three species of scarids that ingest predominantly detritus have also demonstrated that these ﬁshes assimilate over 90% of the protein and lipid and over 50% of the carbohydrate present in the anterior part of their intestine (D. J. Crossman, pers. comm.). As the majority of material ingested by these species is detritus 294

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

(Choat et al. 2002), high levels of assimilation indicate that macronutrients must be assimilated from detritus. Similarly, Nelson & Wilkins (1988) estimated assimilation efﬁciencies of 37% for nitrogen and 20% for total organic matter for the surgeonﬁsh, Ctenochaetus striatus, a species that primarily ingests detritus and inorganic sediment (Choat et al. 2002, Robertson & Gaines 1986). Therefore, as with scarids, any organic matter or nitrogen assimilated by these ﬁshes must originate from detrital particles. The high percentage of detritus ingested by ﬁshes (Table 5), the high nutritional quality of detritus relative to ﬁlamentous algae, and strong evidence of detrital assimilation indicate that these ﬁshes can unequivocally be classiﬁed as detritivores. The list of detritivorous ﬁshes is, however, far from comprehensive, because there are many other EAM feeding species whose gut contents have yet to be accurately quantiﬁed. Further research on the amount of detritus ingested and assimilation of detritus by ﬁshes is therefore required. As the diet of these ﬁshes may vary geographically, future research should also examine spatial variations in detritivory within species, as well as geographic variation in the importance of detritivory within ﬁsh assemblages. With this approach we will gain a better appreciation of the importance of detritus in the diet of individual ﬁsh species and in coral reef trophodynamics as a whole.

Feeding traits of detritivorous ﬁshes The high proportion of detritus in the guts of many detrivorous ﬁshes relative to the amount of detritus in the EAM suggest that many species selectively feed on detritus. Comparisons of blenny and pomacentrid gut contents with samples of the EAM from their territories show that these ﬁshes ingest a higher percentage of detritus and lower percentage of ﬁlamentous algae than is present in the EAM (Wilson & Bellwood 1997, Wilson 2000). The selection of detritus over ﬁlamentous algae by ﬁshes may be related to the structure of the jaws and associated elements, or feeding behaviour. The acanthurid, Ctenochaetus striatus and blennies from the tribe Salariini, have numerous soft or ﬂexible teeth that they use to brush loose detritus and sediment from the EAM, leaving most of the ﬁlamentous algae still attached to the substratum (Purcell & Bellwood 1993, Wilson 2000). Alternatively, some detritivorous ﬁshes may suck loose particles from surfaces (Gerking 1994), a technique that would favour ingestion of detritus rather than attached algae. The selection of detritus over ﬁlamentous algae by these species also infers that it is advantageous to target detrital resources within the EAM, possibly relating to the nutritional value of detritus compared with algae, or partitioning of resources within the EAM by different species (Choat & Bellwood 1985). An inherent problem of many detritus-based diets is that high amounts of inorganic sediment and indigestible material are often ingested. A low percentage of digestible energy and variation in the nutritional quality of detritus dictate that ﬁshes targeting this resource must ingest large amounts of material and practice selective feeding (Bowen et al. 1995). On coral reefs the percentage of organic detritus in particulates collected from within the EAM is generally 20% (Table 1, p. 282), inferring that digestible energy of these particulates is likely to be low. There is a large degree of variation in the consumption rates of coral reef ﬁshes that feed on detritus within the EAM, rates ranging from 2% to 170% of wet body weight ingested as dry material daily (Table 6). Polunin et al. (1995) suggested that variation in the ingestion rates of these ﬁshes may be related to the amount of inorganic sediment in 295

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Table 6 Ingestion rates of detritivorous coral reef ﬁshes. Wet weight of Chlorurus microrhinos and C. sordidus estimated from length weight relationships (Choat et al. 1996) for ﬁsh of 350–440 mm and 150–200 mm standard length respectively. Species

Ingestion rate mg (dw) g (ww)1 day1

Reference

Ctenochaetus striatus Atrosalarias fuscus Stegastes nigricans Chlorurus microrhinos Chlorurus sordidus Chlorurus sordidus

223 22.4 88 881–1692 236 200–496

Polunin et al. 1995 Klumpp & Polunin 1989 Polunin et al. 1995 Bellwood 1995b Polunin et al. 1995 Bellwood 1995b

their diet. Species with a high percentage of inorganic sediment in their diet, such as Chlorurus sordidus and Ctenochaetus striatus, have the highest ingestion rates, whilst Stegastes nigricans, which feeds on particulates with a high percentage of organic detritus (Wilson & Bellwood 1997), has a relatively low ingestion rate. Phylogenetic differences, basal metabolic rate and protein levels of diet may also inﬂuence ingestion rates of these ﬁshes but these results do support the hypothesis that high ingestion rates of detritivores are related to low organic and energy content of the diet (cf. Harmelin-Vivien 2002). To overcome the low energetic value of detritus, it is possible that many detritivorous ﬁsh selectively feed on organically rich particles. Three species of scarid and the pomacentrid, Dischistodus perspicillatus, all have a higher percentage of organic matter in their anterior intestine or stomach than is present in the EAM on which they feed (Wilson & Bellwood 1997, Crossman, pers. comm.). Higher organic content within the gut may be related to selection of detrital material within the EAM. Pomacentrids and blennies show a propensity for small particles in the EAM that have a higher percentage of organic matter (Wilson & Bellwood 1997, Wilson 2000). Thus, preferential ingestion of these particles will increase the percentage of organic matter in their diet. Detritivorous ﬁshes may also feed in locations on the reef where the amount of detritus relative to inorganic sediment in the EAM is high (Purcell & Bellwood 2001, Wilson 2001). Selective feeding by detritivorous ﬁshes on small particles within the EAM and in sediments with a high detrital content will result in proportionally higher amounts of organic matter and energy in the diet. Selective feeding at different locations may also be an important means of obtaining sufﬁcient dietary protein for detritivorous ﬁshes (Bowen et al. 1995). In a South African lake Bowen (1979) proposed that the nearshore distribution of detritivorous ﬁshes was related to the relatively high concentration of protein-rich detrital aggregates in shallow water terraces. On coral reefs, Crossman et al. (2001) found that the amino acid concentration of detritus did not vary between mid- and outer-shelf locations of the GBR, whereas Purcell & Bellwood (2001) showed that the C : N ratio of detrital samples was similar across different zones of a windward reef. Furthermore, although pomacentrids ingested a relatively higher proportion of particles 125 m than was present in the EAM, the C : N ratio of detritus in these particles was similar to larger sized organic particulates (Wilson & Bellwood 1997). Uniformity in amino acid concentrations and C : N ratios in detrital samples collected from the EAM therefore provide little evidence to suggest that detritivorous coral reef ﬁshes selectively feed on detritus with a higher protein content. Relatively high concentrations of 296

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Table 7

Maximum feeding rates of detritivorous ﬁshes.

Species

Bites min1 (max)

Time of max bite rate

Reference

Ctenochaetus striatus Salarias fasciatus Salarias patzneri Stegastes nigricans Chlorurus microrhinos Chlorurus sordidus Chlorurus sordidus Chlorurus sordidus Scarus schlegeli

30 8–12 8–12 5.5 8 27 20 19 25–29

1700 1400–1600 1200–1500 1300 1100–1600 1400 1100–1600 1030–1530 1030–1730

Polunin et al. 1995 Klumpp & Polunin 1990 Wilson unpubl. Polunin et al. 1995 Bellwood 1995b Polunin et al. 1995 Bellwood 1995b Choat & Clements 1993 Choat & Clements 1993

amino acids in detritus compared with ﬁlamentous algae (Crossman et al. 2001) may therefore remove the need to feed selectively on protein-concentrated detritus and selection of organic-rich sediments, which contain easily assimilated amorphic detritus and little inorganic sediment, may take precedence. Diurnal changes in the feeding intensity of detritivorous ﬁshes may also be a means of selective feeding. Feeding rates of detritivorous ﬁshes are generally low in the early morning, then gradually increase, peaking in the late morning or mid afternoon, before declining in the late afternoon (Table 7). Similar patterns are typical of many ﬁshes that are believed to feed on algae and it has been proposed that this pattern relates to diurnal changes in the amount of algal photosynthate (Polunin & Klumpp 1989, Horn et al. 1990, Zoufal & Taborsky 1991, Bruggemann et al. 1994). Polunin & Klumpp (1989) offered the same explanation for the feeding pattern of Ctenochaetus striatus, the assumption being that microalgae within the detritus were an important component of the diet. Accumulation of photosynthetic products from microalgae may also be an important determinant of feeding patterns for detritivorous ﬁshes, although there is also some evidence of increased bacterial production (Moriarty et al. 1985) and an increase in dissolved organic carbon concentration in the mid afternoon (Schramm et al. 1984). The inﬂuence of these variables on the feeding rates of detritivorous ﬁshes is yet to be assessed. However, increases in bacterial productivity and higher concentrations of DOM are also possible explanations for increased feeding rates of detritivores in the mid afternoon.

Contribution of detritivorous ﬁshes to coral reef ﬁsh assemblages Fishes identiﬁed as detritivores in Table 5 represent some of the most abundant species on coral reefs around the world. Members of the families Scaridae and Acanthuridae, which are often the most conspicuous EAM-feeding coral reef ﬁshes (Steneck 1988), have been classiﬁed as detritivores. Visual surveys of ﬁshes from these families in the Seychelles (Jennings et al. 1996), Kenya (McClanahan 1994) and on the GBR (Russ 1984a,b, Choat & Bellwood 1985) all identify Chlorurus sordidus and Ctenochaetus striatus as one of the most abundant species from the families Scaridae and Acanthuridae respectively. It is therefore not surprising that in a recent survey of scarids, acanthurids and siganids in the 297

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Maldives (Sluka & Miller 2001) at least 50% of all the ﬁshes identiﬁed can be classiﬁed as detritivores. Unfortunately, quantitative evaluation of detritus in the guts of reef ﬁshes has been restricted to species collected from the Indo–Paciﬁc, making it difﬁcult to assess the proportion of detritivores in coral reef ﬁsh assemblages in other regions such as the Caribbean. Many of the ﬁsh species in regions like the Caribbean ingest detritus (Randall 1967), although the amount of detritus ingested by these species has not been quantiﬁed. Nonetheless, it is likely that a signiﬁcant proportion of species will be classiﬁed as detritivores. Another problem with many of the surveys conducted on coral reef ﬁshes that feed on the EAM is that they tend to focus on the large conspicuous species and disregard smaller cryptic ﬁshes. Yet in some locations, small ﬁshes from the families Pomacentridae and Blenniidae are responsible for removing more of the EAM than the rest of the ﬁsh community combined (Hatcher 1981). The salariine blennies are of particular interest, as gut content analysis of many species collected from coral reefs indicate they are detritivores (Table 5). Density estimates of blennies also suggest that they represent 20% to 50% of the individuals that feed on the EAM (Townsend & Tibbetts 2000, Wilson 2001). Small body size means blennies only represent 5% to 9% of biomass estimates of the same detritivorous/herbivorous ﬁsh assemblages (Townsend & Tibbetts 2000, Wilson 2001). However, as these ﬁshes are likely to have a higher mass speciﬁc metabolic rate than larger ﬁshes (Clarke & Johnstone 1999), they are likely to have greater trophodynamic signiﬁcance than is indicated by relative biomass estimates. Furthermore, as longevity estimates for salariine blennies are only 1 yr to 4 yr (Labelle & Nursall 1992, S. K. Wilson unpubl. data), they are likely to have much higher turnover rates than acanthurids that live for up to 30 yr to 45 yr (Choat & Axe 1996), or scarids that have maximum life spans of 5 yr to 20 yr (Choat et al. 1996). Similarly, juveniles of EAM feeding ﬁsh are highly abundant on reefs, have high mortality rates and high mass speciﬁc metabolic rates, yet are often excluded from censuses. Many of these juveniles are likely to be detritivores and if combined with small detritivorous ﬁshes like blennies, may contribute substantially to coral reef trophodynamics. To evaluate the relative importance of small versus large species and herbivory versus detritivory amongst ﬁshes that feed on the EAM, the density estimates from three independent studies were compared. These studies used visual censuses to assess ﬁsh abundances and were carried out at similar sites around Lizard Island. Although limited by differences in censusing technique and timing, the combination of these three studies provides a comprehensive list of EAM feeding ﬁsh from a single location and, combined with data in Table 5, offers a unique opportunity to examine the trophic status of ﬁshes within this assemblage. Of the ﬁsh that feed on the EAM at Lizard Island, the pomacentrids were the most abundant family, followed by relatively high numbers of blennies and acanthurids (Fig. 3). Approximately 50% of the pomacentrids and acanthurids can be assigned the trophic status of detritivores, whilst all of the blennies can be similarly classiﬁed. Scarids are also relatively abundant on the reef, although a large percentage of scarids were not assigned a trophic group. These ﬁshes, however, typically ingest the entire EAM (Bellwood & Choat 1990), and because detritus is often the most abundant and nutritionally valuable resource within the EAM, it is likely that most scarids are detritivores. Biomass estimates of the EAM-feeding ﬁshes on the reefs surrounding Lizard Island indicated that scarids and acanthurids had the greatest biomass (Fig. 4). Approximately 50% of the biomass in both these families can be attributed to detritivorous ﬁshes. When estimates from all families are combined detritivorous ﬁsh densities were from 298

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Figure 3 Density estimates of detritivorous and other EAM-feeding ﬁshes on Lizard Island reefs. Blenny density after Wilson (2001), pomacentrids after Meekan et al. (1995), scarids, acanthurids, siganids and kyphosids, J. H. Choat (unpubl. data).

0.32 ﬁsh m2 to 0.10 ﬁsh m2, with highest densities on the reef crest (Fig. 5). When compared with the density of other EAM feeding ﬁshes, detritivores accounted for approximately 20% of all the ﬁsh censused. However, these estimates are highly conservative because the trophic status of approximately 70% of the ﬁshes is unresolved. Consequently, the percentage of detritivorous ﬁshes in this assemblage is likely to be considerably higher than 20%, adding to the undeniably signiﬁcant contribution of detritivorous ﬁshes to this 299

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Figure 4 Biomass estimates of detritivorous and other EAM-feeding ﬁshes on Lizard Island reefs. Biomass estimates of ﬁshes calculated from densities in Figure 3, and published length weight relationships (Kulbicki et al. 1993, Wilson 2001). Total length of individual ﬁsh was estimated during visual census, with the exception of the pomacentrids. A mean total length for each pomacentrid species was estimated by multiplying the maximum total length in Randall et al. (1990) by 0.63. This multiplication factor was based on the mean total length of ﬁshes from other families at Lizard Island, relative to their maximum total length.

300

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Figure 5 Total density and estimated biomass of detritivorous and other EAM-feeding ﬁshes on Lizard Island reefs. Based on data from Figures 3 and 4. Note the proportion of species for which the dietary status is yet to be resolved, and the clear preponderance of detritivores on the crest.

food web. Combining the biomass estimates from all families, the total detritivore biomass was estimated to be between 3.1 g m2 and 16.2 g m2, with greatest biomass on the reef crest (Fig. 5). Overall, detritivores accounted for approximately 40% of the total biomass of EAM feeding ﬁshes, although, as with density estimates, detritivore biomass must be considered conservative and is likely to be much greater than 40%. A high abundance and biomass of detritivorous ﬁshes on the crest may be related to the nutritional quality and productivity of detritus in this reef zone. Purcell & Bellwood (2001) demonstrated that although detrital stocks in the EAM are low on the crest, this detritus is of 301

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

a relatively high nutritional value. The low standing stock of detritus and high abundance of detritivores on the crest also infers that productivity of detritus on the crest is high. Filamentous algae, the primary source of detritus, typically have high productivity on the crest (Klumpp & McKinnon 1989). High water turbity on the crest produces bubbles and laminar shear, which may promote the rate of DOM aggregation and production of amorphic detritus. Deposition rates of detritus are also high on the crest, although it is unclear how much of this material is re-suspended from the substrata (Koop & Larkum 1987). Nonetheless, detritus production is likely to be higher on the reef crest and because this detritus of a relatively high nutritional value, it is not surprising that there is a high abundance and biomass of detritivores on the crest. The consumption of detritus by coral reef ﬁshes also extends beyond those species that feed on the EAM. Detritus can be found amongst the loose sediments between reefs (Johnstone et al. 1990) and some of the ﬁshes that target interstitial resources within these sediments may ingest detritus. Small gobies are often found in such habitats, and gut content analyses have revealed that a large proportion of species ingest signiﬁcant quantities of detritus (Depczynski & Bellwood in press). The coloration patterns and behaviour of many gobies mean they are often overlooked in visual surveys, even though they represent a substantial proportion of cryptobenthic ﬁsh assemblages (Ackerman & Bellwood 2000). Like blennies, small body size and high relative metabolic rate suggest that detritivorous gobies could play an important role in detrital ﬂuxes on coral reefs. Further research on the diet, relative abundance and longevity of these species is required to quantify their contribution to trophodynamic models. Detrital particles suspended in the water column are another dietary resource that may be utilised by some coral reef ﬁshes. Detritus is often abundant in the water column above coral reefs (Marshall 1965, Johannes et al. 1970, Glynn 1973, Gerber & Marshall 1982) and has been found in the guts of ﬁshes that typically feed on plankton (Gerber & Marshall 1974). As many planktivorous ﬁshes occasionally take bites from the substratum, it is possible that some of the detritus in the gut comes from benthic feeding forays, rather than the water column. However, regardless of the source, detritus is ingested by some of these ﬁshes, and, in conjunction with detritivores feeding on the EAM and loose sediments, it is clear that ﬁshes have an important role in transferring detrital-based dietary resources to higher trophic levels on coral reefs.

Future directions in research To appreciate the importance of detritus and detritivorous ﬁsh to coral reef trophodynamics future research should focus on the detritus within the EAM. The EAM is likely to be a major reservoir of detritus and is also where many reef ﬁsh feed. Detritus in the EAM is therefore likely to play a major role in coral reef trophodynamics. To evaluate detrital ﬂuxes on coral reefs it is imperative that rates of EAM detritus production and consumption are quantiﬁed. The primary modes of detrital production in the EAM need to be identiﬁed and their relative signiﬁcance evaluated. Understanding the processes that are involved in the production of detritus and identifying sources of detritus in the EAM will help determine why detritus in the EAM is of a high nutritional value relative to ﬁlamentous algae and identify primary channels for the ﬂux of 302

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

material in reef trophodynamics. Microbes undoubtedly add essential nutrients to the detritus, whilst production of detritus via DOM may result in a reduction of refractory compounds and secondary metabolites that are often associated with algae and can impede feeding and ingestion by consumers. The challenge for future researchers is to obtain reliable estimates of detrital accumulation in the EAM and evaluate the relative importance of different processes of detrital production over a range of spatial and temporal scales. Future research should also work towards compiling a comprehensive list of detritivorous ﬁshes. Future dietary studies need to examine the contribution of detritus relative to other dietary items. Small cryptic species and juveniles, which are likely to be major contributors to reef trophodynamics, require particular investigation. Furthermore, most of the studies to date have been conducted on the GBR, consequently detritivory must be examined on a broader spatial scale, both at the species and community levels.

Acknowledgements This paper was improved as a result of comments and discussions with R. Fisher, M. Depczynski and M. Dommisse. In was supported by a JCU Doctoral Merit Research Scholarship (SKW) and the Australian Research Council (DRB).

References Ackerman, J. & Bellwood, D. R. 2000. Reef ﬁsh assemblages: a re-evaluation using rotenone stations. Marine Ecology Progress Series 206, 227–237. Alber, M. & Valiela, I. 1994. Biochemical composition of organic aggregates produced from the marine macrophyte-derived dissolved organic matter. Limnology and Oceanography 39, 717–723. Alldredge, A. & Silver, M. W. 1988. Characteristics, dynamics and signiﬁcance of marine snow. Progress in Oceanography 20, 41–82. Alongi, D. M. 1989. Detritus in coral reef ecosystems: ﬂuxes and fates. Proceedings of the 6th International Coral Reef Symposium. 1, 29–36. Alonzo, F., Mayzaud, P. & Razouls, S. 2001. Egg production and energy storage in relation to feeding conditions in the subantartic calanoid copepod Drepanopus pectinatus: an experimental study of reproductive strategy. Marine Ecology Progress Series 209, 231–242. Arias-Gonzalez, J. E., Delesalle, B., Salvat, B. & Galzin, R. 1997. Trophic functioning of the Tiahura reef sector, Moorea Island, French Polynesia. Coral Reefs 16, 231–246. Azam, F., Fenchel, T., Field, J. G., Gray, J. S., Meyer-Reil, L. A. & Thingstad, F. 1983. The ecological role of water column microbes in the sea. Marine Ecology Progress Series 10, 257–263. Bailey, T. G. & Robertson, D. R. 1982. Organic and caloric levels of ﬁsh feces relative to its consumption by coprophagus reef ﬁsh. Marine Biology 69, 45–50. Begon, M., Harper, J. L. & Townsend, C. R. 1990 Ecology. Individuals, populations and communities. Cambridge, Mass.: Blackwell. Bellwood, D. R. 1995a. Carbonate transport and within-reef patterns of bioerosion and sediment release by parrotﬁshes (family Scaridae) on the Great Barrier Reef. Marine Ecology Progress Series 117, 127–136. Bellwood, D. R. 1995b. Direct estimate of bioerosion by two parrotﬁsh species, Chlorurus gibbus and C. sordidus, on the Great Barrier Reef, Australia. Marine Biology 121, 419–429.

303

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Bellwood, D. R. & Choat, J. H. 1990. A functional analysis of grazing in parrotﬁshes (family Scaridae): the ecological implications. Environmental Biology of Fishes. 28, 189–214. Biddanda, B. A. 1985 Microbial synthesis of macroparticulate matter. Marine Ecology Progress Series 20, 241–251. Bowen, S. H. 1979. A nutritional constraint in detritivory by ﬁshes: the stunted population of Saratherodon mossambicus in Lake Siaya, South Africa. Ecological Monographs 49, 17–31. Bowen, S. H. 1980. Detrital nonprotein amino acids are the key to rapid growth of Tilapia in Lake Valencia, Venezuela. Science 207, 1216–1218. Bowen, S. H. 1981. Digestion and assimilation of periphytic detrital aggregate by Tilapia mossambica. Transactions of the American Fisheries Society 110, 239–245. Bowen, S. H. 1984. Evidence of a detritus food chain based on consumption of organic precipitates. Bulletin of Marine Science 35, 440–448. Bowen, S. H. 1987. Composition and nutritional value of detritus. In Detritus and microbial ecology in aquaculture. ICLARM conference proceedings, D. S. W. Moriarty & R. S. V. Pullin (eds). Manila, Phillipines: 192–216. Bowen, S. H., Lutz, E. V. & Ahlgren, M. O. 1995. Dietary protein as determinants of food quality: trophic strategies compared. Ecology 76, 899–907. Bruggemann, J. H., Begeman, J., Bosma, E. M., Verburg, P. & Breeman, A. M. 1994. Foraging by the stoplight parrotﬁsh Sparisoma viridae II. Intake and assimilation of food, protein and energy. Marine Ecology Progress Series 106, 57–71. Cahoon, L. B. 1999. The role of benthic microalgae in neritic ecosystems. Oceanography and Marine Biology: an Annual Review 37, 47–86. Charpy, L. 1985. Distribution and composition of particulate organic matter in the lagoon of Tikehau (Tuamotu Archipelago, French Polynesia). Proceedings of the 5th International Coral Reef Congress, Tahiti 3, 353–358. Chin, W., Orellana, M. V. & Verdugo, P. 1998. Spontaneous assembly of marine dissolved organic matter into polymer gels. Nature 391, 568–572. Choat, J. H. 1991. The biology of herbivorous ﬁshes on coral reefs. In The ecology of ﬁshes on coral reefs, P. F. Sale (ed.). London: Academic Press, 120–155. Choat, J. H. & Axe, L. M. 1996. Growth and longevity in acanthurid ﬁshes: an analysis of otolith increments. Marine Ecology Progress Series 134, 15–26. Choat, J. H., Axe, L. M. & Lou D. C. 1996. Growth and longevity in ﬁshes of the family Scaridae. Marine Ecology Progress Series 145, 33–41. Choat, J. H. & Bellwood D. R. 1985. Interactions among herbivorous ﬁshes on a coral reef: inﬂuence of spatial variation. Marine Biology 89, 221–234. Choat, J. H. & Clements, K. D. 1993. Daily feeding rates in herbivorous labroid ﬁshes. Marine Biology 117, 205–211. Choat, J. H. & Clements, K D. 1998. Vertebrate herbivores in marine and terrestrial environments: a nutritional ecology perspective. Annual Review of Ecology and Systematics 29, 375–403. Choat, J. H., Clements, K. D. & Robbins, W. D. 2002. The trophic status of herbivorous ﬁshes on coral reefs I: Dietary analysis. Marine Biology 140, 613–623. Clarke, A. & Johnston, N. M. 1999. Scaling of metabolic rate with body mass and temperature in teleost ﬁsh. Journal of Animal Ecology 68, 893–905. Clavier, J., Chardy, P. & Chevillon, C. 1995. Sedimentation of particulate matter in the south-west lagoon of New Caledonia: spatial and temporal patterns. Estuarine, Coastal and Shelf Science 40, 281–294. Clayton, M. N. & King, R. J. 1990. Biology of marine plants. Melbourne, Australia: Longman Cheshire. Crossman, D. J., Choat, J. H., Clements, K. D., Hardy, T. & McConochie, J. 2001. Detritus as food for grazing ﬁshes on coral reefs. Limnology and Oceanography 46, 1596–1605. Currie, B. R. & Johns, R. B. 1988. Lipids as indicators of the origin of organic matter in ﬁne marine particulate matter. Australian Journal of Marine and Freshwater Research 39, 371–383.

304

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

D’Avanzo, C. D., Alber, M. & Valiela, I. 1991. Nitrogen assimilation from amorphous detritus by two coastal consumers. Estuarine, Coastal and Shelf Science 33, 203–209. D’Avanzo, C. D. & Valiela, I. 1990. Use of detrital foods and assimilation of nitrogen by coastal detritivores. Estuaries 13, 20–24. Depczynski, M. & Bellwood D. R. In press. The role of cryptobenthic reef ﬁshes in coral reef trophodynamics. Marine Ecology Progress Series. Dommisse, M. 2001. The potential nutritional value of detritus depositing onto coral reefs. PhD thesis, Department of Marine Biology, James Cook University, Townsville, Australia. Ducklow, H. W. 1990. The biomass, production and fate of bacteria in coral reefs. In Coral reefs, Z. Dubinsky (ed.). Amsterdam: Elsevier, 265–289. Ducklow, H. W. & Mitchell, R. 1979. Composition of mucus released by coral reef coelenterates. Limnology and Oceanography 24, 706–714. Engel, A. & Schartau, M. 1999 Inﬂuence of transparent exopolymer particles (TEP) on sinking velocities of Nitzschia closterium aggregates. Marine Ecology Progress Series 182, 69–76. Fabricius, K. E., Wild C., Wolanski, E. & Abele, D. In press. Effects of transparent exopolymer particles, marine sediments and muddy marine snow on the survival of hard coral recruits. Estuarine Coastal and Shelf Science. Fris, M. B. & Horn, M. H. 1993. Effects of diets of different protein content on food consumption, gut retention, protein conversion, and growth of Cebidichthys violaceus (Girard), an herbivorous ﬁsh of temperate zone marine waters. Journal of Experimental Marine Biology and Ecology 166, 185–202. Gerber, R. P. & Marshall, N. 1974. Ingestion of detritus by the lagoon pelagic community at Eniwetok Atoll. Limnology and Oceanography 19, 815–824. Gerber, R. P. & Marshall, N. 1982. Characterization of the suspended particulate organic matter and feeding by the marine zooplankton at Enewatok lagoon. Bulletin of Marine Science 32, 290–300. Gerking, S. D. 1994. Feeding ecology of ﬁshes. San Diego: Academic Press. Glynn, P. W. 1973. Ecology of a Caribbean coral reef. The porites reef-ﬂat biotype: Part II. Plankton community with evidence for depletion. Marine Biology 22, 1–21. Gottfried, M. & Roman, M. R. 1983 Ingestion and incorporation of coral-mucus detritus by reef zooplankton. Marine Biology 72, 211–218. Hamner, W. M., Jones, M. S. Carleton, J. H. Hauri I. R. & Williams, D. McB. 1988. Zooplankton, planktivorous ﬁsh, and water currents on a windward reef face: Great Barrier Reef, Australia. Bulletin of Marine Science 42, 459–479. Hansen, J. A., Klumpp, D. W., Alongi, D. M., Dayton, P. K. & Riddle, M. J. 1992. Detrital pathways in a coral reef lagoon. II Detritus deposition, benthic microbial biomass and production. Marine Biology 113, 363–372. Harmelin-Vivien, M. L. 2002. Energetics and ﬁsh diversity on coral reefs. In Coral reef ﬁshes: dynamics and diversity in a complex ecosystem, P. F. Sale (ed.). London: Academic Press, 265–274. Hatcher, B. G. 1981. The interaction between grazing organisms and the epilithic algal community of a coral reef: a quantitative assessment. Proceedings of the 4th International Coral Reef Symposium 2, 515–524. Hatcher, B. G. 1983. The role of detritus in the metabolism and secondary production of coral reef ecosystems. In Proceedings of the Inaugural G.B.R. Conference, J. T. Baker (ed.). Townsville, Australia: James Cook University, 317–325. Hatcher, B. G. 1997. Organic production and decomposition. In Life and death on coral reefs, C. Birkeland (ed.). New York, Chapman & Hall, 140–174. Hiatt, R. W. & Strasburg, D. W. 1960. Ecological relationships of the ﬁsh fauna on coral reefs of the Marshall Islands. Ecological Monographs 30, 65–127. Hong, Y., Smith Jr, W. O. & White, A. 1997. Studies on transparent exopolymer particles (TEP) produced in the Ross Sea (Antartica) and by Phaeocystis antartica (Prymnesiophyceae). Journal of Phycology 33, 368–376.

305

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Horn, M. H. 1989. Biology of marine herbivorous ﬁshes. Oceanography and Marine Biology: an Annual Review 27, 167–272. Horn, M. H., Murray, S. N., Fris, M. B. & Irelan, C. 1990. Diurnal feeding periodicity of an herbivorous blenniid ﬁsh, Parablennius sanguinolentus, in the western Mediterranean. In Trophic relationships in the marine environment, M. Barnes & R. N. Gibson (eds). Proceedings of the 24th European Marine Biology Symposium. Aberdeen: Aberdeen University Press, 170–182. Jennings, S., Boulle, D. P. & Polunin N. V. C. 1996. Habitat correlates of the distribution and biomass of Seychelles reef ﬁshes. Environmental Biology of Fishes 46, 15–25. Johannes, R. E. 1967. Ecology of organic aggregates in the vicinity of a coral reef. Limnology and Oceanography 12, 189–195. Johannes, R. E., Coles, S. L. & Kuenzel, N. T. 1970. The role of zooplankton in the nutrition of some scleractinian corals. Limnology and Oceanography 15, 1579–1586. Johnstone, R. W., Koop, K. & Larkum, A. W. D. 1990. Physical aspects of coral reef lagoon sediments in relation to detritus processing and primary production. Marine Ecology Progress Series 66, 273–283. Klumpp, D. W. & McKinnon, A. D. 1989 Temporal and spatial patterns in primary production of a coral-reef epilithic algal community. Journal of Experimental Marine Biology and Ecology 131, 1–22. Klumpp, D. W. & McKinnon, A. D. 1992. Community structure, biomass and productivity of epilithic algal communities on the Great Barrier Reef: dynamics at different spatial scales. Marine Ecology Progress Series 86, 77–89. Klumpp, D. W. & Polunin N. V. C. 1989. Partitioning among grazers of food resources within damselﬁsh territories on a coral reef. Journal of Experimental Marine Biology and Ecology 125, 145–169. Klumpp, D. W. & Polunin, N. V. C. 1990. Algal production, grazers and habitat partitioning on a coral reef: positive correlation between grazing rate and food availability. In Trophic relationships in the marine environment, M. Barnes & R. N. Gibson (eds). Proceedings of the 24th European Marine Biology Symposium. Aberdeen: Aberdeen University Press, 372–388. Koop, K. & Larkum, A. W. D. 1987. Deposition of organic material in a coral reef lagoon, One Tree Island, Great Barrier Reef. Estuarine, Coastal and Shelf Science 25, 1–9. Kulbicki, M., Mou Tham, G., Thollot, P. & Wantiez, L. 1993. Length-weight relationships of ﬁsh from the lagoon of New Caledonia. NAGA, The ICLARM Quarterly, 26–30. Kuo, S. & Shao, K. 1991. Feeding habits of damselﬁshes (Pomacentridae) from the southern part of Taiwan. Journal of the Fisheries Society of Taiwan 18, 165–176. Labelle, M. & Nursall, J. R. 1992. Population biology of the redlip blenny, Ophioblennius atlanticus macclurei (Sylvester) in Barbados. Bulletin of Marine Science 50, 186–204. Larkum, A. W. D. 1983. The primary productivity of plant communities on coral reefs. In Perspectives on coral reefs, D. J. Banes (ed.). Townsville, Australia: Australian Institute of Marine Science, 221–230. Lourenco, S. O., Barbarino, E., Marquez, U. M. L. & Aidar, E. 1998. Distribution of intracellular nitrogen in marine microalgae: basis for the calculation of speciﬁc nitrogen-to-protein conversion factors. Journal of Phycology 34, 798–811. Mann, K. H. 1988. Production and use of detritus in various freshwater, estuarine and coastal marine ecosystems. Limnology and Oceanography 33, 910–930. Marnane, M. J. 2000. Site ﬁdelity and homing behavior in coral reef cardinalﬁshes. Journal of Fish Biology 57, 1590–1600. Marshall, N. 1965. Detritus over the reef and its potential contribution to adjacent waters of Eniwetok atoll. Ecology 46, 343–344. McClanahan, T. R. 1994. Kenyan coral reef lagoon ﬁsh: effects of ﬁshing, substrate complexity and sea urchins. Coral Reefs 13, 231–241. McClanahan, T. R., McField, M., Huitric, M., Bergman, K., Sala, E., Nyström, M. Nordemar, I.,

306

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Elfwing, T. & Muthiga, N. A. 2001a. Responses of algae, corals and ﬁsh to the reduction of macroalgae in ﬁshed and unﬁshed patch reefs of Glovers Reef Atoll, Belize. Coral Reefs 19, 367–379. McClanahan, T. R., Muthiga, N. A. & Mangi, S. 2001b. Coral and algal changes after the 1998 coral bleaching: interaction with reef management and herbivores on Kenyan reefs. Coral Reefs 19, 380–391. Meekan, M. G., Steven, A. D. L. & Fortin, M. J. 1995. Spatial patterns in the distribution of damselﬁshes on a fringing coral reef. Coral Reefs 14, 151–161. Meyer, J. L., Schultz, E. T. & Helfman, G. S. 1983. Fish schools: an asset to corals. Science 220, 1047–1048. Montgomery, W. L. & Gerking, S. D. 1980. Marine macroalgae as foods for ﬁshes: an evaluation of potential food quality. Environmental Biology of Fishes 5, 143–153. Moriarty, D. J. W. 1982. Feeding of Holothuria atra and Stichopus chloronatus on bacteria, organic carbon and organic nitrogen in sediments of the Great Barrier Reef. Australian Journal of Marine and Freshwater Research 33, 255–263. Moriarty, D. J. W., Pollard, P. C., Hunt, W. G., Moriarty, C. M. & Wassenberg, T. J. 1985. Productivity of bacteria and microalgae and the effect of grazing by holothurians in sediments on a coral reef ﬂat. Marine Biology 85, 293–300. Myers, R. F. 1999. Micronesian reef ﬁshes. Barrigada, territory of Guam, USA: Coral Graphics. Napolitano, G. E. 1999. Fatty acids as chemical and trophic markers in freshwater ecosystems. In Lipids in freshwater ecosystems, M. T. Arts & B. Wainman (eds). New York: Springer, 21–44. Nelson, S. G. & Wilkins, S. De C. 1988. Sediment processing by the surgeonﬁsh Ctenochaetus striatus at Moorea, French Polynesia. Journal of Fish Biology 32, 817–824. Newell, R. 1965. The role of detritus in the nutrition of two marine deposit feeders, the prosobranch Hydrobia ulvae and the bivalve Macoma balthica. Proceedings of the Zoological Society of London 144, 25–45. Otsuki, A. & Wetzel, R. G. 1973. Interaction of yellow organic acids with calcium carbonate in freshwater. Limnology and Oceanography 18, 490–493. Passow, U. 2000. Formation of exopolymer particles, TEP, from dissolved precursor material. Marine Ecology Progress Series 192, 1–11. Passow, U. & Wassman, P. 1994. On the trophic fate of Phaeocystis pouchetii (Hariot): IV. The formation of marine snow by P. pouchetii. Marine Ecology Progress Series 104, 153–161. Polunin, N. V. C. 1996. Trophodynamics of reef ﬁsheries productivity. In Reef ﬁsheries. N. V. C. Polunin & C. M. Roberts (eds). London: Chapman & Hall, 113–135. Polunin, N. V. C., Harmelin-Vivien, M. & Galzin, R. 1995. Contrasts in algal food processing among ﬁve herbivorous coral-reef ﬁsh. Journal of Fish Biology 47, 455–465. Polunin, N. V. C. & Klumpp, D. W. 1989. Ecological correlates of foraging periodicity in herbivorous reef ﬁshes of the Coral Sea. Journal of Experimental Marine Biology and Ecology 126, 1–20. Polunin, N. V. C. & Klumpp, D. W. 1992a. Algal food supply and grazer demand in a very productive coral reef zone. Journal of Experimental Marine Biology and Ecology 164, 1–15. Polunin, N. V. C. & Klumpp, D. W. 1992b. A trophodynamic model of ﬁsh production on a windward reef tract. In Plant–animal interactions in the marine benthos, D. M. Johns, S. J. Hawkins & J. H. Price (eds). Oxford: Clarendon Press, 213–233. Purcell, S. W. 2000. Association of epilithic algae with sediment on a windward reef in the northern Great Barrier Reef, Australia. Bulletin of Marine Science 66, 199–214. Purcell, S. W. & Bellwood D. R. 1993. A functional analysis of food procurement in two surgeonﬁsh species, Acanthurus nigrofuscus and Ctenochaetus striatus (Acanthuridae). Environmental Biology of Fishes 37, 139–159. Purcell, S. W. & Bellwood, D. R. 2001 Spatial patterns of epilithic algal and detrital resources on a windward coral reef. Coral Reefs 20, 117–125. Qasim, S. Q. & Sankaranarayanan, V. N. 1970. Production of particulate matter by the reef on Karavatti Atoll. Limnology and Oceanography 15, 574–578.

307

S. K. WILSON, D. R. BELLWOOD, J. H. CHOAT & M. J. FURNAS

Randall, J. E. 1967. Food habits of reef ﬁshes of the West Indies. Studies in Tropical Oceanography 5, 665–847. Randall, J. E., Allen, G. R. & Steene, R. C. 1990. Fishes of the Great Barrier Reef and Coral Sea. Bathurst, Australia: Crawford House Press. Robertson, D. R. 1982. Fish feces as ﬁsh food on a Paciﬁc coral reef. Marine Ecology Progress Series 7, 253–265. Robertson, D. R. 1984 Cohabitation of competing territorial damselﬁshes on a Caribbean coral reef. Ecology 65, 1121–1135. Robertson, D. R. & Gaines, S. D. 1986. Interference competition structures habitat use in a local assemblage of coral reef surgeonﬁshes. Ecology 67, 1372–1383. Russ, G. R. 1984a. Distribution and abundance of herbivorous grazing ﬁshes in the central Great Barrier Reef. I. Levels of variability across the entire continental shelf. Marine Ecology Progress Series 20, 23–34. Russ, G. R. 1984b. Distribution and abundance of herbivorous grazing ﬁshes in the central Great Barrier Reef. II. Patterns of zonation of mid-shelf and outershelf reefs. Marine Ecology Progress Series 20, 35–44. Sakai, Y. & Kohda, M. 1995. Foraging by mixed species groups involving a small angelﬁsh, Centropyge ferrugattus (Pomacanthidae). Japanese Journal of Ichthyology 41, 429–435. Sano, M., Shimizu, M. & Nose, Y. 1984. Food habits of teleostean reef ﬁshes in Okinawa Island, Japan. The University Museum of the University of Tokyo 25, 1–128. Schramm, W., Gualberto, E. & Orosco, C. 1984. Release of dissolved organic matter from marine tropical reef plants: temperature and desiccation effects. Botanica Marina 28, 71–77. Shaw, P. M. & Johns, R. B. 1985. Comparison of the lipid composition of sediments from three sites in the Venezuela Basin. Marine Geology 68, 205–216. Simon, M. & Azam, F. 1989. Protein content and protein synthesis rates of planktonic marine bacteria. Marine Ecology Progress Series 51, 201–213. Sluka, R. D. & Miller, M. W. 2001. Herbivorous ﬁsh assemblages and herbivory pressure on Laamu Atoll, Republic of Maldives. Coral Reefs 20, 255–262. Sorokin, Y. I. 1974. Bacteria as a component of the coral reef community. Proceedings of the 2nd International Coral Reef Symposium 1, 3–10. Steneck, R. S. 1988. Herbivory on coral reefs: a synthesis. Proceedings of the 6th International Coral Reef Symposium 1, 37–49. Szmant, A. M. 1997. Nutrient effects on coral reefs: a hypothesis on the importance of topographic and trophic complexity to reef nutrient dynamics. Proceedings of the 8th International Coral Reef Symposium 2, 1527–1532. Townsend, K. A. & Tibbetts, I. R. 2000. Biomass and distribution of herbivorous blennies in the southern Great Barrier Reef. Journal of Fish Biology 56, 774–791. Vargas, M. A., Rodriguez, H., Moreno, R. J., Olivares, H., Del Campo, J. A., Rivas, J. & Guerrero, M. G. 1998. Biochemical composition and fatty acid content of ﬁlamentous nitrogen-ﬁxing cyanobacteria. Journal of Phycology 34, 812–817. Viso, A. C. & Marty, J. C. 1993. Fatty acids from 28 marine microalgae. Phytochemistry 34, 1521–1533. Volkman, J. K., Jeffrey, S. W., Nichols, P. D., Rogers, G. I. & Garland, C. D. 1989. Fatty acid and lipid composition of 10 species of microalgae used in mariculture. Journal of Experimental Marine Biology and Ecology 128, 219–240. Wilkinson, C. R. 1987. Microbial ecology on a coral reef. Search 18, 31–33. Wilson, S. K. 2000. Trophic status and feeding selectivity of blennies (Blenniidae: Salariini). Marine Biology 136, 431–437. Wilson, S. K. 2001. Multiscale habitat associations of detritivorous blennies (Blenniidae: Salariini). Coral Reefs 20, 245–251. Wilson, S. K. 2002 Nutritional value of detritus and algae in blenny territories on the Great Barrier Reef. Journal of Experimental Marine Biology and Ecology 271, 155–169.

308

DETRITUS IN THE EPILITHIC ALGAL MATRIX AND ITS USE BY CORAL REEF FISHES

Wilson, S. K. & Bellwood, D. R. 1997. Cryptic dietary components of territorial damselﬁshes (Pomacentridae, Labroidei). Marine Ecology Progress Series 153, 299–310. Wilson, S. K., Burns, K. & Codi, S. 2001a. Identifying sources of organic matter in sediments from a detritivorous coral reef ﬁsh territory. Organic Geochemistry 32, 1257–1269. Wilson, S. K., Burns, K. & Codi, S. 2001b. Sources of dietary lipids in the coral reef blenny Salarias patzneri. Marine Ecology Progress Series 222, 291–296. Zoufal, R. & Taborsky, M. 1991. Fish foraging periodicity correlates with daily changes of diet quality. Marine Biology 108, 193–196.

309

Oceanography and Marine Biology: an Annual Review 2003, 41, 311–354 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR CRAIG R. SMITH 1 & AMY R. BACO 1,2 Department of Oceanography, University of Hawaii at Manoa, 1000 Pope Road, Honolulu, HI, 96822, USA e-mail: [email protected] 2 present address: Biology Department, Woods Hole Oceanographic Institution, Woods Hole, MA 02543, USA e-mail: [email protected]

1

Abstract The falls of large whales (30–160 t adult body weight) yield massive pulses of labile organic matter to the deep-sea ﬂoor. While scientists have long speculated on the ecological roles of such concentrated food inputs, observations have accumulated since the 1850s to suggest that deep-sea whale falls support a widespread, characteristic fauna. Interest in whalefall ecology heightened with the discovery in 1989 of a chemoautotrophic assemblage on a whale skeleton in the northeast Paciﬁc; related communities were soon reported from whale falls in other bathyal and abyssal Paciﬁc and Atlantic sites, and from 30 mya (million years ago) in the northeast Paciﬁc fossil record. Recent time-series studies of natural and implanted deepsea whale falls off California, USA indicate that bathyal carcasses pass through at least three successional stages: (1)

(2)

(3)

a mobile-scavenger stage lasting months to years, during which aggregations of sleeper sharks, hagﬁsh, rat-tails and invertebrate scavengers remove whale soft tissue at high rates (40–60 kg d1); an enrichment opportunist stage (duration of months to years) during which organically enriched sediments and exposed bones are colonised by dense assemblages (up to 40 000 m2) of opportunistic polychaetes and crustaceans; a sulphophilic (“or sulphur-loving”) stage lasting for decades, during which a large, species-rich, trophically complex assemblage lives on the skeleton as it emits sulphide from anaerobic breakdown of bone lipids; this stage includes a chemoautotrophic component deriving nutrition from sulphur-oxidising bacteria. Local species diversity on large whale skeletons during the sulphophilic stage (mean of 185 macrofaunal species) is higher than in any other deep-sea hard substratum community.

Global species richness on whale falls (407 species) is also high compared with cold seeps and rivals that of hydrothermal vents, even though whale-fall habitats are very poorly sampled. Population-level calculations suggest that whale falls are relatively common on the deep-sea ﬂoor, potentially allowing macrofaunal species to specialise on these habitat islands; to date, 21 macrofaunal species are known only from whale falls and may be whale-fall specialists. Nonetheless, whale falls also share 11 species with hydrothermal vents and 20 species with cold seeps, and thus may provide dispersal stepping stones for a subset of the vent and seep faunas. Molecular evidence also suggests that whale falls provided evolutionary stepping stones for the bathymodiolin mussel lineage to move down the continental slope and into deep-sea vent and

311

CRAIG R. SMITH & AMY R. BACO

seep habitats. Finally, whale-fall bacteria have proven to be a novel source of cold-adapted enzymes of potential utility in cold-water detergents. Despite these scientiﬁc advances, major gaps persist in our understanding of the microbial processes, reproductive strategies, population genetics, and biogeography of whale-fall communities.

Scientiﬁc history of whale-fall ecology Whales are the giants of the ocean, with the eight largest cetacean species attaining body weights of 30 t to 160 t (Lockyer 1976). A sunken whale carcass provides a massive food fall to the normally organic-poor deep-sea ﬂoor; for example, the organic carbon contained in a 40-t whale (⬃2 106 g C) is equivalent to that typically sinking from the euphotic zone to a hectare of abyssal sea ﬂoor over 100 yr to 200 yr (e.g. Smith & Demopoulos 2003). The sediments directly underlying a sunken whale carcass (which covers roughly 50 m2) experience an initial pulse of labile organic material equivalent to ⬃2000 yr of background organiccarbon ﬂux. In part because of the massive size of whales, scientists have long speculated on the ecological effects of whale carcasses sinking to the deep-ocean ﬂoor. In 1934, Krogh commented that “whalebone whales represent the maximum energetic efﬁciency attained in the ocean” by feeding low on the oceanic food web and attaining enormous size. He speculated that the dead bodies of large animals (including whales) may “constitute the ultimate food for abyssal fauna” (Krogh 1934b). In particular, Krogh (1934a) calculated for the Southern Ocean that the ﬂux of whale-fall biomass to the sea ﬂoor may be ⬃0.5 g m2 yr1 (Krogh 1934a,b), and decided that it is “practically certain that the bottom fauna must obtain a more than negligible amount of food from fairly large animals sinking down from the surface.” In considering food sources for the deep sea, Bruun (1956) noted that whale ear bones are often trawled or dredged from the abyssal sea ﬂoor, and that a dead whale of 50 t “must attract scavengers for a long time and thus form a local focus of abundant food for predators.” Stockton & DeLaca (1982) speculated similarly that very large food falls, such as dead cetaceans, might yield localised development of dense communities at the deep-sea ﬂoor, possibly with an unusual (or “characteristic”) species structure. They suggested that the rise and fall of such localised benthic populations might take “many years.” In parallel with whale-fall speculations, evidence has accumulated in the taxonomic literature for nearly 150 yr that deep-sea whale remains support a widespread, characteristic fauna (Fig. 1). In 1854, S. P. Woodward described a small mytilid mussel (now known as Adipicola pelagica) living in whale blubber found ﬂoating off the Cape of Good Hope, South Africa (Dell 1987). This species was again reported in 1927 from whale debris in the North Atlantic, and in 1964 living in abundance on a whale skull recovered from 439 m off South Africa (Dell 1987). Dell (1987) concluded that A. pelagica is distributed from the Azores to South Africa living attached to whale remains at the deep-sea ﬂoor (400–1800 m); its occasional recovery in surface waters results from debris ﬂoating up from carcasses rotting at the sea ﬂoor. Another species of mytilid, “Adula” (now Adipicola) simpsoni was noted by Tebble (1966) to live in abundance in “a quite exceptional habitat,” (i.e. on weathered whale skulls trawled from the sea ﬂoor off Scotland, Ireland and the Orkney Islands). From the south Paciﬁc, Marshall (1987) and Gibbs (1987) described, respectively, a new family of limpets (Osteopeltidae) and a sipunculid species (Phascolosoma saprophagicum) 312

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Figure 1 Location of known deep-sea whale-fall sites studied in the world ocean, as well as the locations of known hydrothermal vents, and cold seeps at the deep-sea ﬂoor (Lonsdale 1979, Sibuet and Olu 1998, Van Dover et al. 2002). Note that the whale-skeleton symbol at ⬃30°N in the northeast Paciﬁc represents ﬁve deep-sea whale skeletons studied off southern California.

living in abundance on oily whale skulls trawled from 800–955 m on the Chatham Rise, near New Zealand. Both of these new taxa were surmised to feed on whale bones, with P. saprophagicum ingesting whale-bone oil (Gibbs 1987) and the osteopeltid limpet grazing on bacteria decaying the bones (Marshall 1987). Marshall and Gibbs also noted that the Chatham Rise skulls were encrusted with thousands of two unrecorded species of mytilid bivalves (Marshall 1987) and supported a “a rich fauna of mussels, gastropods, harpacticoid copepods, polychaetes and sipunculans.” Finally, in 1989 Warén reported a second osteopeltid limpet from whale bone trawled off Iceland (Warén 1989). This series of ﬁnds spurred Dell (1987) to note insightfully that the fauna of large organic debris at the deep-sea ﬂoor (including whale remains) was likely to become better known in the future due to increasing commercial trawling in deep water. In 1989, ecological understanding of whale-fall communities advanced substantially with the recognition that deep-sea whale skeletons may harbour chemoautotrophic assemblages (Smith et al. 1989). Based on the ﬁrst submersible observations and quantatitive samples of a deep-sea whale fall, Smith et al. (1989) reported large communities of bacteria, vesicomyid clams, mytilid mussels and gastropods supported by an oil-rich whale skeleton at 1240 m off California, in Santa Catalina Basin. They also noted that several of these species had been recovered from three whale skulls trawled at other bathyal sites off California. Several of the abundant whale-skeleton species (including two species of vesicomyid clams and the extremely abundant mytilid Idas washingtonia) contained sulphur-oxidising 313

CRAIG R. SMITH & AMY R. BACO

chemoautotrophic endosymbionts that appeared to use sulphide derived from anaerobic decomposition of concentrated bone lipids (Smith et al. 1989, Smith 1992, Deming et al. 1997). Based on the calculated abundance of whale falls and faunal similarities to hydrothermal vents, Smith et al. (1989) hypothesised that whale skeletons might provide important dispersal stepping stones for species (including some from hydrothermal vents and cold seeps) dependent on sulphide availability at the deep-sea ﬂoor. The ﬁnds of Smith et al. (1989) were quickly followed by discoveries of chemoautotrophically based assemblages on deep-sea whale skeletons from four other bathyal sites off California (McLean 1992, Smith & Baco 1998, C. Smith unpubl. obs.), and from the western Paciﬁc at 4000 m on the Torishima Seamount (Fujioka et al. 1993, Wada et al. 1994). The demonstration of chemoautotrophic endosymbiosis in Idas washingtonia (Smith et al. 1989, Deming et al. 1997) also suggested that other closely related mytilids (e.g. Adipicola spp., and Idas spp.) found on deep-sea whale bones from numerous locations off New Zealand (Dell 1987, 1996), off Japan (Y. Shiriyama, pers. comm.) and in the Atlantic (Tebble 1966, Dell 1987) might be utilising chemoautotrophy. In addition, apparently chemoautotrophic fossil communities (including Idasola (now Idas)) were discovered on fossilised whale remains from the bathyal northeast Paciﬁc dating from the Oligocene (⬃30 mya) (Squires et al. 1991, Goedert et al. 1995). Concurrently, new species of invertebrates (e.g. ﬁve limpets, a mytilid and two polychaetes) were documented from whale bones recovered off New Zealand and California (Pettibone 1993, Marshall 1994, Bennett et al. 1994, Dell 1996). When considered together, these ﬁndings suggested that chemoautotrophic assemblages, and specialised whale-bone communities, colonise whale falls over wide areas of the modern deep-sea ﬂoor (Smith 1992, Fig. 1) and have been utilising this habitat for tens of millions of years. These results also led to speculation that whale falls contribute signiﬁcantly to deep-sea diversity by providing specialised habitats, and by facilitating the dispersal of some vent-seep taxa (e.g. Committee on Biological Diversity in Marine Systems 1995, Butman et al. 1995). By 1995, the ecology, biogeography and evolution of deep-sea whalefall communities had become topics of broader interest to the oceanographic and marine biological communities, setting the stage for more detailed ecological and phylogenetic studies of whale falls.

Manipulative studies of whale falls off Southern California Although a substantial number of deep-sea whale skeletons had been sampled (mostly accidentally) by 1995, the ecology and biogeography of whale-fall communities remained very poorly understood. In particular, very little was known concerning (a) faunal succession following the arrival of a fresh whale carcass at the deep-sea ﬂoor, (b) persistence times of whale-bone chemoautotrophic assemblages, and (c) faunal relationships between whale-fall, cold-seep and hydrothermal-vent assemblages. Such information is essential for our understanding of the dynamics of whale-fall habitat islands, the recycling of large parcels of labile organic matter, and the ecology and evolution of sulphophiles and opportunistic species at the deep-sea ﬂoor. With funding ultimately derived from a variety of sources (the U.S. National Science Foundation, the National Undersea Research Center – Alaska, The National Geographic Society, and the British Broadcasting Corporation), in 1992 the University of Hawaii initi314

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

ated experimental studies of whale-fall communities at the bathyal sea ﬂoor. Overall goals were to (a) evaluate deep-sea community response to intense pulses of organic enrichment such as that resulting from a whale fall, and (b) elucidate the importance of whale falls as organic and sulphide-rich habitat islands at the deep-sea ﬂoor (Smith et al., in press). The general approach involved use of manned submersibles (e.g. the DSRV ALVIN) and Remotely Operated Vehicles (ROV’s) to conduct time-series studies of natural and experimentally implanted whale carcasses at the ocean ﬂoor. Whale carcasses for experimental emplacement were obtained through NOAA’s Marine Mammals Stranding Network, which monitors the stranding of marine mammals along all USA coasts. When a suitable carcass became available for deep-sea emplacement off California, a team of scientists ﬂew to the site from the University of Hawaii, towed the carcass to an appropriate drop site using a vessel of opportunity, and then sank the carcass to the sea ﬂoor. Because dead whales used in experiments had remained at the sea surface after death and thus had generated decompositional gases (Allison et al. 1991), substantial amounts of steel ballast (600–2700 kg) were used to sink each carcass. Because of the high costs and logistical difﬁculties of sinking dead whales, a limited number of carcasses (n 3) could be implanted at the sea ﬂoor for experimental study (Table 1). These three experimental whale falls, along with two natural skeletons discovered by chance off southern California (Fig. 2), were then visited at time periods ranging from 1 wk to 46 yr after estimated carcass arrival at the sea ﬂoor (Fig. 3). The resulting data provided the basis for the ecological and biogeographic syntheses below.

Patterns of succession on southern California whale falls Patterns of succession on whale carcasses are of broad ecological interest because they provide insights into deep-sea community response to extreme point-source enrichment, both natural (e.g. from whale falls) and anthropogenic. When the whale-fall experiments were initiated, ecologists had detailed understanding of the effects of organic-loading on shallow-water benthos (e.g. Pearson & Rosenberg 1978, Rhoads et al. 1978, Weston 1990, Zmarzly et al. 1994) but could only speculate on the community effects of intense organic loading, in the form of a whale fall, at the deep-sea ﬂoor (e.g. Krogh 1934a, Bruun 1956, Stockton & Delaca 1982, Smith 1985, Bennett et al. 1994). The timescales over which 5–35 t of solid, labile organic material might become assimilated into the seaﬂoor community were unconstrained (Stockton & DeLaca 1982) as were the periods of local community recovery after dissipation of enrichment from a whale fall (although deep-sea successional studies on small scales suggested recovery times 2 yr (Grassle & Morse-Porteous 1987, Smith & Hessler 1987, Kukert & Smith 1992, Snelgrove et al. 1994)). Both issues are of relevance to deep-sea patch dynamics and carbon ﬂux (Stockton & Delaca 1982; Smith 1985, 1986; Snelgrove et al. 1992, 1994, Etter & Caswell 1994, Butman et al. 1995, 1996), and to predicting the effects of analogous anthropogenic organic enrichment in the deep sea (e.g. relocation of sewage sludge, ﬁshery discards, and disposal of animal and medical wastes (Gage & Tyler 1991)). Based on previous studies of deep-sea scavengers, analogies with shallow-water organicenrichment communities, and initial whale-skeleton ﬁnds, fresh whale falls at the bathyal sea ﬂoor off California were expected to pass through four overlapping stages of ecological succession (Bennett et al. 1994, Smith et al. 1998): 315

Whale species

Blue or ﬁn

Balaenopterid? Gray Gray Gray

Santa Catalina Basin (natural)

San Nicolas slope (natural) San Clemente Basin (implanted) San Diego Trough (implanted) Santa Cruz Basin (implanted)

316 1980–90? 1992 1996 1998

⬃1948

⬃60 ⬃40 10 5 35

Time of arrival at sea ﬂoor

Estimated carcass wt (103 kg)

960 1960 1220 1675

1240

Water depth (m)

1988, 1991, 1995, 1999 1995 1995, 2000 1996, 1998, 1999 1998, 1999

Year(s) sampled

33°20 32°26 32°35 33°30

33°12

Latitude (N)

Natural and experimentally implanted whale carcasses studied off California by the University of Hawaii research effort.

Site

Table 1

119°59 118°9 117°30 119°22

118°30

Longitude (W)

CRAIG R. SMITH & AMY R. BACO

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Figure 2 Location of whale falls (sites A, B, C, E and F) studied off southern California. The location of the San Clemente Seep (site D) is also indicated. Depth contours are in metres, the ordinate is in degrees north latitude, and the abscissa in degrees west longitude.

(1) (2)

(3)

(4)

A mobile-scavenger stage, during which soft tissue would be removed from the carcass by dense aggregations of large, active necrophages (cf. Dayton & Hessler 1972, Isaacs & Schwartzlose 1975b, Hessler et al. 1978, Smith 1985). An enrichment-opportunist stage, during which dense assemblages of heterotrophic macrofauna (especially polychaetes and crustaceans) would colonise the bones and organically-enriched sediments surrounding the whale fall (cf. Turner 1977, Pearson & Rosenberg 1978, Levin et al. 1994). A sulphophilic (or “sulphur-loving”) stage, during which a chemoautotrophic assemblage would colonise the bones as they emitted sulphide during anaerobic bacterial decomposition of bone lipids. Methane might also be released during whale-tissue decay (Allison et al. 1991, Naganuma et al. 1996), fostering freeliving or endosymbiotic bacterial methanotrophs. A reef stage, occurring after the depletion of whale organic material, during which the mineral remnants of whale skeletons would be colonised primarily by suspension feeders exploiting ﬂow enhancement (e.g. Jumars & Gallagher 1982) and hard substrata.

Time-series studies of ﬁve carcasses at the bathyal sea ﬂoor, two natural whale falls and three experimentally implanted whale carcasses (Table 1, Fig. 2), provide strong evidence of the ﬁrst three successional stages (Bennett et al. 1994, Smith et al. 1998, Baco-Taylor 2002, Smith et al. 2002, Baco & Smith, unpubl. data). The structure and duration of these stages are discussed below. 317

CRAIG R. SMITH & AMY R. BACO

Figure 3 Times since arrival at the sea ﬂoor at which whale-fall sites off southern California have been sampled. For the San Nicolas carcass, an arrival of 1990 is used because this is the latest likely arrival time.

The mobile-scavenger stage Carcasses studied at times of 0.5–1.5 months after arrival at the sea ﬂoor (n 2) exhibited clear evidence of a mobile-scavenger stage. Both carcasses were largely intact but were covered with hundreds of hagﬁshes (predominantly Eptatretus deani but including Mixine circifrons) consuming soft tissue (Table 2). Sleeper sharks (Somniosus paciﬁcus), ranging in size from approximately 1.5 m to 3.5 m, were observed feeding voraciously on one carcass and in the vicinity of the other (Smith at al. 2002). Observed feeding activities and bite marks suggested that S. paciﬁcus had removed more soft tissue from the carcass at 1.5 months than had any other species. Other notable scavengers at 0.5–1.5 months included huge numbers of small (⬃0.5 cm long) lysianassid amphipods (Santa Cruz Basin carcass) and small numbers of large lithodid crabs, possibly Paralomis multispina (San Diego Trough carcass) (Table 2). Assuming hagﬁsh densities similar to those estimated for the 1300-m deep Santa Catalina Basin (370 km2; Smith 1985, Martini 1998), after 0.5 months to 1.5 months, the whale carcasses had drawn Eptatretus deani from minimum areas of ⬃1–2 km2 (or a radius of 0.6–0.8 km). By 4 months for the 5000-kg San Diego Trough carcass, and by 18 months for the 35 000-kg Santa Cruz Basin carcass, 90% of whale soft tissue had been removed, with only a small number of megafaunal scavengers remaining, indicating that the mobile-scavenger stage was drawing to a close (Smith et al. 2002). Thus, the duration of the mobile-scavenger 318

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

stage for megafauna at 1200–1800 m off California appears to range from the order of 4–5 months to 1.5–2.0 yr, depending on carcass size. Assuming that the soft tissue of a whale carcass constitutes about 90% of its total wet weight (cf. Robineau & de Buffrénil 1993), bathyal scavenger assemblages off California remove tissue from whale carcasses at rates of roughly 40–60 kg day1. These scavenging rates are approximately an order of magnitude higher than recorded on much smaller carcass falls (1–4 kg) off California (Smith 1985). Nonetheless, the scavenging rates for small and large parcels at bathyal depths off southern California are all well ﬁtted by a single logarithmic curve, in contrast to scavenging rates from the abyssal North Atlantic (Jones et al. 1998) (Fig. 4). The location of the whale-fall points on the logarithmic curve suggest that the whale-fall implantations are close to saturating the scavenging capacity of California bathyal ecosystems (i.e. whale falls (or other falls of labile organic material) that are larger in mass than 35 000 kg are likely to be scavenged at similar rates). If this is true, the mobile-scavenger stage for an adult blue-whale carcass of 100 000 kg at bathyal depths off southern California may last ⬃5 yr. Based on the relatively low scavenging rates for small cetaceans obtained by Jones et al. (1998), the mobile scavenger stage may last much longer for large whale falls in the abyssal North Atlantic. It is interesting to note that while megafaunal scavengers had largely dispersed from the San Diego Trough at 4 months and the Santa Cruz carcass at 1.5 yr, highly mobile macrofauna thought to be scavengers, in particular calanoid copepods (K. Wishner, pers. comm.) were very abundant on the whale bones at this time (Baco-Taylor 2002, Baco & Smith, unpubl. data). Thus, it seems likely that the mobile-scavenger stage itself undergoes a temporal succession as the remaining fragments of soft tissue attached to the carcass diminish in size, and are in turn exploited by species of scavengers of successively smaller body size (e.g. starting with sleeper sharks, hagﬁshes and macrourids, passing through lysianassid amphipods, and ending with calanoid copepods).

The enrichment-opportunist stage An enrichment-opportunist stage, during which dense assemblages of heterotrophic macrobenthos colonise organic-rich sediments and bones, was evident on carcasses visited at times of 4 months to 1.5 yr after arrival at the sea ﬂoor (the 5000-kg San Diego Trough carcass and the 35 000-kg Santa Cruz Basin carcass, respectively). This stage appears to begin in organicallyenriched sediments surrounding the skeleton but eventually includes the bone epifauna as well. The organically-enriched sediments within 1–3 m of each carcass were colonised by extremely high densities of macrofauna (Fig. 5). Around the San Diego Trough skeleton, a bed of free-living, centimetre-long polychaetes (Vigtorniella n. sp. and two undescribed species of dorvilleids; Dahlgren et al., unpubl. data, B. Hilbig, pers. comm.) undulated in the near-bottom ﬂow, resembling a ﬁeld of white grass; the bones themselves harboured high densities of dorvilleid polychaetes (Baco-Taylor 2002, Baco & Smith, unpubl. data). Large numbers of minute white gastropods (a new genus; J. McLean, pers. comm.) and juvenile bivalves colonised sediments around the Santa Cruz skeleton, and some bones of this skeleton were densely covered with writhing masses (thousands of individuals) of Vigtorniella n. sp. (Dahlgren et al., unpubl. data). Sediment macrofaunal densities attained 20 000–45 000 m2 within 1 m of the skeletons (Fig. 5); these are the highest ever reported for macrobenthos below 1000 m depths. In contrast, species diversity was dramatically reduced within 1 m of carcasses (Fig. 5). Dominant macrofauna common to both skeletons 319

CRAIG R. SMITH & AMY R. BACO

Figure 4 Top: Linear plot of scavenging rates of soft tissue from carrion falls off southern California as a function of carrion-fall mass. The equation for the plotted logarithmic curve is given. Data from Smith (1985) and Smith et al. (2002). Bottom: Log-linear plot of scavenging rates for carrion falls off southern California (diamonds) (data from Smith 1985 and Smith et al. 2002, as above) and for small cetacean carcasses in the abyssal North Atlantic (squares) (Jones et al. 1998). Note that scavenging rates for small cetacean carcasses in the abyssal North Atlantic fall well below the logarithmic curve ﬁtted for southern California carrion falls.

320

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Figure 5 Top: Sediment macrofaunal densities around whale falls in the San Diego Trough (SDT) at 4 months, and in the Santa Cruz Basin at 18 months. Both carcasses are in the enrichment-opportunist stage. Means one standard error are given. Bottom: Macrofaunal species diversity, based on three diversity indices, versus distance for the Santa Cruz Basin carcass at 18 months. Rarefaction E(51)/20 is the expected number of species in a normalised sample size of 51 individuals, divided by a constant of 20 to allow all three indices to be plotted on a single y-axis. Data from Smith et al. (2002).

321

CRAIG R. SMITH & AMY R. BACO

Table 2 Estimated total megafaunal abundance on whale carcasses at the sea ﬂoor for 0.5 and 1.5 months. Note that the original wet weight of the San Diego Trough carcass was 5000 kg, and that of the Santa Cruz carcass was 35 000 kg. Estimated abundances of lysanassid amphipods are extremely rough. From Smith et al. (2002). Megafaunal taxon

San Diego Trough carcass (t 0.5 months)

Santa Cruz Basin carcass (t 1.5 months)

Eptatretus deani Nezumia stelgidolepis Lithodid crabs (Paralomis multispina?) Small lysianassid amphipods Somniosus paciﬁcus

⬃ 300 1–2 2–4 0 1 observed on periphery

400–800 0 0 105–106? 1–3

included Vigtorniella n. sp., the dorvilleid polychaete Ophryotrocha sp. A, and the cumacean Cumella sp. A, all of which exceeded densities of 3000 m2 within 3 m of the carcass (Table 2). It is worth noting that the dominant species abounding in whale-enriched sediments (e.g. Vigtorniella n. sp., and the undescribed dorvilleids) have not been collected in the background communities. This suggests rapid recruitment and population growth for these relatively sessile species, reaching densities of 3000–10 000 ind. m2 in 4 months. The occurrence of a high-density, low-diversity assemblage in organically-enriched sediments near the whale carcasses is highly reminiscent of macrofaunal community patterns around point sources of organic enrichment in shallow water, for example, sewage outfalls and salmon pens (e.g. Pearson & Rosenberg 1978, Weston 1990, Zmarzly et al. 1994). In addition, there is some faunal similarity at the generic level, with dorvilleids in the genus Ophryotrocha responding to deep-sea whale falls and to sewer outfalls at shelf depths on the California coast (Levin & Smith 1984, Zmarzly et al. 1994). It appears that intense pulses of organic enrichment (e.g. due to whale falls, kelp falls, etc.) are common enough at bathyal depths off California to have allowed the evolution of bathyal enrichment opportunists. The duration of the enrichment-opportunist stage is likely to vary substantially with whale-carcass size and is still difﬁcult to constrain. For the 5000-kg San Diego Trough carcass, enrichment opportunists were abundant in adjacent sediments at 4 months but absent after 2 yr (Smith et al. 2002, C. Smith, unpubl. data), indicating a stage duration of 2 yr.

The sulphophilic stage The fresh bones of large whales, for example, the vertebrae of baleanopterids, may exceed 60% lipid by weight (Allison et al. 1991, D. Schuller, unpubl. data, S. Macko, pers. comm.). Thus, whale-bone lipids may constitute roughly 5–8% of the total body mass (Allison et al. 1991, Robineau & de Buffrénil 1993), and the skeleton of a 40-ton whale carcass may hold 2000–3000 kg of lipids. Following removal of whale soft tissue by scavengers, whaleskeleton decay appears to be dominated by anaerobic microbial decomposition of the large lipid reservoirs within the bones (Smith 1992, Deming et al. 1997; Fig. 6). Sulphate reduction is particularly important, yielding an efﬂux of sulphide from the bones (Deming et al. 1997). As a consequence of the sulphide efﬂux, species exploiting sulphide-based chemoautotrophic production, as well as species of other trophic types (bacterial grazers, bone322

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Figure 6 Schematic of cross section of a whale vertebra resting at the sea ﬂoor during the sulphophilic stage of succession. The predominant decompositional processes occurring within in the bones are illustrated, which include: (1) Diffusion of sulphate from sea water into the bone; (2) Sulphate reduction by anaerobic bacteria decomposing lipids in the lipid-rich bone core; (3) Diffusion of sulphide outward from the bone core, (4) Sulphide oxidation, and organic-matter synthesis, by sulphur-oxidising bacteria living on the bone surface and within the tissues (i.e. endosymbiotically) of vesicomyid clams and other invertebrates (Smith 1992, Deming et al. 1997).

lipid consumers, predators) able to tolerate elevated sulphide concentrations, are expected to colonise the whale skeleton, yielding the sulphophilic stage (Smith et al. 1998). Strong evidence of a recurrent sulphophilic stage comes from carcasses at the bathyal sea ﬂoor off California for periods between 2 yr and 51 yr (n 4 carcasses). This stage is characterised by several key components including: (1) (2) (3) (4)

diverse assemblages of heterotrophic and chemoautotrophic bacteria growing on bone surfaces and within bone cracks and trabaculae (Allison et al. 1991, Deming et al. 1997), large populations (typically 10 000 ind.) of the centimetre-long mytilid Idas washingtonia, which harbours chemoautotrophic endosymbionts (Bennett et al. 1994, Deming et al. 1997), large populations (hundreds to thousands) of the isopod Ilyarachna profunda and galatheid crabs, and frequently, large populations of diverse dorvilleid polychaetes, pyropeltid and cocculinid limpets (in particular Pyropelta musaica and Cocculina craigsmithi), provannid 323

CRAIG R. SMITH & AMY R. BACO

Table 3 Community structure on three whale skeletons during the sulphophilic stage. Estimated population sizes for each carcass are given. Time since whale-carcass arrival at the sea ﬂoor is indicated in parentheses. Data from Bennett et al. (1994), Smith & Baco (1998), Baco et al. (1999), BacoTaylor (2002) Smith et al. (2002), and Baco & Smith (unpubl. data). Taxon Mytilid Idas washingtonia Limpets Cocculina craigsmithi Pyropelta corymba Pyropelta musaica Other limpets Snails Mitrella permodesta Provanna lomana Eulimella lomana Juveniles and others Crustaceans Illyarachna profunda Amphipods Galatheids Misc. crustaceans Polychaetes Nereid sp. 1 Ampharetids Misc. polychaetes Total individuals Total species

San Clemente (3.4 yr)

San Nicolas (5 yr)

Santa Catalina (39–51 yr)

20 000

10 000

10 000

– – – –

300 1200 280 1800

1100 1000 1000 1200

3? – ⬃1000 1800

1800 1500 – 1700

1800 – – 800

900 400 800 9000

500 800 ⬃50 8000

1800 500 ⬃100 4000

⬃50 2500 10 000

⬃50 100 8000

40 000 191

30 000 180

⬃50 50? 1800 40 000 103

gastropods, and the columbellid snail Astyris permodesta (Table 3) (Smith et al. 1989, Allison et al. 1991, Bennett et al. 1994, Deming et al. 1997, Baco-Taylor, 2002, Baco & Smith, unpubl. data). This stage may also include vesicomyid and lucinid clams, and an occasional vestimentiferan worm, in sediments adjacent to the whale bones (Bennett et al. 1994, Feldman et al. 1998, Baco et al. 1999). For large skeletons, several other aspects of the sulphophilic stage on southern California whale falls are noteworthy. (1) (2)

Macrofaunal communities in this stage are large (exceeding 30 000 ind. to 40 000 ind.), species rich and trophically complex (Table 3, and see sections on Trophic relationships and Biodiversity patterns, pp. 326, 329). This successional stage may be remarkably long lasting. A well-developed, chemoautotrophic assemblage has persisted on the Santa Catalina Basin skeleton for at least 15 yr, that is, from 1987 to 2002 (Smith et al. 1989, 2002, Bennett et al. 1994, Baco-Taylor 2002, Smith & Baco, unpubl. data). In addition, radiomet324

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

(3)

ric dating using 226Ra/210Pb disequilibrium, indicates that the Santa Catalina Basin carcass in 1987 had already been at the sea ﬂoor for 39 ( 4) yr (Schuller et al., in prep). Considering that experimentally implanted carcasses have developed sulphophilic communities within 2 yr of reaching the sea ﬂoor (BacoTaylor 2002, Smith et al. 2002, Baco & Smith, unpubl. data), this suggests that large whale skeletons may support sulphophilic communities for at least 50 yr. A number of species (e.g. Idas washingtonia, Ilyarachna profunda, Cocculina craigsmithi, Pyropelta corymba, P. musaica) are extremely abundant on sulphide-rich whale skeletons but have rarely, or never, been collected in other habitats. These species may be whale-fall specialists that have evolved to exploit the productive and persistent habitat of sulphide-rich whale skeletons. The frequency distribution of abundances of macrofaunal species on whale skeletons in this stage also suggest the presence of a “core” group of species that have been associated with whale skeletons over evolutionary time (Bennett et al. 1994).

Although smaller whale skeletons (e.g. those of juvenile gray whales), support many species characteristic of the sulphophilic stage on large skeletons, stable-isotopic evidence suggests that most of the macrofaunal biomass on these small skeletons (including the dominant Idas washingtonia) is not derived from sulphide-based chemoautotrophic production of endosymbionts (e.g. those in I. washingtonia) (Baco-Taylor 2002, Baco & Smith, unpubl. data). Thus, in contrast to large whale skeletons in the sulphophilic stage, the macrofaunal communities on small skeletons are sulphide tolerant but do not appear to be predominantly chemoautotrophic (see Trophic relationships section, p. 326). It should be noted that communities of animals reported on whale skeletons from other oceanic regions, for example, the North Atlantic, South Atlantic, western Paciﬁc, and South Paciﬁc (Tebble 1966, Dell 1987, 1995, Marshall 1987, 1994, Warén 1989, Wada et al. 1994, Naganuma et al. 1996) also appear to fall into the sulphophilic stage. These communities are dominated by mytilid mussels closely related to I. washingtonia (Distel et al. 2000, BacoTaylor 2002, Baco et al., unpubl. data), many of which appear to derive nutrition from chemoautotrophic production (Baco-Taylor 2002, Baco et al., unpubl. data). In addition, where observed in situ or sampled relatively carefully, many of these communities included bacterial mats, cocculinid limpets and galatheid crabs.

The reef stage Time-series studies of whale skeletons thus far have yielded no direct evidence of a reef stage dominated by suspension feeders because the sulphophilic stage has occurred even on very old carcasses (e.g. the Santa Catalina Basin carcass at ⬃50 yr) (Baco-Taylor 2002, Smith et al. 2002). However, the sulphophilic stage does contain a few suspension feeders, including sabellid, chaetopterid and serpulid polychaetes, likely to be exploiting enhanced ﬂow conditions on the bones (Baco-Taylor, 2002, Baco & Smith, unpubl. data). Some of these taxa also occur in the background community on hard substrata (Bennett et al. 1994) and are likely to continue to exploit large, well calciﬁed skeletons even after depletion of whale organic matter, yielding a reef stage. On large skeletons, this stage may not be reached for many decades. On smaller skeletons (e.g., those of juveniles gray whales), this stage may be curtailed by relatively rapid decomposition and dissolution of the poorly calciﬁed bones (C. Smith, pers. obs. from the 5000 kg San Deigo Trough carcass). 325

CRAIG R. SMITH & AMY R. BACO

Trophic relationships on Southern California whale falls Bennett et al. (1994) cite ﬁve sources of organic matter potentially of major signiﬁcance in whale-fall habitats: (1) (2) (3) (4) (5)

whale organic material (e.g. soft tissues and lipids within the bones); free-living hetero- and chemoautotrophic bacteria; endosymbiotic, sulphur-oxidising chemoautotrophic bacteria; tissue of primary consumers; and detrital particles suspended in currents or deposited in sediments around the bones.

Based on 13C and 15N values for epifauna on skeletons at the sea ﬂoor for 4 months to 51 yr, the relative importance of these sources of organic matter varies with successional stage (Baco-Taylor 2002, Baco & Smith, unpubl. data). Our synthesis is based on the assumptions that the whole bodies of consumers have (a) 13C values within 0.3‰ to 1.9‰ of their food material (DeNiro & Epstein 1978, Rau et al. 1983, Fry & Sherr 1984), and (b) 15N values 1.3‰ to 5.3‰ heavier than their food material (DeNiro & Epstein 1981, Minagawa & Wada 1984). Because of the relatively large change in 15N values between consumers and their food source(s) (the so called “trophic shift”), 15N values are frequently useful for delineating trophic levels within food webs.

Mobile scavenger/enrichment opportunist stage Bone epifauna in the mobile scavenger and enrichment-opportunist stages at 4 months to 1.5 yr had 13C and 15N values indicating a single trophic level relying on whale organic material (Baco-Taylor 2002, Baco & Smith, unpubl. data). At the San Diego Trough skeleton at 4 months, the community appeared to derive its nutrition primarily from whale soft tissue. On the Santa Cruz whale community at 1.5 yr, 13C and 15N isotope values were more negative than at San Diego Trough, suggesting a greater dependence on bone lipids.

Sulphophilic stage The communities on all whale skeletons on the sea ﬂoor for 2 yr exhibited high species overlap and appeared to fall into the sulphophilic stage (see above). However, stable isotopic values revealed distinct differences in trophic structure between the communities on juvenile gray whales (San Diego Trough and San Clemente Basin skeletons) and those on the skeletons of larger whales (Santa Catalina Basin and San Nicolas skeletons). Small skeletons The mytilid Idas washingtonia was the community dominant on all skeletons on the bottom for 2 yr (Bennett et al. 1994, Baco et al. 1996, Baco-Taylor 2002, Baco & Smith, unpubl. data). Deming et al. (1997) documented chemoautotrophic endosymbioses in this species from the Santa Catalina Basin skeleton based on microscopy, enzymes assays and isotope values. However, the 13C and 15N values for I. washingtonia from the small skeletons (San Clemente Basin and San Diego Trough) at 2–8.25 yr were much higher than the Santa Catalina Basin and San Nicolas skeletons (Fig. 7), suggesting that I. washingonia did not rely on chemoautotrophy at the small skeletons. 326

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Figure 7 Variation in 13C and 15N values for species common to skeletons that have been on the bottom for 2 yr. (a) Idas washingtonia; (b) Ilyarachna profunda; (c) Amphipod sp. D. Note that for the larger skeletons (i.e. Santa Catalina Basin (SCB) and San Nicolas (SN)), all three species tend to be substantially lighter in both isotopic ratios, indicating a much greater dependence on chemoautotrophic production than on the smaller skeletons in the San Diego Trough (SDT) and San Clemente Basin (Clem) and on wood (from Baco-Taylor 2002, Baco & Smith, unpubl. data).

Rather than dependence on chemoautotrophy, most of the other species on small skeletons at 2–8.25 yr appeared to depend on bone lipids ( 13C values of 20.0 to 13.0 ‰ and

15N values of 13.6‰ to 21.0‰). Even species that were found on all skeletons 2 yr, I. washingtonia, Ilyarachna profunda and Amphipod sp. D, had much higher 13C and 15N values on the San Clemente Basin and San Diego Trough skeletons than on the Santa Catalina Basin and San Nicolas skeletons (Fig. 7). Thus, in contrast to large whale skeletons in the sulphophilic stage (discussed below), the macrofaunal communities on small skeletons are sulphide tolerant but do not appear to be predominantly chemoautotrophic. The San Clemente Basin and San Diego Trough skeletons were from juvenile gray whales, whose vertebrae were poorly calciﬁed compared with adults (Jones et al. 1984). The juvenile skeletons appeared to decompose much more rapidly that than those of adults whales, releasing lipid reservoirs relatively quickly. Because of the relatively small size of juvenile bones, the bone-lipid reservoir is also likely to be much smaller in juveniles than in adult whales. The gradual anaerobic breakdown of bone lipids appears to be the source of sulphides for chemoautotrophic production on whale falls (Smith 1992, Deming et al. 1997), so the lipid reservoir will be depleted more rapidly, and sulphides available for a shorter period of time, on skeletons of younger whales. It appears that there is a minimum size/degree of calciﬁcation required for a whale skeleton to sustain chemoautotrophic communities for extended periods (i.e. years). Juvenile skeletons have not been sampled between 4 months and 2 yr so it is conceivable that during this interval, the community may be at least partially dependent on chemoautotrophic production. Large Whale Falls Whale-fall communities in the sulphophilic stage on large skeletons (Santa Catalina Basin and San Nicolas slope) exhibited much more complex trophic 327

CRAIG R. SMITH & AMY R. BACO

structure, utilising organic material from chemoautotrophic endosymbionts, free-living bacteria, and whale tissue (Baco-Taylor 2002, Baco & Smith, unpubl. data). The majority of the species analysed on the larger skeletons had 13C values 20‰ and appear to be part of a food web dependent on chemoautotrophic sources of production. Nitrogen isotopic ratios of organisms apparently dependent on endosymbiotic chemoautotrophic production ranged from 0.9‰ to 14.4‰. Assuming a trophic shift of 3‰ to 5‰ (DeNiro & Epstein 1981, Minagawa & Wada 1984), there appear to be three to ﬁve trophic levels in the whaleskeleton food web, for example, producers with chemoautotrophic endosymbionts, primary consumers, secondary consumers and scavengers (Baco-Taylor 2002, Baco & Smith, unpubl. data). There were several species with very negative carbon isotopic values ( 13C from 36.5‰ to 29.6‰), indicative of reliance on chemoautotrophic endosymbionts (BacoTaylor 2002, Baco & Smith, unpubl. data). These species include vesicomyid clams, Idas washingtonia, and the polychaete dorvilleid sp. D. To date, polychaetes containing chemoautotrophic and/or nitrogen ﬁxing bacteria have not been reported from any hydrothermal vent habitat. However, dorvilleid sp. D (characterised by paired dorsal pouchlike structures on each segment, Baco-Taylor 2002, Baco & Smith, unpubl. data) had 15N values lighter than any other whale-fall species (0.9‰ to 4.0‰) and 13C values ranging from 31.7‰ to 29.6‰. Such 13C-depleted values are strongly suggestive of chemoautotrophic production, for example, via chemoautotrophic endosymbionts (e.g. Fisher et al. 1994, Deming et al. 1997). A similar dorvilleid with dorsal pouches and extremely depleted

13C values (90‰ to 35‰) has recently been found on northern California, Oregon, and Alaska seeps (Levin et al. 2000, and in prep). The next putative trophic level on large whale skeletons in the sulphophilic stage included the provannid snail Provanna lomana, the columbellid snail, Astyris (Mitrella) permodesta, and two species of ampharetid polychaetes. These four species had light isotopic values ( 13C values of 29.5‰ to 23.5‰ and 15N values of 1.3‰ to 11.1‰) consistent with chemoautotrophic endosymbionts, predation on species with symbionts, or grazing on free-living chemoautotrophic bacteria. Deming et al. (1997) could ﬁnd not evidence of sulphide-oxidising endosymbionts in A. permodesta. Isotope values for the three potential secondary consumers and/or scavenging species, Nereis anoculis, Amphipod sp. D, and Galathaeid sp. 3 ( 13C values of 26.6‰ to 20.0‰ and 15N values of 8.8‰ to 14.4‰), are more positive than expected if they were preying solely on species with chemoautotrophic endosymbionts and may reﬂect a mixed diet. A portion of the food web on large skeletons in the sulphophilic stage appeared to be dependent on bacterial mats. Bacterial mat 13C ranged from 23.4‰ to 19.4‰ and 15N ranged from 4.3‰ to 7.4‰. Two species which appear to feed on bacterial mats, Pyropelta musaica and Ilyarachna profunda, had 15N values ranging from 9.8‰ to 13.8‰, and

13C values similar to, or slightly heavier than, mat material. In contrast to the smaller San Diego Trough, San Clemente Basin and Santa Cruz skeletons, only two species in the Santa Catalina Basin and San Nicolas communities appeared to depend on whale organic material. These were the limpet species, Cocculina craigsmithi and several individuals of Pyropelta musaica, with 15N ranging from 17.6‰ to 19.5‰. This small range of isotope values suggests a single trophic level, with the limpets as secondary consumers of the whale organic material. The trophic structure of the large, old whale-skeleton communities has interesting parallels to hydrothermal-vent communities (Baco-Taylor 2002, Baco & Smith, unpubl. data). 328

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

(1)

(2)

(3)

15N values indicated at least three trophic levels ultimately supported by species with chemoautotrophic endosymbionts. East Paciﬁc hydrothermal-vent communities have similar trophic structure, with the ﬁrst level occupied by freeliving bacteria, which support ⬃2.5 trophic levels of invertebrate consumers (Van Dover & Fry 1989). Like vent communities, the distinctly lower 15N values, 10‰, of the organisms apparently dependent on chemoautotrophic production and bacterial mats indicate that much of the organic nitrogen in the Santa Catalina Basin and San Nicolas whale-fall communities is of local origin (Rau 1981, Van Dover & Fry 1989). biomass distributions on large, old skeletons indicate dominance by species harbouring chemoautotrophic endosymbionts. Bivalves with chemoautotrophic endosymbionts, Vesicomya gigas and Idas washingtonia, comprised 58% of the molluscan biomass collected at the Santa Catalina Basin skeleton in 1991. The three species of limpets, which are most likely dependent on production by free-living chemoautotrophic bacteria and whale organic material, constituted 42% of the biomass (Baco-Taylor 2002, Baco & Smith, unpubl. data). These ﬁndings are similar to those from hydrothermal vents where biomass is dominated by organisms bearing chemoautotrophic endosymbionts (Sarrazin & Juniper 1999).

Biodiversity patterns on Southern California whale falls Diversity in whale-bone epifaunal communities varied with successional stage. The mobilescavenger and enrichment-opportunist stages are characterised by relatively low species richness, with totals of 38 and 18 macrofaunal species, respectively (Baco-Taylor 2002). The sulphophilic stage appears to harbour the greatest species richness, often with 100 macrofaunal species per skeleton. The sulphophilic stage on the San Nicolas skeleton was particularly speciose, with at least 190 species of macrofauna living on the bones (BacoTaylor 2002, Baco & Smith, in review). At all successional stages, roughly half of the known species richness (47–60%) was contributed by the polychaetes. The diversity on the chemoautotrophic whale skeletons of San Nicolas and Santa Catalina Basin was lower than in background sediments in the vicinity of the skeletons (Baco-Taylor 2002, Baco & Smith, in review). However, the average local species richness (185 species) on these two skeletons was higher than on any other type of deep-sea hard substratum, including other reducing habitats. Despite being one of the least-studied deep-sea reducing habitats, whale falls may harbour the highest levels of global species richness; thus far, 407 species are known from whale falls, with 91% of these species coming from California whale falls alone. As more whale falls are sampled in other deep-sea regions, the total number of species known from whale falls certainly will rise dramatically. By comparison, the much more intensively studied hydrothermal vents (Tunnicliffe 1991) are thus far known to harbour ⬃469 species worldwide, (Tunnicliffe et al. 1998) and ⬃230 species are known from cold seeps (Sibuet & Olu 1998, Poehls et al., in prep). The remarkable species richness on whale skeletons on local and (potentially) global scales may be explained by an unusually large number of trophic types found on whale bones including species with chemoautotrophic endosymbionts, bacterial grazers, generalised organic-enrichment respondents, whale-bone consumers, and more typical 329

CRAIG R. SMITH & AMY R. BACO

hard-substratum detritivores such as suspension and deposit feeders (Baco-Taylor 2002, Baco & Smith, in review).

Overall structure and mechanisms of succession on Southern California whale falls The overall structure of succession on deep-sea whale falls is longer and more complex than that on ﬁsh carcasses and other small carrion parcels in the deep sea. For carcass falls ranging up to ⬃50 kg in size, the enrichment-opportunist and sulphophilic stages typically appear to be absent (e.g. Smith 1985, 1986, Jones et al. 1998). Intense organic enrichment of nearby sediments does not occur because the mobile scavengers (including epibenthic forms at bathyal depths) can efﬁciently remove the smaller mass of soft tissue over very short timescales (typically days, Dayton & Hessler 1972, Hessler et al. 1978, Smith 1985, Jones et al. 1998). The primary effect of such food falls on the local sediment community appears to be physical disturbance resulting from the vigorous feeding and swimming activities of scavengers (Smith 1986). On scavenged ﬁsh carcasses, the remaining bones are apparently too small and contain an inadequate organic-matter reservoir to sustain the development of a sulphur-oxidising microbial assemblage; hence, the sulphophilic stage does not develop (Smith 1985). However, the skeletons of small cetaceans such as dolphins, as well as cow bones artiﬁcially placed on the deep-sea ﬂoor, contain large enough organic reservoirs to support mats of Beggiatoa (a sulphate reducing bacterium) (Kitazato & Shirayama 1996) and to sustain limited recruitment of bathymodiolin mussels with chemoautotrophic endosymbionts (Y. Shirayama, pers. comm., Baco et al., unpubl. data). Several aspects of whale-fall community change are of relevance to the consideration of successional mechanisms. As with carrion falls in terrestrial environments (Schoenly & Reid 1987), biotic succession on southern California whale falls appears to be largely a continuum of change, with temporal overlap in the occurrence of the characteristic species from different stages. For example, on 4 month-old and 18 month-old carcasses (the San Diego Trough and Santa Cruz carcasses, respectively), components of both the mobile-scavenger and the enrichment-opportunist stages were present (Smith et al. 2002, Baco-Taylor 2002). In addition, on the Santa Cruz carcass at 18 months, components of the sulphophilic stage had begun to recruit, in particular Idas washingtonia (Baco-Taylor 2002, Baco & Smith, unpubl. data). Nonetheless, there appear to be periods of relatively rapid faunal change on the whale falls that can be considered to be loose successional-stage boundaries. The presence of soft tissue on carcasses elicits active feeding by large aggregations of megafaunal scavengers. When this tissue is depleted (within approximately 4 months for the 5000 kg carcass and 18 months for the 35 000 kg carcass), the abundance of scavenging megafauna drops abruptly (Smith et al. 2002). Similarly, based on analyses using Bray-Curtis similarity and non-metric multidimensional scaling, Baco-Taylor (2002, Baco & Smith, unpubl. data) found abrupt differences in species structure between skeletons at the sea ﬂoor for less than, and greater than, 2 yr. Thus, 2 yr marked the approximate boundary between the enrichment-opportunist and sulphophilic stages. These intervals of relatively rapid community change occurring between periods of relative community stasis are consistent with the concept of “successional stages” (Schoenly & Reid 1987). Successional changes on whale falls not only involve species turnover but also include 330

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

changes in both faunal mobility and trophic structure. The mobile-scavenger stage is dominated by very active swimmers (hagﬁshes, sharks, lysianiassid amphipods), the enrichmentopportunist stage is dominated by moderately mobile epibenthos (e.g. gastropods, dorvilleids, chrysopetalids), and the sulphophilic stage by sessile macrofauna (e.g. I. washingtonia, which attaches with byssal threads) and microbial mats (Bennett et al. 1994, BacoTaylor 2002, Smith et al., in press, Baco & Smith, unpubl. data). Trophic structure shifts from a predominance of scavengers, through carnivore–scavenger–omnivores, to macrofauna harbouring chemoautotrophic, sulphur-oxidising endosymbionts (Baco-Taylor 2002, Smith et al. 2002, Baco & Smith, unpubl. data). In fact, it is the dramatic shifts in mobility patterns and trophic structure that has led to the names of the successional stages. Connell & Slatyer (1977) identiﬁed three general mechanistic models to explain species turnover during successional change: (a) facilitation, (b) tolerance and (c) inhibition. Under facilitation, species arriving early in the successional sequence modify the habitat to facilitate, or “pave the way for”, the colonisation of later-stage species. Under the tolerance model, early-stage species are less tolerant of lower resource levels than are later-stage species, and thus early species are replaced by superior competitors. In the inhibition model, mature individuals of species from all stages inhibit habitat utilisation by all other individuals regardless of species; early successional species dominate initially because they disperse better and/or grow faster, whereas later-stage species ultimately dominate because they live longer and accumulate as early species die off. As originally formulated, all three of the Connell & Slatyer (1977) models rely heavily on competitive interactions. In the facilitation and tolerance models, early species are excluded (or killed) through competition with later-stage species. In the inhibition model, competition, particularly for space, is the primary interaction. It also should be noted that these models do not exhaust the full suite of reasonable interaction scenarios. For example, a null or noninteractive model might be considered, in which species abundances rise and fall independently of other species colonising the habitat. Such a model might apply if all species were held well below carrying capacity due to inadequate larval supply or heavy predation pressure, or if species-speciﬁc pathogens or toxic chemicals (e.g. sulphide) controlled population dynamics. In addition, there is no a priori reason in successional models to link the effects of early-stage species on later colonists (e.g. facilitation) with those of later colonists on early species. For example, it is quite possible (and, in fact, likely for the mobile-scavenger stage of whale falls) that early species facilitate the arrival of later successional species, and then disappear due to reasons other than interspeciﬁc competition. Given this conceptual framework, which mechanisms of succession apply to whale carcasses at the deep-sea ﬂoor? Clearly, facilitation is a dominant mechanism in the transitions between the mobile-scavenger and enrichment-opportunist stages, and between the enrichment-opportunist and sulphopilic stages (Baco-Taylor 2002, Baco & Smith, unpubl. data). Enrichment opportunists cannot colonise bone surfaces until scavengers have stripped off the soft tissue; nor will they recruit to nearby sediments until the frenzied feeding of scavengers has broadcasted a ﬁne rain of whale-tissue fragments over the surrounding sea ﬂoor. The mobile scavengers, however, are not ultimately excluded by competition with enrichment opportunists; the scavengers themselves deplete the carrion resource and then move off in search of other feeding opportunities (cf. Smith 1985, 1986). In turn, many species in the sulphophilic stage, in particular the dominant species I. washingtonia, beneﬁt from the development of sulphate-reducing microbial assemblages on and within the bones (Smith et al. 1989, Smith 1992, Deming et al. 1997). Thus, the colonisation of the bones by anaerobic 331

CRAIG R. SMITH & AMY R. BACO

microbial populations is necessary to facilitate the development of the chemoautotrophic assemblages. Once again, however, the decline of the sulphophilic stage is unlikely to be a consequence of competitive exclusion by later colonists (e.g. by suspension feeders). This stage necessarily declines as sulphate-reducing bacteria deplete lipid reservoirs within the whale bones, and sulphide levels drop below those required to sustain chemoautotrophic endosymbionts. Overall, facilitation appears likely to be the dominant process governing turnover of whale-fall successional stages. However, unlike Connell & Slatyer’s (1977) original facilitation model, the facilitation by early species in whale-fall succession is not ungraciously repaid with competitive exclusion by later colonists. Mechanisms of succession on deep-sea whale falls exhibit some similarities to those on carrion falls in terrestrial environments. As for deep-sea whale falls, facilitation may dominate successional changes in terrestrial carcasses (Connell & Slatyer 1977). In addition, a true mobile-scavenger stage does occur in some terrestrial habitats where large specialised necrophages (e.g. vultures) or facultative scavengers (e.g. minks, foxes, bears, wolves, hyenas) feed on carcasses as mobile adults, removing much of the soft tissue (e.g. Houston 1986, Anderson 2001). However, in many terrestrial ecosystems in North America and Europe, soft tissue persists on large carcasses for substantial periods of time (i.e. scavenging rates appear to be roughly an order of magnitude lower than observed on whale falls), and much of the soft tissue is consumed by the feeding larvae of saprophytic insects (Anderson 2001). In other words, unlike whale falls and other large carrion falls in the deep sea (e.g. Dayton & Hessler 1972, Isaacs & Schwartzlose 1975a, Smith 1985), soft-tissue reduction in many terrestrial ecosystems depends on a reproductive response by saprophytic species (e.g. blow ﬂies, carrion beetles and dermestid beetles) as well as on the decompositional activities of microbes (bacteria and fungi) (Anderson 2001, Byrd & Castner 2001, Merritt & Wallace 2001). The lack of large mobile scavengers in many terrestrial systems is likely to be a consequence of anthropogenic extinction of large vertebrates, which could act as facultative or obligate scavengers (e.g. bears, wolves, wolverines, coyotes, foxes, vultures and condors; Pulliainen 1988, Hewson 1984, 1995, Willey & Snyder 1989, Green et al. 1997). In contrast, in marine environments, human activities may have had the opposite effect, increasing the abundance of large scavengers by enhancing carrion availability through ﬁshery discards and trawling disturbance (Britton & Morton 1994). Thus, anthropogenic impacts may ultimately have driven the recycling of large carrion parcels along substantially different pathways in terrestrial and marine environments.

Biogeographic and evolutionary relationships of whale-fall communities Modern relationships Abundance of whale falls at the deep-sea ﬂoor How common are whale falls at the deep-sea ﬂoor? Are they frequent enough now, or have they been in the past, to allow faunal dispersal (e.g. by planktonic larvae) between adjacent whale falls? Such dispersal is essential for whale falls to serve as sulphide-rich stepping 332

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

stones for species dependent on chemoautotrophy (Smith et al. 1989, Kitazato & Shirayama 1996) and to allow the evolution of whale-fall specialists (Bennett et al. 1994). The abundance of whale-fall communities within a particular successional stage will be a function of the overall frequency of whale falls, and the duration of that particular stage (roughly 0.33–2 yr for the mobile-scavenger stage, 1–2 yr for the enrichment-opportunist stage, and 5–50 yr for the sulphophilic stage at bathyal depths off California). Smith et al. (1989) estimated that 500 gray whales sink to the sea ﬂoor each year within a northeast Paciﬁc habitat area of 8 km2 105 km2. This estimate used a gray-whale population size of 18 000, and assumed that 50% of dying whales sink to, and remain at, the sea ﬂoor. A sinking rate of 90% is probably realistic because most whales suffering natural mortality are in poor nutritional condition and negatively buoyant (Ashley 1926, D. W. Rice, pers. comm.). The percentage of sunken whales that remain at the sea ﬂoor will depend, in part, on the water depth and resultant hydrostatic pressure, which limits the generation of buoyant decompositional gases (Allison et al. 1991). Below a depth of 1000 m, the amount of microbial tissue decay required to reﬂoat a carcass (e.g. ⬃67% of carcass mass through fermentation) is prohibitive; the soft tissue will be scavenged and/or disintegrate long before sufﬁcient buoyancy can be generated (Allison et al. 1991). At shallower depths, there is some probability that gas generation will reﬂoat the carcass, although this will depend on the rate of soft tissue removal by scavengers versus microbial decomposition. A partially scavenged, but otherwise intact, gray-whale carcass has been found at 150 m depth in Alaskan waters (Thomas Shirley, pers. comm.), suggesting that 15 atm of hydrostatic pressure may, at least in cold water, be adequate to prevent decompositional buoyancy for large whales. Given these uncertainties, the assumption that 50% of dying whales sink to, and remain at, the sea ﬂoor seems to be a reasonable (and probably conservative) best guess. Using the approach of Smith et al. (1989), we have estimated current abundances of sea ﬂoor whale-fall communities in the ﬁrst three successional stages (Tables 4, 5). The calculations are in two parts, (a) whale falls resulting from gray whales in the northeast Paciﬁc, and (b) those resulting from the mortality of the nine most common large whale species throughout the global ocean. Within the gray-whale range, rough estimates suggest that whale-fall communities have mean nearest-neighbour distances ranging from 5 km to 16 km, depending on successional stage. The nearest-neighbour distances for the enrichment-opportunist and sulphophilic stages (5–13 km) fall well within documented larval transport distances and scales of gene ﬂow for animals living in other energy-rich, island-type habitats in the deep sea such as hydrothermal vents and cold seeps (Lutz et al. 1984, Black et al. 1994, 1998, Vrijenhoek 1997, Van Dover 2000, Marsh et al. 2001). It is thus entirely feasible that species attaining population sizes of 103–104 on California whale falls (e.g. I. washingtonia, Vigtorniella n. sp., and a number of gastropod species) may routinely disperse between whale falls, potentially using them as their primary habitat, or as dispersal stepping stones between other types of habitat islands (Smith et al. 1989). Our global calculations for the nine large whale species also indicate moderate nearest-neighbour distances of 12–30 km for the enrichment-opportunist and sulphophilic stages, again suggesting that species might routinely disperse between whale falls. In reality, whale falls are likely to be more closely spaced than calculated for the global ocean because whale mortalities are non-randomly distributed; they are likely to be concentrated along whale migration routes and in feeding grounds, which often occur near ocean margins (Gaskin 1982, Katona & Whitehead 1988, Smith et al. 1989, Butman et al. 1995, Perry et al. 1999). 333

Blue Bowhead Fin Gray Humpback Minke Right Sei Sperm

Common name

Balaenoptera musculus Balaenoptera mysticetus Balaenoptera physalis Eschrichtius robustus Megaptera novaeangliae Balaenoptera acutorostrata Balaena glacialis Balaenoptera borealis Physeter macrocephalus

Species

40–160 30–100 25–80 12–40 15–60 5–10 30–80 10–30 13–70

Approximate adult size range (103 kg ww) 8500 9000 128 000 26 000 36 000 935 000 8600 24 000 220 000

Recent population size estimate

0.04 0.05* 0.04 0.06 0.05 0.05* 0.03 0.08 0.05 Ave. 0.05

Natural mortality rate (yr1)

300 500 5100 1600 1800 47 000 300 1900 11 000 69 000

Annual population mortality

Lockyer 1976, Perry et al. 1999 Lockyer 1976, Sheldon & Rugh 1995 Lockyer 1976, Perry et al. 1999 Lockyer 1976, Rice et al. 1984 Lockyer 1976, Perry et al. 1999 Lockyer 1976, IWC** Lockyer 1976, Perry et al. 1999 Lockyer 1976, Perry et al. 1999 Lockyer 1976, Perry et al. 1999

References

Table 4 Body size, recent estimates of worldwide population size, and estimates of natural mortality for 9 of the 10 largest species of Cetacea. Bryde’s whale (Balaenoptera edeni) was not included because we could ﬁnd no recent estimates of population size. *No estimate of natural mortality for this species was available so the mean of the rate estimates for blue, ﬁn, gray, humpback, right, sei and sperm whales was used. **International Whaling Commission website (http://www.iwcofﬁce.org/estimate.htm), July 2002.

CRAIG R. SMITH & AMY R. BACO

334

Gray whales in the NE Paciﬁc Mobile scavenger stage Enrichment opportunist stage Sulphophilic stage Nine large whale species in the global ocean Mobile scavenger stage Enrichment opportunist stage Sulphophilic stage

Successional stage

800 1200 8000 69 000 103 000 690 000

1 (0.33–2) 1.5 (1–2?) 10 (5–50)

Number of sea-ﬂoor carcasses in stage

1 (0.33–2) 1.5 (1–2?) 10 (5–50)

Stage duration (yr) assumed (range)

335 5200 3500 520

1000 670 100

Mean sea-ﬂoor area per carcass (km2)

36 30 12

16 13 5

Mean nearest neighbour distance (km)

Table 5 Estimated average nearest neighbour distances for whale falls in the various successional stages for gray whales in the northeast Paciﬁc and nine large species of whales (combined) in the global ocean. Note that the estimates of stage duration time are rough, and chosen to be conservative. Also, note that the estimates assume population sizes and mortality rates (Table 3) to be at steady state, and that 50% of carcasses sink to, and remain on, the sea ﬂoor. Mean nearest neighbour distance (Pielou 1969), assuming a random distribution, 0.5 (mean area per carcass)0.5.

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

CRAIG R. SMITH & AMY R. BACO

Nearest neighbour distances are somewhat larger for the mobile scavenger stage (i.e. 16 km and 36 km, for gray whales and the nine pooled species, respectively) and it is very unlikely that the dominant members of this successional stage are whale-fall specialists. Deep-sea scavengers typically disperse as large adults, not as low-cost larvae broadcasted in the hundreds to thousands by spawning individuals. These scavengers appear to respond to carrion falls over distances of 10–1000 m (e.g. Smith 1985, Priede et al. 1991, Collins et al. 1998, Klages et al. 2002, discussion above for hagﬁshes) making it very difﬁcult to specialise on whale falls spaced tens of kilometres apart. Even for those species able to fast for many months (such as large lysianassids and hagﬁshes (Hargrave et al. 1994, Tamburri & Barry 1999)), the probability that a single drifting or swimming individual would ﬁnd a whale carcass over a period of months must be very low. In fact, a simple calculation can illustrate this point. Collins et al. (1998) estimated that rat-tails, after feeding, move away from a deep-sea baitfall at radial velocities averaging 1.4 km d1. Thus, an average rat-tail would require of the order of 10 days to cover the nearest-neighbour distance of 16 km between gray-whale falls in the mobile scavenger stage. Assuming (a) that rat-tails move outward in random directions (Collins et al. (1998) and (b) that they can detect a whale fall from a range of 500 m, a rat-tail has roughly a 1/100 chance (i.e. two times the detection range divided by the 100 km circumference of a circle of radius 16 km) of ﬁnding a whale carcass 16 km away. To have, on average, a 50% chance of ﬁnding a whale carcass with this nearest-neighbour spacing, a rat-tail would have to make n randomly oriented steps of 16 km, where n can be calculated from the following equation: 0.5 1 (99/100)n Solving this equation gives an n of approximately 70. Since each of the 70 steps of 16 km would require of order 10 days of rat-tail movement, on average, a rat-tail would have a 50% chance of ﬁnding a whale fall roughly every 700 days. With maximum fasting times of 30–660 days (Hargrave et al. 1994, Tamburri & Barry 1999), it thus appears unlikely that large scavengers could specialise on whale falls. It should be noted that our estimates of whale-fall abundances are based on current whale population sizes that, excluding northeast Paciﬁc gray whales, are typically thought to be 10% to 50% of population sizes prior to large-scale whaling operations, that is, prior to the year 1800 (Gaskin 1982, Braham 1984, Braham & Rice 1984, Gosho et al. 1984, Johnson & Wolman 1984, Mizroch et al. 1984a,b,c, Rice et al. 1984, Shelden & Rugh 1995). Before industrial whaling, whale falls at the sea ﬂoor must have been substantially more abundant (Butman et al. 1995); in fact, Jelmert & Oppen-Bernsten (1996) calculate that, prior to whaling, there were 3.9 105 carcasses sinking per year, making whale falls six times more abundant than at present. Thus, the evolution and survival of whale-fall specialists, and the use of whale skeletons as dispersal stepping stones by vent and seep species, would have occurred much more readily prior to the industrial revolution (Butman et al. 1995). In fact, it is quite feasible that the vast diminution in cetacean populations resulting from whaling reduced deep-sea biodiversity by removing organic-rich habitat islands and sulphide-rich dispersal stepping stones at the deep-sea ﬂoor (Butman et al. 1995, 1996; see Jelmert & Oppen-Bernsten 1996, for a contrasting view). Both whale-fall specialists and some more generalised components of reducing-habitat faunas may have been driven to extinction due to massive loss of whale-fall habitats over the past 200 yr (Butman et al. 1995).

336

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Relationships of California whale falls to other modern communities Scavenger assemblages During the mobile-scavenger stage, whale carcasses off California are consumed by a suite of apparently generalised scavengers. The dominant whale-fall scavengers, in particular hagﬁshes (Eptatretus deani and Mixine circifrons), sleeper sharks (Somniosus paciﬁcus), lysianassid amphipods, macrourids, and lithodid crabs, are known to scavenge ﬁsh falls of a broad range of sizes (Dayton & Hessler 1972, Isaacs & Schwartzlose 1975a, Smith 1985). As might be expected, the aggregation sizes for some of these scavengers, in particular hagﬁshes and amphipods, are at least an order of magnitude larger on whale falls of 5000–35 000 kg than on ﬁsh parcels ranging in size from 1 kg to 40 kg (Table 2 and Smith 1985). It should be noted that the mobile-scavenger stage for bathyal California whales may not entirely overlap scavenger assemblages on ﬁsh falls. Sableﬁsh (Anoplopoma ﬁmbria) and the brittle star Ophiophthalmus normani feed actively on ﬁsh falls (Isaacs & Schwatzlose 1975b, Smith 1985) but have not been observed feeding on whale falls, even though these species occur in the San Diego Trough and, possibly, in the Santa Cruz Basin. The absence of these scavengers from the San Diego Trough whale fall may indicate an avoidance of cetacean ﬂesh or, alternatively, of putrifying ﬂesh. Additional experiments are required to determine whether some necrophagous species off California fail to feed on fresh whale carrion. In other ocean basins, cetacean falls also appear to be consumed by generalised scavengers. For example, Jones et al. (1998) found that scavenger assemblages on small cetacean carcasses (53–100 kg ww) in the North Atlantic abyss included macrourids and lysianassid amphipods, and were similar to those on other types of baitfalls. In addition, the facultative scavenging shark, Centroscyllium coelolepis, is commonly taken at bathyal depths in the North Atlantic with whale tissue (including skinless blubber), in its stomach contents, which suggests feeding on whale falls (Nils-Roar Hareida, pers. comm.). Communities on plant and other organic substrata In addition to whale bones, other submerged organic debris, such as sunken wood, seagrass, and algal holdfasts, harbours a specialised fauna (e.g. Turner 1973, 1977, Wolff 1979). The molluscan fauna of these biogenic substrata are best documented, most likely because molluscs often remain attached to their substratum when recovered in trawls. Cocculiniform limpets are diverse on biogenic substrata (Wolff 1979, Haszprunar 1988) and bathymodiolin mussels (Bivalvia: Mytilidae) can be extremely abundant (Wolff 1979). Not surprisingly, these are also the two main groups that overlap with whale-fall habitats. Whale skeletons in the sulphophilic stage have seven species in common with sunken wood; the polychaete Nereis anoculis (Wolff 1979, Baco & Smith, unpubl. data), the limpet Paracocculina cervae, also found on algal holdfasts (Marshall 1994), and the mussels Idas washingtonia, I. argenteus, I. ghisottii, I. (Adipicola) simpsoni, and Adipicola osseocola (Tebble 1966, Dell 1987, 1995, Warén 1991, 1993, Baco & Smith, unpubl. data). A. osseocola is also found on ﬁsh bones (Dell 1996). Further sampling of both whale falls and sunken wood seems very likely to yield additional species overlap. At higher taxonomic levels, the limpet suborder Cocculiniformia is found almost exclusively on biogenic substrata (e.g. Haszprunar 1988). To date there are eight cocculiniform species known from sulphophilic whale skeletons in the genera Cocculina, Paracocculina, Pyropelta and Osteopelta. Xylodiscula is another whale-fall gastropod genus that overlaps with sunken wood and seagrass (Marshall 1994). The bathymodiolin genera Idas and 337

CRAIG R. SMITH & AMY R. BACO

Adipicola also seem to be associated primarily with deep-sea biogenic substrata (e.g. Dell 1987, 1996). Besides the bathymodiolin species mentioned above, several species in each genus are exclusive to either whale bones or sunken wood. Another, non-molluscan, taxon shared between whale falls and wood is the sipunculan genus Phascolosoma. Two deep-sea species are known in this genus, one from wood falls (P. turnerae), and the other from whale bones (P. saprophagicum) (Gibbs 1987). The substantial overlap at higher taxonomic levels between whale falls and other types of sunken organic debris suggests a close evolutionary history for some of their faunal components (see Evolutionary stepping stone section, p. 343). Enrichment opportunists Whale-fall communities, particularly during the enrichmentopportunist stage, share genera and some species with communities associated with other types of organic enrichment in shallow-water and deep-sea settings. The prominence of dorvilleid polychaetes, particularly the genus Ophryotrocha, is a common feature at whale falls, in communities around sewer outfalls in shallow water, as well as in organicallyenriched sediment trays and Sargassum falls in the deep North Atlantic and North Paciﬁc (Pearson & Rosenberg 1978, Desbruyeres et al. 1980, Levin & Smith 1984, Grassle & Morse-Porteous 1987, Levin et al. 1994). In addition, the polynoid polychaete genus Peinaleopolynoe appears to respond both to whale falls off California and to organically enriched sediment trays at depths of 2000 m in the northeast Atlantic (Desbruyeres and Laubier 1988, Pettibone 1993). This genus bears branchiae and has life-history characteristics allowing it to exploit intense habitat islands of organic enrichment (Desbruyeres and Laubier 1988). Cumaceans in the genus Cumella, which were abundant around whale falls in the organicenrichment stage, may occur also in high densities around ﬁsh falls and in enriched sediment trays in the deep sea (Smith 1986, Snelgrove et al. 1994). In addition, organically-enriched sediments underlying salmon pens in Norwegian fjords harbour chrysopetalids morphologically very similar to Vigtorniella n. sp. found in abundance on the California whale falls. Thus, bathyal whale falls off California do appear to foster species potentially belonging to a generalised enrichment fauna. This similarity contrasts with the faunal response to organically-enriched sediments beneath at least some oxygen minimum zones in the bathyal northeast Paciﬁc, which apparently fail to attract generalised enrichment respondents (Levin et al. 1994). Sulphophilic communities – modern vent and seep afﬁnities Smith et al. (1989) hypothesised that whale skeletons might provide important dispersal stepping stones for vent and seep species dependent on sulphide availability at the deep-sea ﬂoor. This hypothesis has been somewhat controversial (Tunnicliffe & Juniper l990, Martill et al. 1991, Squires et al. 1991, Butman et al. 1995, 1996, Tunnicliffe & Fowler 1996, Jelmert & Oppen-Bernsten 1996) and could be rejected if no overlap were found between the faunas of whale falls, hydrothermal vents, and cold seeps. In fact, a number of species are shared among whale falls and vents or seeps, with some of these being abundant in both habitats. Thus far, 11 macrofaunal/megafaunal species are known to be shared between whale-falls and hydrothermal vents, with the main vent overlap coming from soft-sedimented vent sites in Guaymas Basin, and Middle Valley on the Juan de Fuca Ridge (Table 6). In addition, 20 species are known to occur at both whale-falls and cold seeps (Table 6, Warén & Bouchet 2001). This is a small percentage (2–10%) of the species found in any of these habitats, indicating that, at the species level, the whale-fall, vent and seep biotas are largely distinct. It is, 338

Vesicomya gigas Calyptogena kilmeri Calyptogena elongata Calyptogena paciﬁca? Idas washingtonia Pyropelta corymba Pyropelta musaica Cocculina craigsmithi Neoleptopsis sp? Astyris permodesta Provanna lomana Eulimella lomana Ilyarachna profunda Janiridae sp. Bathykurila guaymasensis Syllid sp. A Sabellid sp. C Maldanid sp. C Dorvilleid sp. Harmothoe craigsmithi Escarpia spicata Entoproct sp. B Octocoral sp. A

Bivalves

339

Totals

Vestimentiferans Entoprocts Cnidarians

Polychaetes

Isopods

Gastropods

Species

Major taxon

Vents 10*

Guaymas vents

Whale skeletons

Habitat Juan de Fuca vents

Seeps 20

Northeast Paciﬁc seeps

Gulf of Mexico seeps

Table 6 Species overlap among whale falls from southern California and hydrothermal vents at Guaymas Basin and on the Juan de Fuca Ridge, Northeast Paciﬁc seeps at various locations, and Gulf of Mexico seeps. For details see: Bennet et al. 1994, Baco et al. 1999, Baco-Taylor 2002, Smith et al. 2002, Baco & Smith, unpubl. data, and Poehls et al., unpubl. data). One cross species present in the habitat; two crosses species abundant in the habitat. *In the North Atlantic, the limpet Protolira thorvaldssoni was described from whale bones and is common at vents (Warén & Bouchet 2001).

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

CRAIG R. SMITH & AMY R. BACO

however, important to note that many whale-fall species (most likely dozens) remain to be identiﬁed. In addition, only one seep in proximity to southern California whale falls, the San Clemente Basin seep, has been sampled for macrofauna (Poehls et al., in prep.), and this site alone shares 12 species with the California-slope whale falls. It can be expected that further sampling of whale skeletons in the proximity of vents and seeps, and increased identiﬁcation of whale-fall species, will likely increase the known species overlap among these habitats. Nonetheless, only a subset of vent and seep species are likely ever to be found on whale falls and potentially use them as dispersal stepping stones. In addition to the species-level overlap, these reducing habitats also share a number of genera. The limpet genus Pyropelta includes two whale-fall species, P. corymba and P. musaica which occur at vents but also P. wakeﬁeldi, which has been found only on whale falls (McLean & Haozprunar 1987, McLean 1992). Two snail genera, Bruciella and Xylodiscula, which were described from vents, have representatives in whale-fall habitats (Marshall 1994). Also, several unidentiﬁed whale-fall species have been preliminarily placed into genera that are known from vents or seeps. For example, “Snail sp. J” from the Santa Cruz skeleton and sediments at 1.5 yr is likely to be a new species of Hyalogyrina (Hyalogyrinidae) (J. McLean, pers. comm.), a genus reported by Warén & Bouchet (1993, 2001) from seep habitats. At least two vescomyid genera, Calyptogena and Vesicomya, also appear to be shared among whale-fall, vent and seep habitats (Baco et al. 1999). Sulphophilic whale-fall communities appear to differ from other reducing habitats in the apportionment of macrofaunal species among phyla. Based on worldwide species lists, Mollusca and Arthropoda are the most speciose phyla at vents (Tunnicliffe et al. 1998), and Mollusca the most species-rich at seeps (Sibuet & Olu 1998). In contrast, annelids account for 47–60% of macrofaunal species in all whale-fall successional stages at all ﬁve whale falls intensively studied to date (Baco-Taylor 2002, Baco & Smith, unpubl. data). Deep-sea hard substratum biota Whale bones appear to share few species with nonreducing, deep-sea hard substrata. Of the 26 macrofaunal species collected on rocks near the San Nicolas skeleton, only two species, an unidentiﬁed amphipod and a scale worm, were also present on the San Nicolas skeleton (Baco-Taylor 2002). There is also very little overlap between sponge stalks collected off Southern California and the Southern California whale falls (Beaulieu 2001, Baco-Taylor 2002, Baco & Smith, unpubl. data). Many species remain to be identiﬁed from both sponges and whale falls, however, raising the possibility that more overlap will be found. As discussed above, whale bones appear to harbour the highest diversity of any deep-sea hard substratum. Densities of macrofaunal individuals on whale skeletons in the sulphophilic stage can also be relatively high, with macrofaunal densities reaching 22 000 ind. m2 (Baco-Taylor 2002, Baco & Smith, unpubl. data). Other deep-sea hard substratum habitats such as manganese nodules and sponge stalks had densities of macrofauna and meiofauna combined of ⬃11 000 ind. m2 (Beaulieu 2001, Mullineaux 1987). Densities on the Santa Catalina Basin (SCB) skeleton in 1999 (⬃22 000 ind. m2) were also much greater than in background SCB sediments (7000 ind. m2, Smith et al. 1998). Whale fall specialists There is substantial evidence that deep-sea whale falls harbour a specialised fauna (i.e. one that is speciﬁcally adapted to live on whale remains). At least 28 macrofaunal species were ﬁrst collected on whale falls, and 21 of these have not been found in any other habitat (Table 7). A number of the species thus far unique to whale carcasses 340

341

Dorvilleidae* Sipuncula

Chrysopetalidae Ampharetidae

Vesicomyid Thyasiridae Aplacophora Arthropoda Anomura Annelida Polychaeta Polynoidae

Bivalvia Bathymodiolinae

Gastropoda

Mollusca Archaegastropoda

Higher taxon

X

X X X X X X X X

Paralomis manningi

Harmathoe craigsmithi Peinaleopolynoe santacatalina Vigtorniella n. sp. Ampharetid gen nov. Asabellides sp. nov. Anobothrus sp. nov. Palpiphitime sp. nov. Phascolosoma saprophagicum

X X X X

X X

X X X

California

South Atlantic Japan, Hawaii New Zealand New Zealand North Atlantic North Atlantic California California California

California California New Zealand New Zealand Iceland New Zealand Iceland New Zealand New Zealand New Zealand California

Location

California California 1000–100 000 California 10 California 10 California California 10 000 California 20–200 New Zealand

200

X X

100 300–1100

Estimated pop. size

X

Known only at whale falls

Adipicola pelagica Myrina (Adipicola) paciﬁca Adipicola (Idas) arcuatilis Adipicola osseocola Idas pelagica Idas ghisottii New species? Axinodon sp. nov. New genus

Pyropelta wakeﬁeldi Cocculina craigsmithi Paracocculina cervae Osteopelta praeceps Osteopelta ceticola Osteopelta mirabilis Protolira thorvaldsoni Bruciella laevigata Bruciella pruinosa Xylodiscula osteophila Hyalogyrina n.sp.

Species

Pettibone 1993 Pettibone 1993 Smith et al. 2002, Dahlgren et al., in prep. B. Hilbig, pers. comm. B. Hilbig, pers. comm. B. Hilbig, pers. comm. B. Hilbig, pers. comm. Gibbs 1987

Williams et al. 2000

Dell 1987 Dell 1987 Dell 1996 Dell 1996 Warén 1993 Warén 1993 Baco et al. 1999 P. Scott, pers. comm. Scheltema, in prep.

McLean 1992 McLean 1992 Marshall 1994 Marshall 1994 Warén 1989 Marshall 1987 Warén 1996 Marshall 1994 Marshall 1994 Marshall 1994 McLean & Warén, pers. comm.

Reference

Table 7 Species (n 29) ﬁrst found at whale falls. The 21 species marked as “known only at whale falls” have been found in no other habitat. Where available, estimated population sizes on whale falls are given. *In addition to Palpiphitime sp. nov., at least 45 unidentiﬁed species of dorvilleids, with population sizes ranging from 10’s to 1000’s of individuals per whale fall, have been collected from whale falls in the Santa Catalina Basin, San Diego Trough, San Clemente Basin and Santa Cruz Basin (Baco & Smith, unpubl. data). Many of these species are likely to be new to science.

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

CRAIG R. SMITH & AMY R. BACO

Table 8 Species overwhelmingly more abundant on whale skeletons than in any of their other known habitats. Estimated population sizes on whale skeletons, and the total number of specimens collected in other habitats, are indicated for each species. Data from Bennett et al. (1994), Smith et al. (1998), Baco-Taylor (2002), Smith et al. (2002), Baco & Smith (unpubl. data), Poehls et al. (in prep.) and McLean (pers. comm.). Species Bivalvia Idas washingtonia Gastropoda Cocculina craigsmithi Pyropelta corymba Pyropelta musaica Crustacea Ilyarachna profunda

Population size on whale skeletons

Number collected in other habitat(s)

10 000–20 000

1–10 (wood, vents, seeps)

300–1100 1000–1200 250

1–10 (vents) 1–10 (vents) 1–10 (vents)

500–1800

1–90 (sediments, seeps)

are very abundant, indicating that they are well adapted to whale falls and can attain large population sizes given suitable conditions. Their absence in samples from other related habitats (e.g. wood falls, algal falls, enriched sediment trays, hydrothermal vents and cold seeps) suggests that these species may indeed be endemic to whale falls. In addition to the 21 potential whale-fall endemics, there are at least ﬁve species which attain very high densities on whale falls, and yet appear to occur only as isolated individuals in other habitats (Table 8). It is quite feasible that a large proportion of the total individuals within these species occur in the whale-fall habitat, essentially making them whale-fall specialists (e.g. with their evolution largely shaped by the selective milieu of whale falls). This brings the total number of potential whale-fall specialists to 26. It should be noted that this number will surely rise as the diverse dorvilleid (45 species), amphipod, and copepod components of the whale-fall fauna are rigorously examined by taxonomists. The taxonomic and functional diversity of the potential whale-fall specialists is noteworthy. These “specialists” come from ﬁve different phyla, and appear to include whale-bone feeders (the sipunculid and some limpets), bacterial grazers (some limpets, Ilyarachna profunda), species utilising chemoautotrophic endosymbionts (the bathymodiolins, thyasirid and vesicomyid), deposit feeders (the ampharetids), facultative suspensions feeders (the bathymodiolins), and predators (the polynoids and Paralomis manningi) (see discussion of food webs above). This diversity, in combination with the abundance patterns of macrofaunal species on whale skeletons (Bennett et al. 1994), suggest that a variety of taxa and trophic types may have become speciﬁcally adapted to whale-fall niches.

Ancient/evolutionary relationships Ancient whales and reptiles Large cetaceans have existed for 40 myr (million years) (Briggs & Crowther 1990). Because ancient oceans contained scavengers, decomposers, and molluscs with chemoautotrophic endosymbionts functionally similar to those in the modern ocean (Hogler 1994), it 342

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

seems very likely that whale-fall succession has generally followed the patterns we describe above for tens of millions of years (cf. Hogler 1994). Fossil chemoautotrophic communities have been found on fossil whale skeletons as old as 30 myr (Squires et al. 1991, Goedert et al. 1995). During the Mesozoic, before the existence of whales, it is likely that large marine reptiles, particularly ichthyosaurs and plesiosaurs, supported chemoautotrophy-based communities (Martill et al. 1991, Hogler 1994, Marshall 1994). Squires et al. (1991) and Goedert et al. (1995) provide fossil evidence of chemosynthetic communities associated with a variety of deep-sea whale skeletons as early as the Oligocene (30 mya). Eight whale skeletons in the Makah and Pysht formations on the Olympic Peninsula (Washington State) harboured a molluscan fauna characteristic of reducing habitats, including mytilid, thyasirid, and lucinid bivalves; modern representatives of these families are known to harbour chemoautotrophic endosymbionts. Based on these ﬁndings, Goedert et al. (1995) estimate that whale skeletons have been able to support chemoautotrophic communities for at least the past 30–35 myr. Martill et al. (1991) suggest that other large marine vertebrates, such as tetrapods and marine reptiles, may have supported chemoautotrophic fauna as early as 200 mya. From ichthyosaur and plesiosaur remains, they found evidence of molluscs that are also associated with Eocene seeps. Marshall (1994) found a fossilised limpet, Osteopelta cf. mirabilis, in close association with bones of a fossil leatherback turtle from the Middle Eocene. Similar limpets are also known from modern whale falls in New Zealand and Iceland (Marshall 1987, Warén 1989). Kitazato & Shirayama’s (1996) experiment with cow bones also showed that bones of other vertebrates are capable of supporting chemoautotrophic production.

Dispersal stepping stones in ecological and evolutionary time Smith et al. (1989) hypothesised that whale skeletons might provide important dispersal stepping stones for species (e.g. some from hydrothermal vents and cold seeps) dependent on sulphide availability at the deep-sea ﬂoor. When initially posed, this hypothesis was controversial (e.g. Tunnicliffe & Juniper 1990, Goedert et al. 1995), although the data were clearly inadequate to provide a deﬁnitive test. It now appears reasonable that at least a few taxa may have used whale falls for dispersal among reducing habitats in ecological and evolutionary time. One group of species which may have used whale falls as dispersal stepping stones are the vesicomyid clams. Using mitochondrial COI DNA sequences, three to four species of vesicomyid clams have been identiﬁed on whale falls (Baco et al. 1999). These clams were conspeciﬁc with (a) Vesicomya gigas, a species collected from northeast Paciﬁc vent sites, (b) Calyptogena kilmeri, a species found at northeast Paciﬁc cold seeps, and (c) Calyptogena elongata, a species found in anoxic California basins (Baco et al. 1999). Because the whalefall clams are conspeciﬁc with vent and seep species, and because they occur in reproductively viable population sizes at whale falls, these results offer support for the dispersal stepping-stone hypothesis. Baco et al. (1999) also showed that whale falls may have played a role in the evolution of vesicomyid clams. Peek et al. (1997) suggested that most vesicomyid lineages are restricted to a single type of reducing habitat (i.e. vents, seeps or anoxic basins). However, whale-fall vesicomyids deviate from this pattern, containing vesicomyid lineages found also at 343

CRAIG R. SMITH & AMY R. BACO

soft-sediment hydrothermal vents, cold seeps and anoxic basins. This suggests that whale falls may offer habitat conditions intermediate to, or broader than, those found in other reducing habitats. Whale falls may well represent an intermediate habitat type between softsediment vents and seeps, with the potential to provide evolutionary stepping-stones between divergent soft-sediment reducing habitats at the deep-sea ﬂoor (Baco et al. 1999). Based on very rough estimates, the diversiﬁcation of vesicomyid clams was approximately synchronous with the diversiﬁcation of large cetaceans, suggesting that the relationship between whale and vesicomyid evolution merits further scrutiny (Baco et al. 1999). Whale falls may also have been important in the evolution of vent-seep mytilids. The evolutionary origins of hydrothermal-vent and cold-seep species have been the subject of speculation. Many vent species are thought to have evolved from seep ancestors, with evolution progressing from shallow water to the deep sea (Hecker 1985, McLean 1985, Craddock et al. 1995). Until very recently, little attention had been given to the potential importance of organic-remain habitats (i.e. whale falls, wood falls, algal falls) in the evolution of vent-seep faunas. By studying DNA sequences of the nuclear 18S gene in mytilids from a range of deep-sea reducing habitats including hydrothermal vents, cold seeps, whale falls and wood islands, Distel et al. (2000) showed that whale fall and wood mussels in the genera Idas, Adipicola and Benthomodiolus were closely related to vent and seep mussels in the genera Tamu and Bathymodiolus (Distel et al. 2000). Baco et al. (Baco-Taylor 2002, Baco et al., unpubl. data) then used Mitochondrial 16S and COI DNA gene sequences to demonstrate an evolutionary sequence from sunken wood to whale falls to seeps and ﬁnally to vents, suggesting organic-remains mytilids preceded vent and seep mytilids in evolutionary time (Baco-Taylor 2002, Baco et al., unpubl. data). All three genes revealed that the organicremains, vent, and seep mytilids form a monophyletic subfamily that evolved 30 mya from a shallow water ancestor (Distel et al. 2000, Baco-Taylor 2002, Baco et al., unpubl. data), consistent with the estimated diversiﬁcation times for vesicomyid clams and large whales (see above). Baco et al. (unpubl. data) also used carbon isotopic data combined with the mitochondrial DNA phylogenies to yield insights into the evolutionary history of mytilid-endosymbiont associations (Baco-Taylor 2002, Baco et al., unpubl. data). Many vent and seep mytilids are known to harbour sulphur-oxidising and/or methanotrophic endosymbionts. Based on 13C values, Baco et al. provided evidence that species on organic remains exhibited an increasing dependence on sulphur-oxidising chemoautotrophy over evolutionary time (Baco-Taylor 2002, Baco et al., unpubl. data). Stable isotope data also suggest that the mytilid-endosymbiont relationship evolved in organic remains-habitats, rather than in vent and seep environments (Baco-Taylor 2002, Baco et al., unpubl. data). All of these results provide strong support for the hypothesis that organic remains, including whale falls, have provided evolutionary stepping-stones as mytilids have radiated from shallow water into deep-sea vent and seep habitats.

Biotechnological spinoffs When a whale carcass arrives at the deep-sea ﬂoor, a diverse assemblage of microbes colonises and decomposes the lipids and proteins contained in the remains (e.g. Allison et al. 1991, Deming et al. 1997). Because deep-sea habitats generally are cold (2–4°C), the bacter344

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

ial decomposers on whale falls are typically psychrophilic (i.e. they have optimal growth temperatures below 20°C) or psychrotrophic (i.e. are facultative psychrophiles). The enzymes of psychrotrophic bacteria are of particular commercial interest because they sustain high activities at low temperatures and yet remain relatively stable at high temperatures (J. Stein, pers. comm.). Enzymes with these characteristics are desirable in the detergent, pharmaceutical and food-processing industries. This potential for discovering psychrotrophs has led to the exploration of lipid-rich whale-fall habitats for novel bacterial enzymes (e.g. lipases and proteases) for use in cold-water detergents and other industrial applications. Through use of recombinant cloning techniques, the biotechnology company Diversa, Inc. identiﬁed a large number of bacterial clones from whale carcasses with cold-adapted lipase activity. This approach allows direct access to the genomic information of natural microbial assemblages, in which 99% of the diversity remains unculturable. Some of the whale-carcass lipases appear to have promise as detergent additives, potentially allowing stains to be removed more efﬁciently from laundry during cold-water washing. The successful application of such enzyme to detergents could yield signiﬁcant energy savings and prove proﬁtable; the USA demand for detergent enzymes currently supports a market estimated at roughly $150 million yr1 (J. Stein, pers. comm.).

Anthropogenic inﬂuences on whale-fall communities The populations sizes of large cetaceans have suffered major depredations from human whaling activities over the last 200 yr. In particular, the abundance of all the great whale species were drastically reduced, and some species exterminated (e.g. the North Atlantic gray whale), between 1860 and 1986 (Butman et al. 1995). Clearly, whaling has dramatically altered the rates and geographic distribution of whale falls to the deep-sea ﬂoor (Butman et al. 1995, 1996). Because whale falls harbour a specialised fauna and may provide dispersal stepping stones for some deep-sea sulphophiles, this reduction in whale falls may have caused species extinctions, and reduced species diversity, in deep-sea ecosystems ranging from whale falls to hydrothermal vents (Butman et al. 1995, 1996). Those species most dependent on whale falls are the most likely to have been exterminated, raising the possibility that whale-carcass habitats now retain only the most generalised subset of their original biota. Unfortunately, the structure of whale-fall communities, and assemblages in other deep-reducing habitats such as vents and seeps, has been studied only very recently, with data collection initiated in 1977 (Van Dover 2000). Thus, it will be very difﬁcult to evaluate the biodiversity losses in whale-fall communities, and other deep-sea habitats, caused by intensive whaling. Some insights into the effects of ﬂuctuating whale-carcass supply may be gained by studying whale-fall ecology and biogeography as global whale populations rebound from their hunting-induced lows (Butman et al. 1995). However, even such studies will fail to elucidate the identity and characteristics of species driven to extinction as an indirect consequence of whaling. This sobering thought highlights the need to explore the remote, poorly known ecosystems of the deep ocean prior to the further anthropogenic alteration of marine ecosystems (e.g. due to pollution, overﬁshing, and most signiﬁcantly, global climate change) if we wish to reveal (and preserve) the ecological and evolutionary wonders of the deep sea. 345

CRAIG R. SMITH & AMY R. BACO

Future directions The last 15 yr have witnessed dramatic advances in our understanding of the ecology of whale falls. Nonetheless large gaps in our knowledge remain. A few research areas that could yield dramatic progress are highlighted below.

Microbial community structure and dynamics Deep-sea whale bones and surrounding organically-enriched sediments are extreme environments in terms of organic loading, electron-acceptor availability, low temperature, and high hydrostatic pressure. In addition, whale falls are ephemeral, eutrophic habitat islands embedded in a generally oligotrophic sea ﬂoor. Such conditions may select for novel microbial metabolic strategies, dynamics, consortia and symbioses within the lipid-rich bone matrix, on bone surfaces, within the tissues of Metazoa, and in surrounding impacted sediments. While limited bio-prospecting for novel microbial enzymes has occurred in the whale-fall habitat (see above), virtually nothing is known about microbial biodiversity or the dynamics of microbially mediated biogeochemical transformations in deep-sea whale falls. In addition, the nature of microbial symbioses in bathymodiolin mussels on whale falls, and in organic-remains habitats is poorly understood. Because whale- and wood-fall bathymodiolins show evidence of increasing reliance on chemoautotrophic production (Baco et al. unpubl. data), studies of the mussels may provide insights into the evolution of chemoautotrophic endosymbiosis.

Macrofaunal reproduction, dispersal and gene ﬂow Because of the fragmented, relatively ephemeral nature of whale-fall habitats, whale-fall specialists are likely to exhibit reproduction and dispersal strategies atypical for the general deep sea, but potentially similar to those from hydrothermal vents and cold seeps (Van Dover 2000). Reproductive and dispersal strategies for whale-fall biota remain largely unknown, as do rates of gene ﬂow among whale falls, and between whale falls and other types of reducing habitats. Settling cues for whale-fall specialists may be particularly unusual by deep-sea standards, and might include compounds characteristic of putrefaction, such as the diamines putrescine and cadaverine (Hart & Schuetz 1972).

Succession Many issues concerning the structure and dynamics of whale-succession remain unresolved. How long can the sulphophilic stage last? Is faunal succession functionally and taxonomically similar on sunken whale carcasses in regions beyond the California slope, on the carcasses of other large invertebrates (e.g. whale sharks), or on other large concentrations of labile organic matter at the deep-sea ﬂoor (e.g. packages of sewage sludge, boluses of trawl discard)? How far back in the fossil record can such patterns of succession be documented for large carcasses (e.g. ichthyosaurs and plesiosaurs; see Hogler 1994 for speculation)? The answers to such questions are essential to understanding the dynamics of whale-fall habitat 346

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

islands, the recycling of large parcels of organic matter, and the evolution of sulphophiles and opportunistic species at the deep-sea ﬂoor.

Relationships between whale-fall, kelp-fall and wood-fall communities While the communities associated with plant debris have been documented in many parts of the deep sea (e.g. Wolff 1976), faunal assemblages associated with large kelp falls and wood falls remain largely unstudied in the deep northeast Paciﬁc, even though kelp falls may be common (e.g. Smith 1983, Harrold et al. 1998). Because large plant falls may provide concentrated and persistent sources of organic enrichment and reduced inorganic species (e.g. sulphide and methane; Smith 1983, Vetter 1994, Distel et al. 2000), they may foster assemblages closely related to the whale fall biota. In fact, it is quite conceivable that some of the species now regarded as potential whale-fall specialists utilise large kelp or wood falls as their primary habitat.

Biogeography and evolution of whale-fall communities The structure of whale-fall assemblages on the California slope is reasonably well known but the sampling of whale-fall communities in other oceanic regions is extremely fragmentary. Thus far, several whale-fall species are known to be widespread within ocean basins (e.g. Vigtorniella n. sp. on Californian and Hawaiian bones (Dahlgren et al., unpubl. data), Adipicola pelagica in the North and South Atlantic (Dell 1987)) but it is impossible to say whether pan-basin distributions are the rule or exception. We are even further from describing the biogeographic provinces of the whale-fall biota, and how their structure is related to the distribution of whale feeding grounds and migration corridors, and to the documented biogeographic patterns of hydrothermal vents and cold seeps (Van Dover et al. 2002). Knowledge of these biogeographic patterns is essential to rigorous evaluation of the evolutionary history of the whale-fall, vent and seep biotas. Rather than speculate on these patterns, we look forward to more widespread studies of the ecology and biogeography of whale falls and other reducing habitats within the framework of such programmes as the Census of Marine Life’s Chemosynthetic Ecosystems Project (ChEss; Tyler et al., in press).

Acknowledgements This paper is dedicated to the late Jacqueline R. Smith, whose love, guidance, support and enthusiasm launched the senior author on a career in science. We warmly thank the people too numerous to name who assisted at sea and on land during our whale-fall studies. We also thank Nils-Roar Haeida for allowing us to cite unpublished data. Our own whale-fall studies have beneﬁted immensely from the extraordinary talents and efforts of the crews of the DSRV ALVIN, the ATV, and DSV TURTLE. Our whale-fall work has been generously supported by the National Undersea Research Center Alaska (now the West Coast & Polar Regions Undersea Research Center), the U.S. National Science Foundation, the National Geographic Society, the British Broadcasting Corporation, the U.S. E.P.A., and the University of Hawaii 347

CRAIG R. SMITH & AMY R. BACO

Research Council. Adrian Glover and Iris Stowe assisted heroically in manuscript preparation. This is contribution no. 6088 from the School of Ocean and Earth Science and Technology, University of Hawaii at Manoa.

References Allison, P. A., Smith, C. R., Kukert, H., Deming, J. W. & Bennett, B. A. 1991. Deep-water taphonomy of vertebrate carcasses: a whale skeleton in the bathyal Santa Catalina Basin. Paleobiology 17, 78–89. Anderson, G. S. 2001. Insect succession on carrion and its relationship to determining time of death. In Forensic entomology, J. H. Byrd & J. L. Castner (eds). Boca Raton: CRC Press, 143–176. Ashley, C. W. 1926. The Yankee whaler. Boston: Houghton Mifﬂin. Baco, A. R. & Smith, C. R. (in review). High biodiversity levels on a deep-sea whale skeleton. Marine Ecology Progress Series. Baco, A. R., Smith, C. R., Peek, A. S., Roderick, G. K. & Vrijenhoek, R. C. 1999. The phylogenetic relationships of whale-fall vesicomyid clams based on mitochondrial COI DNA sequences. Marine Ecology Progress Series 182, 137–147. Baco, A. R., Smith, C. R. & Vrijenhoek, R. C. 1996. Deep-sea whale skeleton communities on the California slope: structure, dynamics and vent-seep afﬁnities. EOS, Transactions of the American Geophysical Union 76, OS68 only. Baco-Taylor, A. R. 2002. Food-web structure, succession and phylogenetics on deep-sea whale skeletons. PhD thesis, University of Hawaii. Beaulieu, S. E. 2001. Life on glass houses: sponge stalk communities in the deep sea. Marine Biology 138, 803–817. Bennett, B. A., Smith, C. R., Glaser, B. & Maybaum, H. L. 1994. Faunal community structure of a chemoautotrophic assemblage on whale bones in the deep northeast Paciﬁc Ocean. Marine Ecology Progress Series 108, 205–223. Black, M. B., Lutz, R. A. & Vrijenhoek, R. C. 1994. Gene ﬂow among vestimentiferan tube worm (Riftia pachyptila) populations from hydrothermal vents of the eastern Paciﬁc. Marine Biology 120, 33–39. Black, M. B., Trivedi, A., Maas, P. A. Y., Lutz, R. A. & Vrijenhoek, R. C. 1998. Population genetics and biogeography of vestimentiferan tube worms. Deep-Sea Research II 45, 365–382. Braham, H. 1984. The status of endangered whales: an overview. Marine Fisheries Review 46, 2–6. Braham, H. & Rice, D. 1984. The right whale, Balaena glacialis. Marine Fisheries Review 46, 38–44. Briggs, D. E. G. & Crowther, P. R. 1990. Palaeobiology: a synthesis. Oxford: Blackwell Scientiﬁc Publications. Britton, J. C. & Morton, B. 1994. Marine carrion and scavengers. Oceanography and Marine Biology: an Annual Review 32, 369–434. Bruun, A. F. 1956. The abyssal fauna: its ecology, distribution and origin. Nature 177, 1105–1108. Butman, C. A., Carlton, J. T. & Palumbi, S. R. 1995. Whaling effects on deep-sea biodiversity. Conservation Biology 9, 462–464. Butman, C. A., Carlton, J. T. & Palumbi, S. R. 1996. Whales don’t fall like snow: reply to Jelmert. Conservation Biology 10, 655–657. Byrd, J. H. & Castner, J. L. 2001. Forensic entomology; the utility of arthropods in legal investigations. Boca Raton: CRS Press. Collins, M. A., Priede, I. G., Addison, S., Smith, A. & Bagley, P. M. 1998. Acoustic tracking of the dispersal of organic matter by scavenging ﬁshes in the deep-sea. Hydrobiologia 371/372, 181–186. Committee on Biological Diversity in Marine Systems 1995. Understanding marine biodiversity: a research agenda for the nation. Washington, DC: National Academy Press.

348

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Connell, J. H. & Slatyer, R. O. 1977. Mechanisms of succession in natural communities and their role in community stability and organization. American Naturalist 111, 1119–1144. Craddock, C., Hoeh, W. R., Gustafson, R. G., Lutz, R. A., Hashimoto, J. & Vrijenhoek, R. C. 1995. Evolutionary relationships among deep-sea mussels (Bivalvia: Mytilidae) from hydrothermal vents and cold-water methane/sulﬁde seeps. Marine Biology 121, 477–485. Dayton, P. K. & Hessler, R. R. 1972. Role of biological disturbance in maintaining diversity in the deep sea. Deep-Sea Research 19, 199–208. Dell, R. K. 1987. Mollusca of the Family Mytilidae (Bivalvia) associated with organic remains from deep water off New Zealand, with revisions of the genera Adipicola Dautzenborg, 1927 and Idasola Iredale, 1915. National Museum of New Zealand Records 3, 17–36. Dell, R. K. 1995. New species and records of deep-water mollusca from off New Zealand. National Museum of New Zealand Records 2, 1–26. Deming, J., Reysenbach, A. L., Macko, S. A. & Smith, C. R. 1997. The microbial diversity at a whale fall on the seaﬂoor: bone-colonizing mats and animal-associated symbionts. Microscopy Research and Technique 37, 162–170. DeNiro, M. J. & Epstein, S. 1978. Inﬂuence of diet on the distribution of carbon isotopes in animals. Geochimica et Cosmochimica Acta 42, 495–506. DeNiro, M. J. & Epstein, S. 1981. Inﬂuence of diet on the distribution of nitrogen isotopes in animals. Geochimica et Cosmochimica Acta 45, 341–351. Desbruyeres, D., Bervas, J. Y. & Khripounoff, A. 1980. Un cas de colonization rapide d’un sediment profound. Oceanologica Acta 3, 285–291. Desbruyeres, D. & Laubier, L. 1988. Exploitation d’une source de matière organique concentrée dans l’océan profound: intervention d’une annelide polychete nouvelle. Comptes Rendus de L’Academie des Sciences Série III 307, 329–335. Distel, D. L., Baco, A. R., Chuang, E., Morril, W., Cavanaugh, C. & Smith, C. R. 2000. Do mussels take wooden steps to deep-sea vents? Nature 403, 725–726. Etter, R. J. & Caswell, H. 1994. The advantages of dispersal in a patchy environment: effects of disturbance in a cellular automaton model. In Reproduction, larval biology, and recruitment of the deepsea benthos, C. Young & K. Eckelbarger (eds). New York: Columbia University Press, 284–305. Feldman, R. A., Shank, T. M., Black, M. B., Baco, A. R., Smith, C. R. & Vrijenhoek, R. C. 1998. Vestimentiferan on a whale fall. Biological Bulletin 194, 116–119. Fisher, C. R., Childress, J. J., Macko, S. A. & Brooks, J. M. 1994. Nutritional interactions in Galapagos Rift hydrothermal communities: inferences from stable carbon and nitrogen analyses. Marine Ecology Progress Series 103, 45–55. Fry, B. & Sherr, E. B. 1984. 13C measurements as indicators of carbon ﬂow in marine and freshwater ecosystems. Contributions in Marine Science 27, 13–47. Fujioka, K., Wada, H. & Okano, H. 1993. Torishima whale deep-sea animal community assemblagenew ﬁndings by “Shinkai 6500”. Journal of Geography 102, 507–517. Gage, J. D. & Tyler, P. A. 1991. Deep-sea biology: a natural history of organisms at the deep-sea ﬂoor. Cambridge: Cambridge University Press. Gaskin, D. E. 1982. Ecology of whales and dolphins. Portsmouth: Heinemann Educational. Gibbs, P. E. 1987. A new species of Phascolosoma (Sipuncula) associated with a decaying whale’s skull trawled at 880 m depth in the southwest Paciﬁc. New Zealand Journal of Zoology 14, 135–137. Goedert, J. L., Squires, R. L. & Barnes, L. G. 1995. Paleoecology of whale-fall habitats from deepwater Oligocene rocks, Olympic Peninsula, Washington State. Palaeogeography, Palaeoclimatology, Palaeoecology 118, 151–158. Gosho, M., Rice, D. & Breiwick, J. 1984. The sperm whale, Physeter macrocephalus. Marine Fisheries Review 46, 54–64. Grassle, J. F. & Morse-Porteous, L. S. 1987. Macrofaunal colonization of disturbed deep-sea environments and the structure of deep-sea benthic communities. Deep-Sea Research I 34, 1911–1950.

349

CRAIG R. SMITH & AMY R. BACO

Green, G., Mattson, D. J. & Peek, J. M. 1997. Spring feeding on ungulate carcasses by grizzly bears in Yellowstone National Park. Journal of Wildlife Management 61, 140–155. Hargrave, B. T., Prouse, N. J., Phillips, G. A. & Cranford, P. J. 1994. Meal size and sustenance time in the deep-sea amphipod Eurythenes gryllus collected from the Arctic Ocean. Deep-Sea Research I 41, 1489–1508. Harrold, C., Light, K. & Lisin, S. 1998. Organic enrichment of continental shelf and deep-sea benthic communities by macroalgal drift imported from nearshore kelp forests. Limnology and Oceanography 43, 669–678. Hart, H. & Schuetz, R. D. 1972. Organic chemistry: a short course. Boston: Houghton Mifﬂin. Haszprunar, G. 1988. Anatomy and relationships of the bone-feeding limpets Cocculinella minutissima (Smith) and Osteopelta mirabilis Marshall (Archaeogastropoda). Journal of Molluscan Studies 54, 1–20. Hecker, B. 1985. Fauna from a cold sulfur seep in the Gulf of Mexico: comparison with hydrothermal vent communities and evolutionary implications. Bulletin of the Biological Society of Washington 6, 465–473. Hessler, R. R., Ingram, C. L., Yayanos, A. A. & Burnett, B. R. 1978. Scavenging amphipods from the ﬂoor of the Philippine Trench. Deep-Sea Research I 25, 1029–1047. Hewson, R. 1984. Scavenging and predation upon sheep and lambs in West Scotland, UK. Journal of Applied Ecology 21, 843–868. Hewson, R. 1995. Use of salmonid carcasses by vertebrate scavengers. Journal of Zoology, London 235, 53–65. Hogler, J. A. 1994. Speculations on the role of marine reptile deadfalls in Mesozoic deep-sea paleoecology. Palaios 9, 42–47. Houston, D. C. 1986. Scavenging efﬁciency of turkey vultures in tropical forest. The Condor 88, 318–323. Isaacs, J. D. & Schwartzlose, R. A. 1975a. Active animals of the deep-sea ﬂoor. Scientiﬁc American 233, 85–91. Isaacs, J. D. & Schwartzlose, R. A. 1975b. Biological applications of underwater photography. Oceanus 18, 24–30. Jelmert, A. & Oppen-Bernsten, D. O. 1996. Whaling and deep-sea biodiversity. Conservation Biology 10, 653–654. Johnson, J. & Wolman, A. 1984. The humpback whale, Megaptera novaeangliae. Marine Fisheries Review 46, 30–37. Jones, E. G., Collins, M. A., Bagley, P. M., Addison, S. & Priede, I. G. 1998. The fate of cetacean carcasses in the deep sea: observations on consumption rates and succession of scavenging species in the abyssal northeast Atlantic Ocean. Proceedings of the Royal Society of London Series BBiological Sciences 265, 1119–1127. Jones, M. L., Swartz, S. L. & Leatherwood, S. 1984. The gray whale Eschrictius robustus. San Diego: Academic Press. Jumars, P. A. & Gallagher, E. D.1982. Deep-sea community structure: three plays on the benthic proscenium. In The environment of the deep sea, W. G. Ernst & J. G. Morin (eds). Englewood Cliffs: Prentice-Hall, 217–285. Katona, S. & Whitehead, H. 1988. Are Cetacea ecologically important? Oceanography and Marine Biology: an Annual Review 26, 553–568. Kitazato, H. & Shirayama, Y. 1996. Rapid creation of a reduced environment and an early stage of a chemosynthetic community on cattle bones at the deep-sea bottom in Sagami Bay, central Japan. Vie Milieu 46, 1–5. Klages, M., Muyakshin, S., Soltwedel, T. & Arntz, W. E. 2002. Mechanoreception, a possible mechanism for food fall detection in deep-sea scavengers. Deep-Sea Research I 49, 143–155. Krogh, A. 1934a. Conditions of life at great depths in the ocean. Ecological Monographs 4, 430–439. Krogh, A. 1934b. Conditions of life in the ocean. Ecological Monographs 4, 421–429.

350

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Kukert, H. & Smith, C. R. 1992. Disturbance, colonization and succession in a deep-sea sediment community: artiﬁcial-mound experiments. Deep-Sea Research I 39, 1349–1371. Levin, L. A., James, D. W., Martin, C. M., Rathburn, A. E., Harris, L. H. & Michener, A. E. 2000. Do methane seeps support distinct macrofaunal assemblages? Observations on community structure and nutrition from the northern California slope and shelf. Marine Ecology Progress Series 208, 21–39. Levin, L. A., Plaia, G. R. & Huggett, C. L. 1994. The inﬂuence of natural organic enhancement on life histories and community structure of bathyal polychaetes. In Reproduction, larval biology and recruitment of the deep-sea benthos, C. Young & K. Eckelbarger (eds). New York: Columbia University Press, 261–283. Levin, L. A. & Smith, C. R. 1984. Response of background fauna to disturbance and enrichment in the deep sea: a sediment tray experiment. Deep-Sea Research I 31, 1277–1285. Lockyer, C. 1976. Body weights of some species of large whales. Journal du Conseil International pour l’Exploration de la Mer 36, 259–273. Lonsdale, P. 1979. A deep-sea hydrothermal site on a strike-slip fault. Nature 281, 531–534. Lutz, R., Jablonski, D. & Turner, R. 1984. Larval development and dispersal at deep-sea hydrothermal vents. Science 226, 1451–1453. Marsh, A. G., Mullineaux, L. S., Young, C. M. & Manahan, D. T. 2001. Larval dispersal potential of the tubeworm Riftia pachyptila at deep-sea hydrothermal vents. Nature 411, 77–80. Marshall, B. A. 1987. Osteopeltidae (Mollusca: Gastropoda): a new family of limpets associated with whale bone in the deep sea. Journal of Molluscan Studies 53, 121–127. Marshall, B. A. 1994. Deep-sea gastropods from the New Zealand region associated with recent whale bones and an Eocene turtle. Nautilus 108, 1–8. Martill, D. M., Cruickshank, A. R. I. & Taylor, M. A. 1991. Dispersal via whale bones. Nature 351, 193 only. Martini, F. H. 1998. Ecology of hagﬁshes. In The biology of hagﬁshes, J. Jørgensen et al. (eds). London: Chapman & Hall, 57–78. McLean, J. H. 1985. Preliminary report on the limpets at hydrothermal vents. Biological Society of Washington Bulletin 6, 159–166. McLean, J. H. 1992. Cocculiniform limpets (Cocculinidae and Pyropeltidae) living on whale bone in the deep sea off California. Journal of Molluscan Studies 58, 401–414. McLean, J. H. & Haszprunar, G. 1987. Pyropeltidae, a new family of cocculiniform limpets from hydrothermal vents. Veliger 30, 196–205. Merritt, R. W. & Wallace, J. R. 2001. In Forensic entomology; the utility of arthropods in legal investigations, J. H. Byrd & J. L. Castner (eds). 2001. Boca Raton: CRS Press, 177–222. Minigawa, M. & Wada, E. 1984. Stepwise enrichment of 15N along food chains: further evidence and the relation between 15N and animal age. Geochimica et Cosmochimica Acta 48, 1135–1140. Mizroch, S. A., Rice, D. W. & Breiwick, J. M. 1984a. The blue whale, Balaenoptera musculus. Marine Fisheries Review 46, 15–19. Mizroch, S. A., Rice, D. W. & Breiwick, J. M. 1984b. The ﬁn whale, Balaenoptera physalus. Marine Fisheries Review 46, 20–24. Mizroch, S. A., Rice, D. W. & Breiwick, J. M. 1984c. The sei whale, Balaenoptera borealis. Marine Fisheries Review 46, 25–29. Mullineaux, L. S. 1987. Organisms living on manganese nodules and crusts: distribution and abundance at three North Paciﬁc sites. Deep-Sea Research 34, 165–184. Naganuma, R., Wada, H. & Fujioka, K. 1996. Biological community and sediment fatty acids associated with the deep-sea whale skeleton at the Torishima seamount. Journal of Oceanography 52, 1–15. Pearson, T. H. & Rosenberg, R. 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: an Annual Review 16, 229–311.

351

CRAIG R. SMITH & AMY R. BACO

Peek, A. S., Gustafson, R. G., Lutz, R. A. & Vrijenhoek, R. C. 1997. Evolutionary relationships of deep-sea hydrothermal vent and cold-water seep clams (Bivalvia: Vesicomyidae): results from mitochondrial cytochrome oxidase subunit I. Marine Biology 130, 151–161. Perry, S. L., DeMaster, D. P. & Silber, G. K. 1999. The great whales: history and status of six species listed as endangered under the U. S. Endangered Species Act of 1973. Marine Fisheries Review 61, 1–74. Pettibone, M. H. 1993. Polynoid polychaetes associated with a whole skeleton in the bathyal Santa Catalina Basin. Proceedings of the Biological Society of Washington 106, 678–688. Pielou, E. C. 1969. An introduction to mathematical ecology. New York: Wiley-Interscience. Priede, I. G., Bagley, P. M., Armstrong, J. D., Smith, K. L. & Merrett, N. R. 1991. Direct measurement of active dispersal of food-falls by deep-sea demersal ﬁshes. Nature 351, 647–649. Pulliainen, E. 1988. Ecology status and management of Finnish wolverine gulo-gulo populations. Lutra 31, 21–28. Rau, G. H. 1981. Low 15N/14N in hydrothermal vent animals: ecological implications. Nature 289, 484–485. Rau, G. H., Mearns, A. J., Young, D. R., Olson, R. J., Schafer, H. A. & Kaplan, I. R. 1983. Animal 13 C/12C correlates with trophic level in pelagic food webs. Ecology 64, 1314–1318. Rhoads, D. C., McCall, P. L. & Yingst, J. Y. 1978. Disturbance and production on the estuarine seaﬂoor. American Scientist 66, 577–586. Rice, D. W., Wolman, A. A. & Braham, H. W. 1984. The gray whale, Eschrichtius robustus. Marine Fisheries Review 46, 7–14. Robineau, D. & de Buffrénil, V. 1993. Nouvelles données sur la masse du squelette chez les grands cétacés (Mammalia, Cetacea). Canadian Journal of Zoology 71, 828–834. Sarrazin, J. & and Juniper, S. K. 1999. Biological characteristics of a hydrothermal ediﬁce mosaic community. Marine Ecology Progress Series 185, 1–19. Schoenly, D. & Reid, W. 1987. Dynamics of heterotrophic succession in carrion arthropod assemblages: discrete seres or a continuum of change? Oecologia 73, 192–202. Shelden, K. E. W. & Rugh, D. J. 1995. The bowhead whale, Balaena mysticetus: its historic and current status. Marine Fisheries Review 57, 3–4. Sibuet, M. & Olu, K. 1998. Biogeography, biodiversity, and ﬂuid dependence of deep-sea cold seep communities at active and passive margins. Deep-Sea Research II 45, 517–567. Smith, C. R. 1983. Enrichment, disturbance and deep-sea community structure: the signiﬁcance of large organic falls to bathyal benthos in Santa Catalina Basin. PhD dissertation, University of California at San Diego. Smith, C. R. 1985. Food for the deep sea: utilization, dispersal and ﬂux of nekton falls at the Santa Catalina Basin ﬂoor. Deep-Sea Research I 32, 417–442. Smith, C. R. 1986. Nekton falls, low-intensity disturbance and community structure of infaunal benthos in the deep-sea. Journal of Marine Research 44, 567–600. Smith, C. R. 1992. Whale falls: chemosynthesis on the deep-sea ﬂoor. Oceanus 36, 74–78. Smith, C. R. & Baco, A. R. 1998. Phylogenetic and functional afﬁnities between whale-fall, seep, and vent communities. Cahiers de Biologie Marine 39, 345–346. Smith, C. R., Baco, A. R. & Glover, A. 2002. Faunal succession on replicate deep-sea whale falls: time scales and vent-seep afﬁnities. Cahiers de Marine Biologie 43, 293–297. Smith, C. R. & Demopoulos, A. W. J. 2003. The deep Paciﬁc Ocean ﬂoor. In Ecosystems of the world Volume 28: Ecosystems of the deep ocean, P. A. Tyler (ed.). Amsterdam: Elsevier, 181–220. Smith, C. R. & Hessler, R. R. 1987. Colonization and succession in deep-sea ecosystems. Trends in Ecology and Evolution 2, 359–363. Smith, C. R., Kukert, H., Wheatcroft, R. A., Jumars, P. A. & Deming, J. W. 1989. Vent fauna on whale remains. Nature 34, 27–128. Smith, C. R., Mullineaux, L. S. & Levin, L. A. 1998. Deep-sea biodiversity: a compilation of recent advances in honor of Robert R. Hessler. Deep-Sea Research II 45, 1–12.

352

ECOLOGY OF WHALE FALLS AT THE DEEP-SEA FLOOR

Snelgrove, P. V. R., Grassle, J. F. & Petrecca, R. F. 1992. The role of food patches in maintaining high deep-sea species diversity: ﬁeld experiments using hydrodynamically unbiased colonization trays. Limnology and Oceanography 37, 1543–1550. Snelgrove, P. V. R., Grassle, J. F. & Petrecca, R. F. 1994. Macrofaunal response to artiﬁcial enrichments and depressions in a deep-sea habitat. Journal of Marine Research 52, 345–369. Squires, R. L., Goedert, J. L. & Barnes, L. G. 1991. Whale carcasses. Nature 349, 574 only. Stockton, W. L. & DeLaca, T. E. 1982. Food falls in the deep sea: occurrence, quality, and signiﬁcance. Deep-Sea Research I 29, 157–169. Tamburri, M. N. & Barry, J. P. 1999. Adaptations for scavenging by three diverse bathyal species, Eptatretus stouti, Neptunea amianta & Orchomente obtusus. Deep-Sea Research I 46, 2079–2093. Tebble, N. 1966. British bivalve seashells: a handbook for identiﬁcation. London: The British Museum (Natural History). Tunnicliffe, V. 1991. The biology of hydrothermal vents: ecology and evolution. Oceanography and Marine Biology: an Annual Review 29, 319–407. Tunnicliffe, V. & Fowler, M. R. 1996. Inﬂuence of sea-ﬂoor spreading on the global hydrothermal vent fauna. Nature 379, 531–533. Tunnicliffe, V. & Juniper, S. K. 1990. Cosmopolitan underwater fauna. Nature 344, 300 only. Tunnicliffe, V., McArthur, A. G. & McHugh, D. 1998. A biogeographical perspective of deep-sea hydrothermal vent fauna. Advances in Marine Biology 34, 353–442. Turner, R. D. 1973. Wood-boring bivalves, opportunistic species in the deep sea. Science 180, 1377–1379. Turner, R. D. 1977. Wood, mollusks, and deep-sea food chains. Bulletin of the American Malacology Union 213, 13–19. Tyler, P. A., German, C. R., Ramirez-Llodra, E. & Van Dover, C. L. ChEss: understanding the biogeography of chemosynthetic ecosystems. Oceanologica Acta (in press). Van Dover, C. L. 2000. The ecology of deep-sea hydrothermal vents. Princeton: Princeton University Press. Van Dover, C. L. & Fry, B. 1989. Stable isotope composition of hydrothermal vent organisms. Marine Biology 102, 257–263. Van Dover, C. L., German, C. R., Speer, K. G., Parson, L. M. & Vrijenhoek, R. C. 2002. Evolution and biogeography of deep-sea vent and seep invertebrates. Science 295, 1253–1257. Vetter, E. W. 1994. Hotspots of benthic production. Nature 372, 47 only. Vrijenhoek, R. C. 1997. Gene ﬂow and genetic diversity in naturally fragmented metapopulations of deep-sea hydrothermal vent animals. Journal of Heredity 88, 295–293. Wada, H., Naganuma, T., Fujioka, K., Ditazato, H., Kawamura, K. & Akazawa, Y. 1994. The discovery of the Torishima whale bone animal community and its meaning – the results of revisit dives by the “Shinkai 6500”. Japan Marine Science & Technology Center/Deep-Sea Research 10, 38–47. Warén, A. 1989. New and little known Mollusca from Iceland. Sarsia 74, 1–28. Warén, A. 1991. New and little known Mollusca from Iceland and Scandinavia. Sarsia 76, 53–124. Warén, A. 1993. New and little known Mollusca from Iceland and Scandinavia. Part 2. Sarsia 78, 159–201. Warén, A. 1996. New and little known Mollusca from Iceland and Scandinavia. Part 3. Sarsia 81, 197–245. Warén, A. & Bouchet, P. 1993. New records, species, genera, and a new family of gastropods from hydrothermal vents and hydrocarbon seeps. Zoologica Scripta 22, 1–90. Warén, A. & Bouchet, P. 2001. Gastropoda and Monoplacophora from hydrothermal vents and seeps; new taxa and records. Viliger 44, 116–231. Weston, D. P. 1990. Quantitative examination of macrobenthic community changes along an organic enrichment gradient. Marine Ecology Progress Series 61, 233–244. Willey, P. & Snyder, L. M. 1989. Canid modiﬁcation of human remains implications for time-sincedeath. Journal of Forensic Sciences 34, 894–901.

353

CRAIG R. SMITH & AMY R. BACO

Williams, A. B., Smith, C. R. & Baco, A. R. 2000. New species of the genus Paralomis White 1856 (Crustacea, Decapoda, Anomura, Lithodidae) from a sunken whale carcass in the San Clemente Basin off southern California. Journal of Crustacean Biology 20, Special No. 2, 281–285. Wolff, T. 1976. Utilization of seagrass in the deep sea. Aquatic Botany 2, 161–174. Wolff, T. 1979. Macrofaunal utilization of plant remains in the deep sea. Sarsia 64, 117–136. Zmarzly, D. L., Stebbins, T. D., Pasko, D., Duggan, R. M. & Barwick, K. L. 1994. Spatial patterns and temporal succession in soft-bottom macroinvertebrate assemblages surrounding an ocean outfall on the southern San Diego shelf: relation to anthropogenic and natural events. Marine Biology 118, 293–307.

354

Oceanography and Marine Biology: an Annual Review 2003, 41, 355–390 © R. N. Gibson and R. J. A. Atkinson, Editors Taylor & Francis

THE DIET OF HARBOUR PORPOISE (PHOCOENA PHOCOENA) IN THE NORTHEAST ATLANTIC M. B. SANTOS & G. J. PIERCE Department of Zoology, University of Aberdeen, Tillydrone Avenue, Aberdeen, AB24 2TZ, UK e-mail: [email protected]

Abstract The harbour porpoise (Phocoena phocoena) is probably the most abundant small cetacean in the northeast Atlantic and as such is an important top predator. It is also one of the most threatened species, particularly as a consequence of ﬁshery by-catch. Porpoises feed mainly on small shoaling ﬁshes from both demersal and pelagic habitats. Many prey items are probably taken on, or very close to, the sea bed. Even though a wide range of species has been recorded in the diet, porpoises in any one area tend to feed primarily on two to four main species (e.g. whiting (Merlangius merlangus) and sandeels (Ammodytidae) in Scottish waters). Evidence for selective predation is equivocal. Many studies provide evidence of geographic, seasonal, interannual, ontogenetic or sexual differences in prey types or prey sizes, and such differences are often (speculatively) interpreted in terms of prey availability. A few studies demonstrate trends in diet selection that are consistent with changes in prey abundance. However, lack of availability of prey abundance data at an appropriate spatial and temporal scale is often a problem. Porpoise diets overlap extensively with diets of other piscivorous marine predators (notably seals). Many of the main prey species are also taken by commercial ﬁsheries, although porpoises tend to take smaller ﬁshes than those targeted by ﬁsheries. Given their high abundance, porpoises clearly remove substantial quantities of ﬁsh. The literature on porpoise diets in the northeast Atlantic suggests that there has been a longterm shift from predation on clupeid ﬁsh (mainly herring Clupea harengus) to predation on sandeels and gadoid ﬁsh, possibly related to the decline in herring stocks since the mid-1960s. Evidence from studies on seals suggests that such a shift could have adverse health consequences. Food consumption brings porpoises into contact with two important threats – persistent organic contaminants and ﬁshing nets, both of which have potentially serious impacts.

Introduction The harbour or common porpoise, Phocoena phocoena (Linnaeus 1758), is one of the six species recognised in the family Phocoenidae (Read 1999). Its common name, porpoise, derives from the Latin porcus piscus (pigﬁsh) and was used in Ancient Rome. Linnaeus (1758) distinguished it from the common dolphin by calling it Delphinus phocoena, from the Greek word phokia (seal), due to its lack of beak and its seal-like appearance. 355

M. B. SANTOS & G. J. PIERCE

Harbour porpoises, one of the most common cetaceans in European waters (Watson 1985), are small with an average adult length of 150 cm to 160 cm and an average weight of 45 kg to 60 kg (Gaskin et al. 1974). Maximum sizes of animals stranded in the UK have been reported as 163 cm and 54 kg in males and 189 cm and 81 kg in females (Lockyer 1995) although more recent work by Santos et al. (2001a) reported maximum sizes for males and females as 170 cm (55 kg) and 171 cm (55.5 kg), respectively, for porpoises stranded in Scotland. Harbour porpoises stranded in northwest Spain and Portugal seem to be larger, and several specimens have measured more than 200 cm (Donovan & Bjørge 1995, Sequeira 1996). There is some variation in the maximum ages reported for different harbour porpoise populations: no porpoises over 17 yr of age have been found in the Bay of Fundy (eastern Canada) (Read & Gaskin 1990, Read & Hohn 1995) while animals up to 24 yr old have been reported in UK waters (Lockyer 1995) and off California (Hohn & Brownell 1990). Harbour porpoises are widespread in coastal waters of the Northern Hemisphere; their range extends northwards from 14–15°N in the North Atlantic and from 30°N in the North Paciﬁc (Evans 1980). They are also found in the Black Sea and some records suggest the existence of a separate population off northwestern Africa (Smeenk et al. 1992). Historically, the species was also found in the western Mediterranean but there are no conﬁrmed records since the nineteenth century (Marchessaux 1980, Donovan & Bjørge 1995). The species is also found over offshore shallows (e.g. the Georges and Grand Banks, stretching from Newfoundland to southern New England on the edge of the North American continental shelf) and around islands such as the Faeroes and Iceland. Individuals have also been recorded considerable distances up rivers (e.g. 320 km up the river Mass in Holland, Gaskin 1984). Intraspeciﬁc differences in morphometric and meristic skull characters have been demonstrated between the North Paciﬁc, North Atlantic and Black Sea (Tomilin 1957, Kinze 1985, Miyazaki et al. 1987, Yurick & Gaskin 1987, Amano & Miyazaki 1992). The differences between these three areas, suggesting reproductive isolation, have been conﬁrmed by studies on mitochondrial DNA (Rosel et al. 1995, Wang et al. 1996). Rosel et al. (1995) suggested that there are at least three subspecies: P. phocoena phocoena (Atlantic), P. phocoena vomerina (Paciﬁc) and P. phocoena relicta (Black Sea). The existence of further population subdivisions within these three major areas has been proposed by several authors using a variety of methods (e.g. Yurick 1977, Gaskin 1984, Yurick & Gaskin 1987, Andersen 1993, Rosel et al. 1995, 1999, Tiedemann et al. 1996, Wang et al. 1996, Börjesson & Berggren 1997, Walton 1997, Wang & Berggren 1997, Berrow et al. 1998, Lockyer 1999, Westgate & Tolley 1999, Tolley et al. 1999, 2001, Tolley & Heldal 2002) and putative stock boundaries have been suggested by the International Whaling Commission’s (IWC) Scientiﬁc Committee (Donovan & Bjørge 1995). Historically, harbour porpoises have been taken in European waters for human consumption and for ﬁsh bait (Watson 1985). It is believed that, in some areas, the porpoise hunt goes back to the Stone Age (e.g. inner Danish waters, Möhl 1970 quoted in Kinze 1995). Porpoise meat was greatly appreciated during the Middle Ages. It was served in the English court at coronation banquets and on other important occasions (Whymper 1883). In Denmark, a harbour porpoise ﬁshery was ﬁrst mentioned in 1357 and continued until the Second World War (Kinze 1995). Kinze estimated an average catch of at least 1000 animals took place every year in Danish waters for three centuries. Along the coast of Norway, large numbers were taken as early as the eleventh century (Watson 1985). At present hunting continues, on a small scale, only in Greenland and the Faeroes (IWC 1996). 356

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Accidental (by-catch) mortality in ﬁshing gear is recognised as the main threat at present for harbour porpoise populations in the northeast Atlantic (ASCOBANS 1994, IWC 1994, 1996, Donovan & Bjørge 1995, Tregenza et al. 1997, Vinther 1999). The harbour porpoise suffers a high degree of by-catch mortality across its range (Read & Gaskin 1988). In northern Europe, it is the most frequently caught cetacean species in ﬁshing nets and it is caught in a wide variety of ﬁshing gear, mainly gill nets but also salmon drift nets, pound nets for herring and salmon and mid-water trawls (Lindroth 1962, van Utrech 1978, Andersen & Clausen 1983, Kinze 1990a, Northridge & Lankester 1992, Clausen & Kinze 1993 quoted in Lowry & Teilmann 1994, Tregenza et al. 1997, Vinther 1999). There are historical records of porpoises caught in gill nets as long ago as the sixteenth century in Europe (Belon 1551, quoted in Harmer 1927), but it is clear that increased ﬁshing effort together with innovations in net design and use of thinner synthetic ﬁbres (which reduces detectability) have resulted in steeply increased catch rates. The impact of by-catches on harbour porpoise populations is, at the moment, difﬁcult to assess but there is concern about the sustainability of these catches. In 1990 and 1991, the Scientiﬁc Committee of the IWC recommended, with the “highest priority”, the reduction of by-catch for harbour porpoise (IWC 1991, 1992). It also noted the need to improve knowledge of stock identity and migration, and obtain reliable ﬁgures on by-catches and on abundance. An additional cause of harbour porpoise mortality in Scotland is the bottlenose dolphin (Tursiops truncatus), particularly in the Moray Firth area. Post-mortem examinations of a number of porpoise carcasses (all with multiple skeletal fractures and damage to internal organs), provided the ﬁrst evidence of dolphin attacks, when measurements of tooth marks present in the skin corresponded with the tooth spacing in Tursiops mandibles (Ross & Wilson 1996). Subsequently, bottlenose dolphins were observed attacking porpoises and repeatedly throwing them out of the water. The reasons for these interactions are unknown, but could include competition for food, play, practice-ﬁshing or sexual behaviour (Ross & Wilson 1996). In July 1994, a survey of Small Cetacean Abundance in the North Sea (SCANS) was carried out to estimate numbers of harbour porpoises and other small cetaceans in the North Sea and adjacent waters. To date this survey has provided the most accurate and complete information on population ﬁgures for harbour porpoises in the area. Previous ship and aerial surveys were on a smaller scale (e.g. Heide-Jørgensen et al. 1992, 1993, Bjørge & Øien 1995). The population estimates for harbour porpoise obtained from SCANS are given in Table 1. As pointed out by Hammond et al. (1995), these ﬁgures were calculated from observations that took place in only one month. Porpoises clearly exist in large numbers in the northeast Atlantic (Hammond et al. 1995) and, as such, are likely to have a quantitatively important place in the marine food web. In addition to their ecological importance, as ﬁsh eaters they may compete with commercial ﬁsheries. Diet has consequences for individual ﬁtness and, ultimately, population status and it is of interest to determine whether the large-scale ﬂuctuations in ﬁsh stocks seen in the second half of the twentieth century have inﬂuenced porpoise diet and population trends. In so far as porpoises prey on species also eaten by humans, the health consequences of diet choice (e.g. in terms of nutrition or bioaccumulation of contaminants) are of direct relevance to the human population. The general biology of harbour porpoises is reviewed in Bjørge & Donovan (1995) and Read (1999). However, to date there has been no major review of diets of harbour porpoises 357

M. B. SANTOS & G. J. PIERCE

Table 1 Abundance estimates of harbour porpoise for the different blocks surveyed by the Small Cetacean Abundance in the North Sea (SCANS) survey (data from Hammond et al. 1995), CV coefﬁcient of variation. Block

Area covered in the survey

CV

Porpoise abundance

A B C D E F G H I J L M X Y Total

Celtic Shelf Channel and South tip North Sea East coast of Britain West northern North Sea (excluding J) North central North Sea Central North Sea (56°N–58°N) South central North Sea South eastern North Sea Kattegat Waters off Shetland and Orkney Coastal area S and W of Jutland Off SW coast of Norway Bay of Kiel Northern Wadden Sea

0.57 0 0.18 0.25 0.49 0.25 0.34 0.29 0.34 0.34 0.47 0.27 0.48 0.27 0.14

36 280 0 16 939 37 144 31 419 92 340 38 616 4211 36 046 24 335 11 870 5666 588 5912 341 366 (260 000–449 000)

(95% conﬁdence intervals)

and little attention has been given to their ecological role as marine top predators. The present paper reviews information on the diet of harbour porpoises in the northeast Atlantic and will try to answer the following questions: (1) (2) (3) (4) (5)

(6)

What is their feeding niche: what do harbour porpoises eat and where do they obtain their food? Are they specialist or generalist predators and is their diet related to prey abundance (opportunistic or selective)? Do porpoises compete with other top predators such as seals and other cetaceans, and do they compete with ﬁsheries? Is there evidence of long-term trends or changes in diet composition? What are the consequences of diet composition for individual health and population status (e.g. is the diet nutritionally adequate, does diet choice play a role in ﬁshery by-catch mortality, and are there adverse consequences of contaminant bioaccumulation)? What important questions require further research?

Study methods Almost all published accounts of porpoise diet are based on stomach contents analysis. The largest scale studies are associated with hunting of porpoises in the Black Sea (Tsalkin 1940, quoted in Tomilin 1957) but most studies are on by-caught and stranded animals. The methodology of stomach contents analysis was reviewed by Pierce & Boyle (1991). The present review focuses mainly on results from stomach contents analysis. However, 358

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

some information on diet can also be obtained indirectly from studies on distribution and behaviour. Relatively new techniques for obtaining information on marine mammal diets such as fatty acid signature analysis of blubber and stable isotope analysis have not yet been applied to porpoises on a large scale. Fatty acids are mainly of dietary origin and stable isotope values vary geographically and, in a systematic way, with trophic level. In contrast with stomach contents analysis, these new techniques can provide information on the prey ingested over a longer period (days to months; e.g. Tiezen 1978, Hobson 1990, Iverson et al. 1995, Kirsch et al. 1998, 2000, Hooker et al. 2001). However, stable isotope analysis provides relatively coarse level data on diet, providing data mainly on the trophic level at which an animal feeds (see Pauly et al. 1998). On the other hand, fatty acid analysis has the potential to provide quantitative data on diet, although realisation of this potential requires information on the fatty acid composition of all putative prey, including data on variability, and the computing power to solve the equivalent of an extensive series of simultaneous equations. Fatty acid signature analysis is well-established as a technique for studies on seals (e.g. Iverson et al. 1997). Doubts were raised as to its applicability to cetaceans due to blubber stratiﬁcation: the outer blubber layers being relatively metabolically inert while the inner layers, closer to the muscle, are metabolically active. Also, blubber from different parts of the body may show differences in fatty acid composition, as found for some cetaceans (Ackman & Lamothe 1989). Recent studies in harbour porpoise have shown that dietary fatty acids are concentrated in the inner blubber (Koopman et al. 1996, Koopman 1998) and the fatty acid composition seems to be uniform in the body with the exception of the caudal peduncle.

What do harbour porpoises eat? The earliest published information available on the diet of harbour porpoises derives from the examination of single specimens, which were captured accidentally in ﬁshing gear or stranded, during the nineteenth century. Dewhurst (1834) records that “porpoises live upon small ﬁsh though they will eat any offal and garbage that is thrown into the sea”. Southwell (1881) observed that “the food consists of ﬁsh and it follows the shoals of herring, etc., amongst which it commits great depredation; it has a taste for salmon (Salmo salar) and is sometimes taken in the salmon-nets”. Van Beneden (1889) stated that “the porpoise preys on ﬁsh, including herring, but may also eat crustaceans, cephalopods and even marine plants”. Descriptions of porpoise diets from the late nineteenth and early twentieth centuries include records of a wide variety of ﬁshes and, in some cases, cephalopods and crustaceans in stomachs. Scott (1903), examining the stomach contents of a porpoise caught in a salmon net in Scotland, found remains of ﬁsh (ﬂesh and otoliths), of which the most frequent was whiting (Merlangius merlangus). The high numbers of otoliths found (240), in the opinion of the author, showed “how destructive these cetaceans can be when they get among a shoal of ﬁshes”. Various small-scale studies in the northeast Atlantic from the early part of the twentieth century through to the 1970s document predation by porpoises on whiting, herring (Clupea harengus) and sometimes other ﬁsh species such as capelin (Mallotus villosus), mackerel (Scomber scombrus), sole (Solea solea), cod (Gadus morhua), eel (Anguilla anguilla), as well as shrimps and cuttleﬁsh (Millais 1906, Stephen 1926, Harmer 1927, Freund 1932 quoted in Tomilin 1957, Fraser 1946, Darling 1947, Matthews 1952, Hardy 1959, Slijper 1962, Andersen 1965, Källquist 1975 quoted in Otterlind 1976). 359

M. B. SANTOS & G. J. PIERCE

Similar types of prey were recorded in other areas: for example, Paciﬁc herring (Clupea pallasii), capelin, Paciﬁc sardine (Sardinops coerulea) and smelt (Osmeridae) in the North Paciﬁc (Sleptsov 1952 quoted in Tomilin 1957, Wilke & Kenyon 1952, Scheffer 1953, Fink 1959); herring, cod, mackerel (Scomber scombrus), hake (Urophycis tenuis), pollack (Pollachius virens), squid (Loligo pealii), smelt (Osmerus mordax), silver hake (Merluccius bilinearis) and redﬁsh (Sebastes marinus) off eastern Canada (Sergeant & Fisher 1957, Smith & Gaskin 1974). The most extensive published study on harbour porpoise diet was carried out in the Black Sea, where 4000 stomachs were examined by Tsalkin (1940 quoted in Tomilin 1957). The diet consisted mainly of benthic ﬁsh (several goby species, Gobius rotan, G. melanostomus, G. syrman, Mesogobius batrachocephalus; Black Sea ﬂounder, Pleuronectes ﬂesus ﬂesus; Black Sea sole, Solea nasuta; bream, Abramis brama and Black Sea whiting, Gadus euxinus). Pelagic species (Black Sea silverside, Atherina pontica; Black Sea anchovy, Engraulis encrasicholus; pikeperch, Lucioperca lucioperca; mullet, Mugil sp. and Black Sea shad, Caspialosa sp.) were also found but were thought to be taken only when they occurred in large and dense schools. In the northeast Atlantic, the ﬁrst detailed study on harbour porpoise diet was carried out by Rae (1965). This author, concerned with the possible role of porpoises as predators of salmon, examined the stomach contents of 52 animals by-caught in different ﬁshing gears in Scotland from 1959 to 1965. Herring and whiting were the main prey found. In a later study, from 1965 to 1971, a further 30 by-caught and 11 stranded porpoises were examined (Rae 1973). In these 41 stomachs, the main prey were clupeids (herring and sprat Sprattus sprattus) and small gadoids (mainly whiting). No evidence of predation on salmon was found in either study. Lindroth (1962) examined the stomach contents of 50 harbour porpoises from the Baltic Sea captured from 1960–1, again with the aim of ﬁnding out whether porpoises prey on salmon, and also found no evidence of salmon in the diet. Remains of sprat, herring, Baltic cod (G. m. callarias), gobies (Aphya minuta) and sandeels (Ammodytes sp.) were found. As schemes to record and examine stranded cetaceans were set up in different countries and research programmes started collecting by-caught specimens during the 1980s and 1990s, more studies on harbour porpoise diets were carried out (see Table 2). The results of these studies demonstrated that harbour porpoises feed on both pelagic schooling ﬁsh species, for example, herring, capelin, whiting, blue whiting Micromesistius potassou, sardine, northern anchovy, and demersal or benthic ﬁsh, for example, hake, Trisopterus spp., sandeels (Ammodytidae), gobies, sole, dab Limanda limanda, Greenland halibut Reinhardtius hippoglossoides (Desportes 1985, Sekiguchi 1987, Kinze 1989, Lick 1991a,b, Aarefjord et al. 1995, Börjesson & Berggren 1996, Malinga & Kuklil 1996, Rogan & Berrow 1996, Santos 1998). Other prey species such as cephalopods (ommastrephids, sepiolids, loliginids, gonatids), other molluscs, crustaceans (e.g. euphausiids) and polychaetes (Nereis) were also recorded (Smith & Gaskin 1974, Desportes 1985, Sekiguchi 1987, Kinze 1989, Smith & Read 1992, Gaskin et al. 1993, Gearin et al. 1994, Rogan & Berrow 1996, Santos 1998). Plastic and other foreign objects such as nylon ﬁshing line and a banana peel have also been recorded in the stomachs (Kastelein & Lavaleije 1992, Baird & Hooker 2000).

360

361

20 (b)

100 (b) 15 (m)

Washington California

94 9

5

18 20

127 136 c 24 111 95

160 (b) 149 (b) c 31 (b) 138 (b) 95 (b)

18 (d) 20 (d)

42 * 14

48 (m) 27 (m) 16 (b) a

145

171 (m)

a

8 179 36 * 19 100 72 62 58 6

Stomachs with food remains

17 (s) 247 (m) 54 (m) 34 (m) 26 (m) ⬃250 (b) 72 (m) 62 (m) 58 (m) 6 (m)

Stomachs analysed

Paciﬁc Ocean Northern Japan

Greenland

Poland NW Atlantic Eastern Canada

Baltic Sea Germany

Ireland United Kingdom Scotland The Netherlands Denmark NW Spain Skagerrak and Kattegat Seas Sweden

NE Atlantic France Denmark, Sweden, Norway Germany

Area

Ommastrephid squids, herring, anchovy, hake Paciﬁc herring, squid, smelts Northern anchovy

Herring, silver hake, cod Clupeids, gadids Euphausiids Capelin, herring Herring, silver hake, red and white hakes Greenland halibut, haddock Capelin, halibut

Cod, gobies, herring Gobies, herring, cod Cod, gobies, herring

Herring, sprat, whiting

Blue whiting, scad, hake Herring, gadids Sole, cod Sandeels, sole Trisopterus spp., whiting, herring Gadids, sandeels, gobies Whiting, sandeels Whiting Cod, viviparous blenny, whiting Scad, sandeels, Trisopterus spp.

Main Prey

Gearin et al. 1994 Sekiguchi 1987

Gaskin et al. 1993

Kinze 1989 Kinze 1990b

Fontaine et al. 1994 Gannon et al. 1998

Recchia & Read 1989 Smith & Read 1992

Lick 1991a,b Benke & Siebert 1996 Malinga & Kuklik 1996

Berggren 1996, Börjesson & Berggren 1996

Desportes 1985 Aarefjord et al. 1995 Lick 1991a,b Benke & Siebert 1996 Rogan & Berrow 1996 Martin 1996 Santos 1998 Santos 1998 Santos 1998 Santos 1998

Reference

Table 2 Main studies on harbour porpoise diets since the 1980s (s stranded animals, b by-caught, m stranded and by-caught, d direct hunt). *No data on number of stomachs with food provided, aadult animals, ccalves (from Smith & Read 1992).

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

M. B. SANTOS & G. J. PIERCE

Are porpoises opportunistic or selective predators? As Dunnet (1996) pointed out, opportunism, selection and availability are “in fact shorthand for very complex biological interactions about which we know only a little in quantitative terms”. Strictly applied, the term “opportunistic” predation can be taken to imply that prey are taken as encountered, with the implication that prey availability is the only criterion affecting diet choice. Following the theory of optimal diet selection (e.g. Pulliam 1974), observation of opportunistic predation (sensu stricto) might be taken to imply that high quality prey are relatively rarely encountered. However, the term is probably used at least as often as a catch-all phrase which indicates no more than that evidence for active selection was not found (and often was not sought). The diversity of prey eaten and the geographical variation found in the diet have led some authors to consider the harbour porpoise to be an opportunistic feeder (e.g. Martin 1996, Teilmann & Dietz 1996), not limited to shallow waters, but able also to feed pelagically on midwater species from deeper habitats (IWC 1996). In fact, few studies in the northeast Atlantic have been able to address this question because information on prey abundance, for all the species at an appropriate spatial and temporal scale, is rarely available. Donovan & Bjørge (1995) note that, to answer the question of whether harbour porpoise feeding patterns follow prey availability, it would be necessary to study “the distribution of prey and target species on a very small spatial scale, much smaller than presently documented in ﬁshery literature”. Santos (1998) compared the rank order of importance (by weight) of different ﬁsh species in harbour porpoise diets and ﬁshery landings in Scotland, excluding species eaten by porpoises but not commercially ﬁshed, and found a positive correlation in 3 yr out of the 5 yr studied. Species that are not ﬁshed made up between 0.3% (1996) and 2.6% (1994) of the diet by weight but their absence in landings data simply reﬂected low (or zero) market value. Using catch statistic data as a measure of species abundance has potential risks since ﬁshery landings are also affected by changes in market demand, ﬁshing effort, establishment of management measures such as minimum landing size, total annual catches, area closures, etc. (Hislop 1996). Nevertheless, ﬁshery-catch data can sometimes provide a reasonable estimate of the abundance, distribution and availability of the prey species (Evans 1975), and the correlation between the importance rankings of certain species in commercial catches and in porpoise diets (Santos 1998) remains of interest. To the extent that ﬁsheries take species in proportion to their abundance, this result provides weak support for the notion of porpoises as an opportunistic species. Porpoise distribution around Shetland was spatially correlated with sandeel distribution in 1992 and 1993 but there was no correlation between the seasonal patterns of porpoise numbers and prey abundance (Evans & Borges 1995). The nature of feeding strategies can be revealed by studies of dietary variation. Against a background of varying ﬁsh abundance, piscivorous predators, especially those foraging opportunistically, might be expected to show regional, seasonal or interannual variation in diet. On the other hand, evidence of sex- or age-related variation in diet may be consistent with opportunistic predation (if there is sex- or age-related habitat segregation) or provide evidence for the role of other factors (e.g. if both sexes occupy the same habitat). Despite the existence of numerous studies on harbour porpoise diet there has been little quantitative analysis of patterns in diet in the northeast Atlantic. This is mainly due to the lack of appropriate sample sizes that would allow disentanglement of the effects of all the factors potentially affecting diet variability. 362

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Geographical variation Several studies on harbour porpoise diets in the northeast Atlantic have revealed geographical variation in the main prey consumed (in the UK, Martin 1996, Santos 1998; in Norway, Denmark and Sweden, Aarefjord et al. 1995, Berggren 1996; in Germany, Lick 1991a,b, Benke & Siebert 1996; in Ireland, Rogan & Berrow 1996). One caveat is that not all of these studies were contemporaneous and the confounding effect of temporal shifts in diet cannot always be ruled out. Based on examination of prey remains from the stomachs of 100 porpoises stranded and by-caught on the British coast from 1989 to 1994, the most important prey in terms of biomass were gadoids (whiting, haddock, Norway pout Trisopterus esmarkii and pollack), while sandeels and gobies were by far the most frequently eaten. Differences were found between diets in different areas of the British coast, with sandeels taken in bigger numbers on the east coast, while Norway pout was taken by more than half of the porpoises from Shetland, but was not eaten elsewhere (Martin 1996). The greater dietary importance of Trisopterus in Shetland than elsewhere in Scotland is also supported by results in Santos (1998). Harbour porpoises from Irish waters fed mainly on Trisopterus spp. (Rogan & Berrow 1996). Signiﬁcant between-area differences in the diet were reported by Santos (1998), who analysed the stomach contents of 198 stranded and by-caught porpoises from Denmark, The Netherlands, Scotland and Galicia (NW Spain) mainly from 1989 to 1996. The author found various signiﬁcant between-area differences in the diet. Porpoises from Scotland had eaten signiﬁcantly more sandeels than had those from Denmark and Holland, also more sepiolids and fewer cod than in Denmark. Danish porpoises took signiﬁcantly more gobies than did porpoises in Scotland, while viviparous blennies (Zoarces viviparus) were present only in the Danish diet. In Holland, porpoises had taken signiﬁcantly more gobies, dragonets and squid Loligo forbesi than had porpoises from Scotland, and more sepiolids than porpoises from Denmark. In Galicia, despite a small sample size, a wider range of prey was recorded than in other areas, including some species found only in the Galician diet (e.g. silvery pout Gadiculus argenteus thori, argentine Argentina sp.). Regional differences in the diet were also found by Aarefjord et al. (1995), who examined stranded and by-caught specimens from Norwegian, Danish and Swedish waters (1985–90). Overall, herring was the most important single prey species, while gadoids made up more than half of the total prey weight. However, harbour porpoises from the Danish North Sea and the Baltic took mainly cod, whiting, sandeels and gobies, while saithe, blue whiting and capelin were more frequent in porpoises from the Norwegian waters. Herring and sprat were found to be the main food of harbour porpoises stranded and by-caught during 1988–93 in the Swedish Skagerrak and Kattegat Seas (Berggren 1996). In German waters, results from Lick (1991a,b) and Benke & Siebert (1996), from strandings and by-catches in 1985–90 and 1991–3, respectively, indicate geographic differences in the diet of harbour porpoises. In porpoises from the North Sea from 1991–3, sandeels made up almost 40% of the prey weight, with a further 30% being common sole (Solea solea). In contrast, in the Baltic Sea over 50% of the total prey weight was made up of gobies, while 23% was herring and 15% was cod (Benke & Siebert 1996).

363

M. B. SANTOS & G. J. PIERCE

Seasonal variation Seasonal variation in harbour porpoise distribution has been described as a general inshore movement in summer and offshore movement in winter, although east–west and north–south migrations have also been proposed in different locations (e.g. Tomilin 1957, Gaskin et al. 1974, Gaskin 1977, 1984, Taylor & Dawson 1984, Gaskin & Watson 1985, Northridge et al. 1995, Read & Westgate 1997). Seasonal movements are believed to be related to prey availability or to breeding habitat (Gaskin 1977, Northridge et al. 1995, Read & Westgate 1997). Camphuysen & Leopold (1993) interpreted the higher numbers of porpoise sightings and strandings on the Dutch coast in winter than in summer as indicating a seasonal east-west movement. Aspects of the ecology of some of the prey species may assist in interpreting seasonal differences in diet (Santos 1998). Sandeels spend most of the autumn and winter (September to March) buried in the sand, with the exception of the spawning period between December and January (Macer 1966, Reay 1970, Langham 1971, Winslade 1974, Wright 1996). This is consistent with the higher prevalence of sandeels in spring and summer diets than in autumn and winter diets of Scottish harbour porpoises (Santos 1998). Although, arguably, echolocating porpoises should be able to detect sandeels in the sand it is perhaps energetically more costly to catch them. The higher importance of whiting in the winter diet could relate to the lower availability of sandeels but is also consistent with trends in whiting abundance. Trawl survey data for the east coast of Scotland indicate that whiting, poor cod and Norway pout are more abundant in inshore waters in winter than in summer (Santos et al., unpubl. data). Seasonal variation was also documented in the diet of harbour porpoises by-caught in Swedish waters: while herring was the main prey all year round, the contribution of sprat and whiting varied seasonally (Börjesson & Berggren 1996). Seasonal differences have also been reported in the size of prey eaten by porpoises (Santos 1998). In Scotland, smaller whiting were taken in autumn and bigger whiting were eaten in spring and summer, while bigger sandeels were taken in winter and spring than summer. In Denmark, smaller viviparous blennies and whiting were taken in spring than summer. In Holland, smaller gobies were taken in autumn than in winter and spring. Such trends may also be interpreted with reference to ﬁsh life cycles, in that small ﬁshes are most likely to be taken when 0-group ﬁshes move into the area used by porpoises. Thus, large numbers of 0-group sandeels become available in the summer months, when they are preyed upon by harbour porpoises and a wide variety of birds, ﬁshes and other marine mammals (e.g. Furness & Hislop 1981, Jonsgåard 1982, Perkins et al. 1982, Furness 1987, Daan 1989, Harris & Riddiford 1989, Monaghan et al. 1989, Pierce et al. 1989, 1991a,b, Harris & Wanless 1991, Storey 1993, Thompson et al. 1991). Seasonal differences in the diet of harbour porpoises have been reported also in studies outside the northeast Atlantic. In Atlantic Canadian and United States waters, prey diversity was higher in winter than in summer, with porpoises eating mainly herring, silver hake and pearlsides (Maurolicus weitzmani) in the autumn (Palka et al. 1996). Gannon et al. (1998) analysed the stomach contents of 95 harbour porpoises by-caught in autumn in the Gulf of Maine and compared the results with previous studies of harbour porpoises by-caught in summer in the Bay of Fundy (Smith & Gaskin 1974, Recchia & Read 1989, Smith & Read 1992). They noted that herring was the main prey for porpoises in both autumn and summer but was less dominant in the autumn diet than in summer. They also found that porpoises ate a wider variety of prey and of prey sizes in autumn than in summer. 364

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Interannual variation Interannual variation in the diet might be expected to follow variation in the availability of preferred prey. Declines in the availability of common prey could lead harbour porpoises to switch to other prey species and/or prey sizes. For some of the commercial ﬁsh species eaten by porpoises, estimates of abundance and size distributions are available. However, this is not the case for many other prey species (e.g. the smaller inshore ﬁsh such as blennies, gobies and eels). Signiﬁcant interannual differences in the average size of ﬁshes eaten (e.g. for sandeels and whiting in Scotland, viviparous blenny and whiting in Denmark, and gobies in Holland) were described by Santos (1998). In terms of amounts eaten, signiﬁcant interannual variation was found only for herring, and the signiﬁcance of this variation was strongly inﬂuenced by a single porpoise killed by bottlenose dolphins in the Moray Firth in November 1994, which had eaten many small herring (150 mm). In any case, the interannual changes were apparently unrelated to changes in herring abundance. However, it is worth noting that of the three main studies on porpoise diets in UK waters, only the earliest (Rae 1965, 1973) records herring as forming a major part of the diet and this change could reﬂect the decline in herring abundance in the North Sea since the 1960s. This topic is revisited in more detail below. Outside the northeast Atlantic, Recchia & Read (1989) found some differences in the diet of harbour porpoises by-caught in groundﬁsh gill nets in the Bay of Fundy (Canada) during two time periods (1969–72 and 1985–7). Porpoises from 1969–72 had taken mackerel and silver hake less often than porpoises from 1985 to 1987 although herring, silver hake and cod remained the main prey in the diet for both samples. The main prey (by percentage occurrence) of harbour porpoises off northern Washington differed between the years 1988–90. In 1988 it was Paciﬁc herring, followed by squid and smelt, while in 1989 smelt was the main prey followed by squid (Loligo opalescens) and gadoids. In 1990, Paciﬁc herring was again the main prey followed by smelt and gadoids (Gearin et al. 1994).

Ontogenetic variation Differences in diet between adult and juvenile porpoises in the northeast Atlantic have been found in several studies (Lick 1991a,b, Benke & Siebert 1996, Börjesson & Berggren 1996, Santos 1998). Juveniles cannot dive as deep as adults and could be prevented by their small size from catching and eating big prey. Differences in the diet of young (120 cm total length) and adult porpoises were found in a sample of 78 stomachs from porpoises stranded and by-caught in Germany. Young porpoises took more gobies, while adult porpoise took more ﬂatﬁshes and gadoids and had a bigger variety of prey species in the stomach (Lick 1991a,b). Similar results were found in a sample of 61 porpoises by-caught and stranded in Germany during 1991–3 (Benke & Siebert 1996). Börjesson & Berggren (1996) also noted that gobies were important in the diet of calves (1-yr old) from porpoises by-caught off Swedish waters. The authors concluded that the small size of gobies could make them a suitable prey for calves. Santos (1998) found that, in Scotland, adult harbour porpoises ate bigger whiting than did juveniles, while in Denmark juveniles ate bigger viviparous blenny and whiting than adults, and in Holland 365

M. B. SANTOS & G. J. PIERCE

adults took bigger gobies and sandeels than juveniles. The author considered that it was possible that most of these differences related to adult porpoises feeding further offshore than juveniles. In addition, the analysis showed that, in Holland, smaller porpoises took fewer whiting but more gobies than did bigger porpoises. In other studies in the northeast Atlantic, no signiﬁcant differences were found between the diets of calves (113 cm total length) and adult porpoises in Scandinavian waters, although gobies were the most frequent prey in the stomach of 0 and 1-yr old porpoises (Aarefjord et al. 1995), or between diets of juvenile and adult porpoises in a sample of animals stranded and by-caught in the UK (Martin 1996). Outside the northeast Atlantic, although no differences were found between the diets of juvenile and adult porpoises in Canadian waters (Smith & Gaskin 1974), comparison of a sample of 31 calves (1-yr old) and 149 adult porpoises by-caught in gill nets in the Bay of Fundy from 1985 to 1991 revealed that calves were taking euphausiids during their ﬁrst summer while adult porpoises ate herring (Smith & Read 1992). The authors suggested that calves “learn” to forage on euphausiids before starting to take larger prey such as ﬁshes. A comparison of the diets of 13 calves and 74 juvenile and adult porpoises by-caught in gill nets in the Gulf of Maine in 1989 and 1991–4 showed that calves had eaten a greater proportion of pearlsides and euphausiids than adults and had also taken smaller herring and silver hake (Gannon et al. 1998).

Diet of male and female harbour porpoises Differences in diets of males and females might be expected if the sexes tend to inhabit different areas, as a consequence of the larger average body size in females, and if the foraging behaviour of females is affected by the presence of nursing calves. Segregation of harbour porpoises in groups of different sex and/or age has been proposed by several authors to explain differences in by-catch ﬁgures. The predominance of mature males in the catch in the Baltic Sea was explained by Tomilin (1957) as a consequence of adult males forming separate groups that are more “mobile” than groups comprising juveniles or females with calves. Such segregation by age and sex has also been suggested as an explanation of high catches of sub-adult males in nets in offshore Canadian waters (Kinze 1994). In contrast, females accompanied by calves would tend to be associated with shallower waters (Kinze 1994). If this segregation takes place, females with calves would not only have a different distribution from males but could also be restricted in their search for food (e.g. by not being able to dive very deep or search long distances). Seasonal differences in the prey composition of adult female harbour porpoises, in a sample of 119 porpoises bycaught in the Swedish Kattegat and Skagerrak ﬁsheries, were interpreted by Börjesson & Berggren (1996) as indicating that habitat preferences of females could be “dictated by their association with young calves”. The absence of milk in the stomachs of six calves by-caught with their mothers in gill nets in the Bay of Fundy led Smith & Read (1992) to suggest calves are unable to nurse while their mothers are actively foraging. It is also possible that, rather than merely being a consequence of segregation, any differences in diet between the sexes could be a mechanism to reduce competition. Few differences were found between the diets of male and female porpoises in Scotland, Denmark and Holland (Santos 1998). In Scotland, male porpoises ate more sepiolids and had a higher overall prey diversity than females as well as there being some difference in 366

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

prey size. In Denmark, female porpoises had signiﬁcantly more prey in the stomach than males. In Holland, female porpoises ate signiﬁcantly more gobies than did male porpoises. This higher prey diversity in Scottish male porpoises could indicate different feeding grounds or less selectivity in the prey eaten. Female porpoises are signiﬁcantly bigger and heavier than males and the higher number of prey in females’ stomachs in Denmark could reﬂect higher energetic needs. However, this result was not found for the Scottish and Dutch porpoises. Aarefjord et al. (1995) found no signiﬁcant differences in diet between seven adult females and 48 adult males, although they recognised that the number of females examined was very low. On the other hand, signiﬁcant differences in the number of prey in the stomach were found between male and female porpoises of 1-yr old or less, with males eating more ﬁsh than females. Outside the northeast Atlantic, results suggest an absence of dietary differences between the sexes. No differences in the diet between sexes were found in a sample of 81 harbour porpoises collected from eastern Canadian waters between 1969 and 1972 (Smith & Gaskin 1974) or in a sample of 100 harbour porpoises by-caught along the northern coast of Washington State (Gearin et al. 1994). Prey weight showed no signiﬁcant differences between the sexes in a sample of 138 harbour porpoises by-caught in the Gulf of St Lawrence, Canada, in 1989 (Fontaine et al. 1994). Finally, no signiﬁcant differences were seen in the diet of males and non-lactating females in a sample of 95 harbour porpoises by-caught in the Gulf of Maine (Gannon et al. 1998). It should be noted that many studies do not distinguish between diets of lactating and non-lactating females and it seems that differences between diets of males and females are most likely to be seen when females are nursing calves.

Cause of death and diet variability A basic problem with most recent studies on porpoise diets is that they are based on dead animals. Differences in stomach contents from animals dying from different causes are difﬁcult to relate to feeding strategies. Thus, if sick animals had a diet different from healthy animals, does this difference indicate active changes in diet selection or simply a consequence of reduced mobility leading to lower prey encounter rates? Certainly, the cause of death represents a potential confounding factor and source of bias in dietary studies. If nothing else, different components of the population are represented in proportion to the frequency with which they die rather than their relative abundance in the living population. The problems arising from the use of stranded specimens in dietary analysis have been extensively discussed and reviewed elsewhere (e.g. Pierce & Boyle 1991, Sekiguchi et al. 1992). Strandings of cetaceans can be considered an “opportunistic” resource, the composition of which depends on many factors (wind and currents carrying the carcasses to the coast, accessibility of coastal locations, state of preservation, etc.). With the increase in interactions between marine mammals and ﬁsheries, by-catches have become another source of samples for dietary studies. In contrast to strandings, which could represent injured or ill individuals, by-catches may provide samples of “healthy” animals (Kuiken et al. 1994a). However, the use of by-caught individuals is not free of potential biases, notably the possible bias of the diet towards the target species of the ﬁshery and associated species (Waring et al. 1990) and the possible “net selection” of particular porpoise size and age classes. A disproportionately high number of juvenile porpoises amongst by-catches was reported in 367

M. B. SANTOS & G. J. PIERCE

Denmark (Clausen & Andersen 1988, Kinze 1994) and for the German ﬂeet ﬁshing in the Baltic and the North Sea (Kock & Benke 1996). It has been suggested that avoidance of nets could be related to experience, making young animals more vulnerable. Young animals could try to explore and play with the nets and become entrapped and this fact could also put females at risk if they try to rescue their calves (IWC 1994, Kinze 1994). Kinze also pointed out that the existence of age-related segregation in harbour porpoises would make some groups more vulnerable to by-catch. Gaskin & Blair (1977) found sub-adult males segregating from other groups in Canadian waters and staying closer to the coast and thus becoming more frequently entrapped in nets. A problem with identifying by-catches, as such, arises if by-caught animals are freed from the nets (either by the ﬁshermen or by other causes) and are found ﬂoating at sea or stranded on the coast. The diagnosis of by-catch in these animals is at present a difﬁcult task. Carcasses are often too decomposed to allow any post-mortem study to be done and some net types do not cause net marks on the skin – perhaps the clearest indication of bycatch (Kuiken 1996). At present only a small percentage of cetaceans found stranded can be diagnosed clearly as by-catches (Siebert et al. 1996). In Scotland, harbour porpoises killed by bottlenose dolphins is another source of samples available for analysis. However, these samples are also not free from potential biases. Ross & Wilson (1996) observed that, although there were no signiﬁcant differences in the number of males and females killed, there was a bias towards porpoises of 100 cm to 140 cm, which corresponded with juvenile or prepubertal animals between 1 yr and 3 yr of age. They also noted that there is seasonal variation in the total number of strandings in the Moray Firth area, with a peak value in June, and that the “injured porpoises represented a relatively constant proportion of this number, such that within each month the number of injured porpoises was signiﬁcantly correlated with the number that died of other causes”. In the northeast Atlantic, by-caught porpoises in Irish waters had eaten less clupeids and whiting than stranded porpoises but both groups had a similar proportion of gadoids in the diet (Rogan & Berrow 1996). However, the sample size in this study was small (nine and ten animals, respectively). Santos (1998) found signiﬁcant differences in the diet between porpoises killed by bottlenose dolphins and porpoises that died of other causes. Taking into consideration that the seasonal distribution of the deaths caused by bottlenose dolphins was not homogeneous, with more deaths occurring in the second and third quarters (spring and summer), the author noted that it was not surprising that porpoises killed by dolphins had taken more sandeels (since the importance of sandeels in the diet was also found to be signiﬁcantly related to season with sandeels eaten mainly in spring and summer). The differences in the importance of the other species in the diet could be related not only to seasonal but also geographical differences in abundance and/or availability. The sample of harbour porpoises killed by bottlenose dolphins came mainly from the Moray Firth and the surrounding areas. A further seven came from further south, near the Firth of Forth. In all cases harbour porpoises killed by bottlenose dolphins came exclusively from the east coast of Scotland. Signiﬁcant differences in the size of prey eaten, between porpoises killed by bottlenose dolphins and the remaining porpoises, could also be explained by the seasonal distribution of the deaths caused by dolphins. During the summer months, when more deaths take place, younger sandeels and bigger whiting are taken by harbour porpoises in bigger numbers. Differences in diet between by-caught and stranded animals have been found for other cetacean species, for example, common dolphins, dusky dolphins (Lagenorhynchus obscu368

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

rus) and Heaviside’s dolphins (Cephalorhynchus heavisidii) in South Africa (Sekiguchi et al. 1992).

Where and how do porpoises feed? Direct information on where porpoises feed comes from studies of distribution and diving behaviour. Westgate et al. (1995) recorded diving behaviour of harbour porpoises in the Bay of Fundy by attaching time-depth recorders. Typical dives were short and shallow (mean duration 44 s and mean depth 14 m), although dives to depths of up to 226 m (the maximum water depth in the area) and durations up to 321 s were recorded. Some evidence was found that dives were less frequent, but deeper, at night. Most dives were characterised as “ﬂat bottomed”, with around one-third of the dive time being bottom time, which would be consistent with foraging at the sea bed. Pierpoint et al. (1999) recorded porpoise echolocation activity on the Welsh coast using acoustic data loggers and found that porpoise activity was highest at night and during the ebb tide. Goodson (1994), in the context of discussing by-catches in set gill nets, comments that very little is known about the foraging strategies of harbour porpoises. However, records of by-catches themselves provide evidence about where and how porpoises feed, some information can be gleaned from records of surface observations, and relevant data are emerging from recent studies on echolocation behaviour. There is little doubt that porpoises often feed near the sea bottom, as indicated by several lines of evidence: the importance of sandeels in the diet, the presence of sepiolids in the diet, the characteristics of the sonar system and the fact that porpoises are often caught in bottom-set gill nets.

Surface observations Harbour porpoises seem to be gregarious and schools consist normally of few animals (nine in the Bay of Fundy, Gaskin et al. 1974) although aggregations of several hundred individuals have been reported in the literature (Fink 1959, Rae 1965). However, porpoises are believed to hunt independently rather than in groups (Read 1999). Pierpoint et al. (1994) observed porpoises in tidal races surfacing repeatedly at the same location, always orientated so as to face into the tidal stream, which they interpreted as foraging activity. The presence of gulls scavenging at the water surface supports this interpretation and, although associations between individuals were temporary, groups of up to 10 porpoises were seen. Silva et al. (1999) observed porpoises from land on the Portuguese coast during daylight hours. Numbers of sightings were highest at 09.00 and gradually declined through the day. In relation to the tidal cycle, numbers were low at both low and high tide. Maximum numbers were seen when tide height was 1 m below the height at high tide, but it is not stated whether this was during the ebb tide or ﬂood tide. Watson & Gaskin (1993) used surface observations to estimate dive times, which they record as being between 35 s and 4 min for feeding porpoises in the Bay of Fundy.

369

M. B. SANTOS & G. J. PIERCE

Echolocation in porpoises Sturtivant et al. (1994) reviewed information on porpoise echolocation. Evans (1973) suggested that sonar in porpoises is used mainly for target detection rather than target classiﬁcation owing to its monochromatic nature; Amundin et al. (1988) also noted that porpoise clicks show a narrow-band width. Hatakeyama & Soeda (1990) recorded the clicks to be mainly in the range 125–140 kHz. Prey searching involves the use of a narrow-beam, narrow-band high-frequency sonar (with a peak frequency around 130 kHz, Kastelein et al. 1999) believed by Goodson & Sturtivant (1996) to have evolved for short-range foraging, particularly near to the sea surface or the sea bottom. Kastelein et al. (1997a) also suggest that porpoise echolocation is likely to be adapted for detecting prey on the sea bottom. Their experiments showed that porpoises could detect steel and plastic discs buried up to 7 cm deep in sand and the authors’ comment that similar objects should be detectable when buried at greater depths in a muddy substratum because the substratum density would be closer to that of water. The occurrence of very small prey such as bobtail squid (Sepiolidae, e.g. Sepiola atlantica, the adults of which are no more than 2 cm in length) naturally leads to questions about how they are caught. Sepiolids are sit-and-wait predators which normally remain partially buried in the substratum. They are probably detected by porpoises directing their sonar into the substratum, and it is probably the acoustic signal from the hole rather than the animal that allows them to be detected (D. Goodson, pers. comm.)

Inferences from dietary studies Some further general inferences about feeding areas and the feeding niche can be made with reference to the ecology of the prey species. A wide variety of prey has been found in the stomachs of harbour porpoises (see previous sections), including pelagic, mesopelagic and benthic species. Prey species found in the diet of harbour porpoises are mainly small schooling ﬁsh 400 mm long, indeed in most cases 300 mm (Read 1999). It has been suggested that harbour porpoise teeth are used only to hold the prey but not to break it up into smaller pieces and, therefore, porpoises are limited as to the size of prey they can consume. In the northeast Atlantic, a mixture of mainly demersal species (whiting, cod, sandeels, Trisopterus spp., gobies) has been cited as the main prey in most cases (Lick 1991a,b, Benke & Siebert 1996, Martin 1996, Rogan & Berrow 1996, Santos 1998) although in some areas a higher proportion of pelagic prey (mainly herring) has been recorded (e.g. Berggren 1996). Aarefjord et al. (1995) found a predominance of pelagic species (herring, capelin) in the diet of harbour porpoises by-caught in Norwegian waters, while benthic prey (cod, whiting, sandeels, Pleuronectidae) predominated in the stomach contents of porpoises by-caught and stranded in Sweden and Denmark. Outside the northeast Atlantic, pelagic species such as herring, capelin, smelt (Smith & Gaskin 1974, Recchia & Read 1989, Fontaine et al. 1994, Gannon et al. 1998) have been quoted as the main prey for harbour porpoises in the Bay of Fundy, the Gulf of Maine and the Gulf of St Lawrence (eastern Canada). In the North Paciﬁc, clupeids (herring, capelin, sardine) were again recorded as the main prey for harbour porpoises (Wilke & Kenyon 1952, Scheffer 1953, Fink 1959). In Scottish waters, the two main prey types recorded in stomach contents during the 1990s were whiting and sandeels (Santos 1998) and it may be proposed that porpoises spend 370

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

a substantial amount of time in areas frequented by these species. Whiting is a demersal species living in shallow waters, usually from 39–200 m over sandy or muddy grounds. It can reach up to 70 cm standard length, although normally the size is 30–40 cm (Whitehead et al. 1989). Its distribution extends from northern Norway towards Iceland to the west and towards the northern coasts of Portugal to the south. It is also present in the Mediterranean, Aegean, Adriatic and Baltic Seas (Hislop 1972). Whiting mature at around 2 yr of age and at a modal length (for females) of 26 cm. The spawning season is extended, beginning in February in the southern North Sea and March in the northern North Sea and ending in June. Spawning normally takes place in waters less than 100 m depth (Hislop 1984). During the ﬁrst year of life, whiting are found in shallow waters, concentrating in the central and southern North Sea and in Scottish coastal waters. Most of the whiting taken by Scottish, Dutch and Danish harbour porpoises in Santos (1998) were estimated to be 23 cm in length and were therefore probably younger than 2-yr old (Hislop 1984). Concentrations of 1-yr olds are found mainly in Scottish coastal waters and in the central and southern North Sea (Hislop 1984), which would explain the greater importance of whiting in Scottish and Dutch porpoise diets than in diets of porpoises from more northern areas such as Norwegian Waters. Sandeels are a group of demersal ﬁshes which, in the northeast Atlantic, comprise three main species with very similar otoliths, Ammodytes marinus, A. tobianus and Gymnammodytes semisquamatus. Two other species, Hyperoplus immaculatus and H. lanceolatus (commonly called greater sandeels) attain a bigger size, which distinguishes them from the common or lesser sandeels, and they are also less abundant. Sandeels receive their name because of their unique way of life, spending the hours of darkness and most of the winter buried in the sand (Macer 1966, Reay 1970, Langham 1971, Winslade 1974). Because of this characteristic they are considered demersal, depending on a suitable bottom substratum to burrow, but during their activity periods they lead a pelagic life (Storey 1993). Sandeels feed on plankton and move throughout the water column during the day. Of the three main species found in the Northeast Atlantic, Ammodytes marinus is the most common, also being one of the commonest species on the continental shelf of northwest Europe and accounting for 10–15% of the total ﬁsh biomass of the North Sea (Sparholt 1990).

Food consumption by harbour porpoises Harbour porpoises have some unique characteristics among cetaceans. They are one of the smallest cetaceans and most of their range is in cold waters. Their life history includes a very short nursing period (usually less than 1 yr), sexual maturity is attained at around 3 yr of age and there is a very short resting period between pregnancies (usually females give birth each year), so that females are often pregnant and lactating at the same time (Read et al. 1997). Smeenk (pers. comm.) found that, in Dutch waters, most of the females do not give birth every year, perhaps due to unfavourable food conditions or impaired health. Harbour porpoise habitat and life history impose very high energetic demands. Furthermore, their small size means that they cannot store much energy and this makes them more dependent on a year-round proximity to food sources (Brodie 1995). According to Brodie, for harbour porpoise this dependence has “the consequence that its distribution and nutritive condition may more strongly reﬂect the distribution and energy density of its prey than for other cetaceans”. 371

M. B. SANTOS & G. J. PIERCE

Yasui & Gaskin (1986) estimated that the daily feeding rate of a non-lactating adult harbour porpoise would be 3.5% of its total body weight per day, considerably lower than values from previous studies based on food intake records from captive animals (e.g. 8.26% quoted by Sergeant 1969, based on data from Andersen 1965). However, more recent captive feeding studies tend to support the higher values. Kastelein et al. (1997b) recorded food consumption in captivity for the species (based on six individuals) to be between 4% and 9.5% of body weight. Lockyer et al. (2001) recorded food consumption to change seasonally from 7% to 9.5% of body weight for two harbour porpoises in captivity. Santos (1998) used the more conservative estimate of 3.5% and population size estimates from Hammond et al. (1995) to calculate the amount of prey removed each year by harbour porpoises in Scottish, Danish and Dutch waters. Her ﬁgures indicate that porpoises could be removing signiﬁcant amounts of several commercial ﬁsh species. For example, the estimated consumption of whiting by porpoises surpasses the landings of this species for human consumption in the North Sea. Thus, extrapolating from Scottish dietary data, harbour porpoises off Scotland and the east coast of England (SCANS blocks C, D and J) could consume around 14 640 t of whiting, 13 800 t of sandeels and 1000 t of herring per year. Off the Danish coast (SCANS blocks I and L), harbour porpoises could eat around 2880 t of herring, 6660 t of cod and 6230 t of viviparous blenny, while off the Dutch coast and west coast of Germany (SCANS blocks H and Y) porpoises could eat around 1800 t of whiting, 650 t of cod and 300 t of sandeels (assuming porpoises off the east coast of Germany to have a diet similar to the combined diet of Danish and Dutch porpoises). Finally, using combined dietary data for Scotland, Denmark and Holland, harbour porpoises in the central North Sea could eat around 3900 t of herring, 33 400 t of whiting and 14 000 t of sandeels. It should be noted that conﬁdence limits on all these estimates are wide: regardless of accuracy, the precision available is low, reﬂecting the level of uncertainty associated with sampling error, the regressions used to estimate size of ﬁsh prey eaten, the population size and the energy requirements (Santos 1998, see Santos et al. 2001b for a similar calculation for sperm whales).

Competition with other predators Comparing the harbour porpoise with other predators in the northeast Atlantic, the predator with the most similar average body size is the common (or harbour) seal, which has a range in weight in UK waters of around 45–106 kg in females and 55–130 kg in males (Corbet & Harris 1991). Several studies have been carried out on the diet of common seals in the Moray Firth (e.g. Pierce et al. 1989, 1991a,b, Thompson et al. 1991, Tollit & Thompson 1996). In general, common seal diet is dominated by sandeels in summer and other species such as gadoids (whiting, cod) and clupeids (herring, sprat) in winter. Thus, in recent studies, diets of harbour porpoise and common seals are seen to follow a similar pattern, which is consistent with both types of predators exploiting the same locally abundant resources. Tollit & Thompson (1996) also found considerable interannual variation in the diet of common seals in the Moray Firth between 1989 and 1992. In the Skagerrak and Kattegat (May–September 1988), Härkönen (1988) found that the most important species in common seal diet were cod, plaice, dab, lemon sole Microstomus kitt and sandeels. Härkö372

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

nen & Heide-Jørgensen (1991) reported the main prey of common seals in the same area in July–December 1989 to be mainly gadoids, particularly cod. Cod was also the most important species by weight in Danish harbour porpoise diets (Santos 1998), although viviparous blennies were the second most important category and very few ﬂatﬁshes were eaten. Grey seals in the Moray Firth have a similar summer diet to common seals in the same area (i.e. predominantly sandeels (Pierce et al. 1991a)). Studies elsewhere in Scotland have shown that other types of ﬁsh, particularly gadoids, dominate the winter diet (e.g. Hammond & Prime 1990, Pierce et al. 1990, Hammond et al. 1994). Again, broadly speaking, there are similarities with harbour porpoise diet. There are few data on bottlenose dolphin diet in Scottish waters, but cod, saithe and whiting were the main prey in 10 stomachs (eight from the Moray Firth) examined by Santos et al. (2001c). It has been speculated that bottlenose dolphins kill porpoises as a result of food competition (Ross & Wilson 1996) but clearly more data are needed to test this hypothesis. Apart from parallels with seals, the high importance of sandeel in harbour porpoise diets is shared with a large range of other predators, for example, whiting and cod (Daan 1989), pleuronectids and salmonids (Storey 1993), Arctic tern Sterna paradisaea (Monaghan et al. 1989), pufﬁns Fratercula arctica, guillemots Uria aalge, razorbills Alca torda (Harris & Riddiford 1989), great skuas Catharacta skua (Furness & Hislop 1981), Arctic skuas Stercorarius parasiticus, kittiwakes Rissa tridactyla (Furness 1987), shags Phalacrocorax aristotelis (Harris & Wanless 1991), minke whales Balaenoptera acutorostrata (Jonsgåard 1982) and humpback whales Megaptera novaengliae (Perkins et al. 1982).

Interactions with ﬁsheries Interactions of marine mammals with ﬁsheries are of two general types, operational and biological (Harwood & Greenwood 1985). The former include ﬁshery by-catch of porpoises and the latter include predation by porpoises on ﬁshed species. Both types of interaction arguably reﬂect diet choice, the latter more directly. The results of most of the studies on harbour porpoise diets in the northeast Atlantic show an overlap between the ﬁsh species consumed by porpoises and those targeted by ﬁsheries. This potential for competition was already noted at the beginning of last century in the North Sea when direct observations of porpoises reputedly “chasing salmon” or “playing with salmon” led ﬁshermen and naturalists to believe in the possibility of porpoises competing with local salmon ﬁsheries (e.g. Macintyre 1934, Berry 1935). Concern over the status of the Baltic salmon stock led Svärdson (1955) to propose: The relation between porpoise and salmon can be and ought to be tested by an experiment. The porpoise-hunting tradition . . . must be revived and as many of the migrating porpoises as possible caught for some years, so as to see what happens to the salmon in the Baltic. If the relation is once again positive a method has been found for conserving in the Baltic a permanent salmon population more abundant than the present one. In fact, studies from this period found no evidence of predation on salmon in porpoise stomach contents (Rae 1965, 1973, Lindroth 1962). 373

M. B. SANTOS & G. J. PIERCE

The North Sea and adjacent areas (waters west of Scotland and the Skagerrak/Kattegat area) have a long history of ﬁshery exploitation. The types of ﬁshery include pelagic and demersal ﬁsheries for human consumption, and industrial ﬁsheries (where the catch is used for reduction purposes). The pelagic ﬁshery mainly targets herring, mackerel and horse mackerel, while the demersal ﬁsheries usually catch a mixture of roundﬁsh species (e.g. cod, haddock, whiting) and/or a mixture of ﬂatﬁsh species (plaice Pleuronectes platessa and sole) with a by-catch of roundﬁsh. The industrial ﬁshery mainly takes sandeels, Norway pout and sprat, although catches also include herring, haddock and whiting (Anonymous 2002a). Whiting is the third most important species of commercial demersal ﬁsh in the North Sea. Catches of this species in the North Sea increased during the 1950s and 1960s and reached a maximum in 1969 with 200 000 t. After this date maximum landings started to decline and reached an historical low in 1998. At present the stock is considered to be outside safe biological limits (Anonymous 2002a). In recent years, a signiﬁcant part of the whiting landed is taken as by-catch in the industrial ﬁshery, mainly for Norway pout, and in addition large quantities of whiting are being discarded in favour of higher priced species. The sandeel Ammodytes marinus supports the largest single-species ﬁshery in the North Sea. Its distribution extends in the eastern North Atlantic from 74ºN to 49ºN (Channel Islands and western English Channel), including eastern Greenland, Barents Sea and the Baltic Sea (Whitehead et al. 1989). Sandeels are used for bait and food on a small scale in many areas, but the major ﬁsheries are for the production of ﬁshmeal with between 600 000–1 100 000 t being taken from the North Sea each year (mainly by Denmark) (Anonymous 2002a). Santos (1998) estimated that harbour porpoises take more whiting than are landed by ﬁsheries in the North Sea, although the whiting taken by porpoises were mainly smaller than those targeted by the ﬁshery. In Scotland, just over 99% of the whiting consumed were below the minimum landing size established for the species (27 cm), compared with 99.5% for Denmark and 70% for Holland. However, sizes of ﬁshes may be underestimated from measurements on otoliths because no correction was applied for otolith erosion. Wijnsma et al. (1999) carried out in vitro digestions of ﬁsh otoliths to estimate consequences of otoliths erosion for dietary studies on porpoises and suggested that the overall picture of diet composition for porpoises in Scottish waters was relatively robust to such errors. Discards of smaller whiting and other species in the North Sea are at present a cause for concern because the amount of whiting discarded is estimated to be equivalent to 60% of the amount landed (Anonymous 2002a) and discards are similarly high for haddock. Discards are known to be eaten by seabirds (Hudson & Furness 1989, Berghahn & Rösner 1992, Furness et al. 1992) and it is interesting to speculate whether harbour porpoises might also take advantage of this resource. Some evidence of feeding on discards exists for other cetaceans, for example, killer whales (Couperus 1994).

Long-term trends in porpoise diet Harbour porpoises in the northeast Atlantic may already have switched prey species following the decline in herring stocks to a diet based on sandeels, whiting and other gadoid species. The studies by Rae (1965, 1973) on harbour porpoise diet in Scotland between 1959 and 1971 showed clupeids (herring and sprat) to be the most frequent prey. Gadoids (mainly 374

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

whiting) were found to be second in importance in the diet, while sandeels represented a minor proportion (being identiﬁed in less than 8% of the stomachs). Although the fact that most of the animals were obtained during the winter months could explain the lack of sandeels in the diet, the importance of herring in the stomachs of harbour porpoises analysed by Rae is clearly greater than in the diet of harbour porpoises in more recent studies. The decline in herring stocks in the North Sea between the 1950s and 1970s is well documented (Cushing & Burd 1957, Burd 1978) and has been proposed as one of several hypothesis (together with organochlorine pollution, noise pollution produced by increasing boat trafﬁc and by-catches in ﬁshing gear) to explain the apparent decline in numbers of harbour porpoises in the North Sea (e.g. Verwey 1975, Otterlind 1976, Van Bree 1977, Andersen & Clausen 1983, Gaskin 1984, Kayes 1985, Smeenk 1987, IWC 1994, Kleivane et al. 1995). The collapse of the herring stocks did indeed coincide with the apparent decline in harbour porpoise numbers in the southern North Sea. However, herring was also overﬁshed on the Scottish coasts but there is no evidence of a parallel decline in porpoise numbers (Evans 1980). Furthermore, an earlier disappearance of Zuiderzee herring (a brackish water population) in Dutch waters during the 1930s, was not accompanied by a decline in harbour porpoise numbers in the area (Smeenk 1987). On the other hand, Camphuysen (1994) noted that an increase in herring stocks in recent years had been followed by a slight increase in sightings of harbour porpoises in the southeast North Sea. There has been speculation about the likelihood and consequences of porpoises switching to other prey species if their main prey were depleted by overﬁshing, because many of the prey species eaten by harbour porpoise are also commercially exploited (e.g. herring, sprat, sardine, cod, whiting, sole, sandeels) (IWC 1996). Fish stocks have shown considerable variation in abundance and distribution in the past, some of which has been the result of over-exploitation. Well known cases include the collapse of the North Sea herring stocks (Burd 1978, Corten 1990) and the massive decrease of North Sea mackerel population in the early 1970s (Cushing 1980). Not only have pelagic species shown large ﬂuctuations in abundance, landings from the Shetland sandeel ﬁshery fell during the 1980s and a parallel decline in seabird breeding success followed (Monaghan et al. 1989). Changes in the structure of ﬁsh communities due to ﬁshery exploitation have already been described for some areas (e.g. Celtic Sea, Pinnegar et al. 2002). At present, many stocks in the North Sea are outside or close to safe biological limits, with high ﬁshing mortality that is believed to be unsustainable in the longer term and spawning stock biomasses below safe levels or declining towards critical levels. Over-exploitation of herring for the human consumption ﬁshery, together with considerable by-catches of juveniles in the industrial ﬁshery in the North Sea and Skagerrak/Kattegat area, caused a rapid decline of the stock and in the 1990s emergency regulations were introduced to reduce ﬁshing mortality. The stock is at present believed to be inside safe biological limits (Anonymous 2002b) – although reported to be outside safe biological limits in 2001 (Anonymous 2001) – but stock rebuilding has been delayed by too optimistic assessments and misreporting. The state of the sprat stock is not well known with large natural ﬂuctuations in annual stock biomass (Anonymous 2001). The North Sea component of the mackerel stock is still severely depleted and considered to be in need of maximum protection (Anonymous 2002c). For the species taken in the demersal ﬁshery for human consumption, the stock of cod is considered to be outside safe biological limits and there is concern that, if the rate of ﬁshing continues, the stock will collapse. For both haddock and whiting the North Sea stocks are 375

M. B. SANTOS & G. J. PIERCE

also considered to be outside safe biological limits, with spawning stock biomass reaching an historical low in 1998 for whiting. The status of saithe stocks in the North Sea (including the Skagerrak area) and the West of Scotland is also causing concern, having ﬂuctuated around safe biological limits in recent years. Of the ﬂatﬁsh species, plaice and North Sea sole are considered to be outside safe biological limits with the historical minimum of spawning stock biomass recorded in 1997 for plaice and in 1998 for sole (Anonymous 2002a). Finally, for the industrial ﬁshery, both Norway pout and sandeel stocks are considered to be inside safe biological limits but recruitment for both species appear to be highly variable and can inﬂuence the abundance of the species rapidly due to the short life-span (Anonymous 2002a).

Consequences of diet for individual health and population status Possible causes for the decline of harbour porpoises in most European waters noted by Smeenk (1987) included the decline in herring stocks and other factors such as organochlorine pollution. Other threats to populations include ﬁshery by-catch, habitat degradation through pollution, disturbance by ship trafﬁc and boats and coastal development (Read et al. 1997). Of these threats, by-catch is arguably the most serious, the ﬁshery by-catch of harbour porpoises in northeast Atlantic waters being regarded as unsustainable by the IWC (IWC 1995). As argued above, harbour porpoises in Scottish waters have apparently already switched prey species following the decline in herring stocks to a diet based on sandeel, whiting and other gadoid species. Switching from a prey with high caloriﬁc value such as herring (Murray & Burt 1977) to one with lower caloriﬁc density could have long-term effects on survivorship and productivity. Short-term effects have been described by Dudok van Heel (1962) who observed that captive porpoises fed on young cod lost weight but this weight loss was halted when the diet was changed to the same amount of herring. Evidence of physiological changes in harbour seals, related to changes in diet composition, was found by Thompson et al. (1997). Analysing haematological parameters of a harbour seal population, they found that in years when herring and sprat dominated the diet, leukocyte counts were signiﬁcantly higher than in years when alternative prey dominated the diet. Evidence of widespread macrocytic anaemia was also found in years when an alternative prey dominated the diet. Diet is the route of entry of persistent organic pollutants and toxic elements. Of all the different types of organochlorine compounds used, only two groups (especially resistant to biodegradation) have entered the marine food chain in high concentrations and are present in marine mammal tissues. These two groups are the DDTs (dichloro-diphenyltrichloroethanes, used as pesticides in agriculture until late 1970s) and PCBs (polychlorinated biphenyls, used mainly in the electricity industry) (Aguilar & Borrell 1995). Aguilar & Borrell (1995) found that, although levels of DDT (and other contaminants such as heavy metals) in tissues were low, levels of PCBs in harbour porpoises from the eastern North Atlantic were high enough to cause concern about their possible effects on the population. Organochlorine compounds are thought to depress reproductive performance (Subramanian et al. 1987, Addison 1989) and the immune system (Wassermann et al. 1979, Brouwer et al. 1989, Vos & Luster 1989, Swart et al. 1994, Ross et al. 1996) and have been 376

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

shown in experimental studies to adversely affect mammalian reproduction (e.g. Merson & Kirkpatrick 1976, Fuller & Hobson 1986, Reijnders 1986, Boon et al. 1987, Gray et al. 1998). The harbour porpoise was the ﬁrst cetacean analysed for these compounds (Holden & Marsden 1967). Harbour porpoise, with its coastal habitat, its position at the top of the food chain and its small body size (and high metabolic rate) could be especially affected by these pollutants. Moreover, it has been suggested that the capacity of small cetaceans to metabolise certain PCB congeners is very low compared with that of birds and terrestrial mammals (Tanabe et al. 1988) and is possibly lower in harbour porpoises than in other odontocetes (Duinker et al. 1989). Jepson et al. (1999) found a signiﬁcant association between elevated blubber chlorobiphenyl concentrations and mortality due to infectious diseases in harbour porpoises from England and Wales stranded and by-caught between 1990 and 1996. However, Kuiken and co-authors found blubber chlorobiphenyl concentrations to be unrelated to adrenocortical hyperplasia in porpoises stranded and by-caught in 1990 and 1991 (Kuiken et al. 1993) or mortality due to infectious diseases in porpoises stranded and by-caught between 1989 and 1992 (Kuiken et al. 1994b). At present, while the proximate origin of PCBs in porpoise blubber is clearly to be found in the prey species, it is not clear whether particular prey species (or marine habitats) are responsible for this transfer. However, it may be noted that in general the pattern of abundance of different PCB congeners in marine mammal blubber differs markedly between ﬁsh eaters (such as the harbour porpoise) and cephalopod eaters (Wells & Mckenzie 1994). The coastal distribution of harbour porpoises makes them vulnerable to high levels of incidental ﬁshery mortality, particularly in bottom-set gill nets but also in other ﬁshing gear, e.g. salmon drift nets, pound nets for herring and salmon and mid-water trawls (e.g. Lindroth 1962, van Utrech 1978, Kinze 1990a, Northridge & Lankester 1992, Lowry & Teilmann 1994, Read 1999). Unlike dolphins, harbour porpoises seem likely to become entangled when the nets are on the sea bottom (Read & Gaskin 1988). However, it is unclear whether porpoises become entangled because they attempt to take ﬁshes already caught in the nets or if they are simply foraging in the area where nets are set and fail to detect the nets (see Tregenza et al. 1997). One possibility is that, when porpoises are echolocating ﬁshes buried in the substratum, objects in the water column (such as nets) are not detected.

Future research Most recent results on porpoise diet derive from examination of stomach contents of stranded animals. Inevitably this leads to an incomplete and potentially biased view of diet, and makes it difﬁcult to partition variation reliably. It is clear that there are regional, seasonal, sex- and size-related differences in diet and there may well be individual differences in food preferences. However, stomach contents of dead animals provide only a single snapshot of diet, with no possibility of repeated samples from the same animal. Furthermore, the most recent meal is not necessarily representative of the typical diet, especially if the animal was weakened by disease, but also if it was by-caught in a ﬁshing net as a result of pursuing a particular type of ﬁsh. Last and by no means least, stomach contents analysis is complicated by the digestive erosion of prey tissues and hard parts (see review by Pierce & Boyle 1991). 377

M. B. SANTOS & G. J. PIERCE

In other marine mammals, notably seals (e.g. Iverson et al. 1997), analysis of fatty acid composition of the blubber has allowed inferences to be made about the average diet composition of individuals over extended periods, both overcoming some of the biases of stomach contents analysis and allowing good data to be collected from animals with empty stomachs. There are both logistical and technical obstacles to be overcome before this method is used routinely on porpoises. Quantitative interpretation of fatty acid proﬁles requires extensive libraries of the fatty acid proﬁles of putative prey species. Calculating the most likely diet composition requires considerable computing power: the higher the number of possible prey species and the more fatty acids that are taken into account, the more possible combinations of different proportions of prey types need to be screened. Finding the relative importance of different prey in the diet is equivalent to solving a series of simultaneous equations, with data on each fatty acid expressed as a separate equation, each of which has a term for each prey species. For example, for three fatty acids and two prey species: I1Ca,1 I2Ca,2 Ca,obs I1Cb,1 I2Cb,2 Cb,obs where: Iy importance of prey type y in the diet Cx,y concentration of fatty acid x in the body of prey species y Cx,obs concentration of fatty acid x in porpoise blubber In these equations, all the C terms are known and it is necessary to ﬁnd the set of values for I values. Further complications are provided by variation in fatty acid proﬁles within prey species (e.g. in relation to the prey life cycle and reproductive cycle or variation in the prey species’ own diet) and differences between the overall dietary fatty acid proﬁle and that of the blubber (e.g. due to variation in assimilation, metabolism and de novo synthesis in different fatty acids). Given variation in fatty acid proﬁles of individual prey species, the set of simultaneous equations is unlikely to have an exact solution. The most probable solution could be calculated using computer simulations incorporating known variability in prey fatty acid proﬁles. The only realistic way of completely overcoming the latter problem would be to derive correction factors based on captive feeding experiments involving animals on controlled diets. However, while relatively straightforward for seals, this is unlikely to be feasible for porpoises, nor would it be ethically acceptable in some countries. To follow the diet of an individual porpoise over its lifetime, repeated biopsy samples of blubber could be taken. Individual identiﬁcation could be conﬁrmed from DNA analysis of the tissue sample. However, fatty acids of recent dietary origin are concentrated in the lower blubber, so that complete blubber cores would be needed, potentially providing a route for infection and again raising ethical questions. Regardless of whether fatty acid analysis could or should be extended to studies on live animals, it offers the most likely method for obtaining good dietary data on this species. For some large cetaceans, faecal analysis has proved to be viable, collecting material from behind a swimming animal using a net. It seems unlikely that this would be successfully for an animal as small (and generally shy of human contact) as a porpoise. Attachment of cameras (“crittercams”) has also been successful for pinnipeds and larger cetaceans and, 378

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

given a sufﬁciently small package, it should be possible to obtain a porpoise-eye view of its feeding activity. Other kinds of recording devices (e.g. time-depth recorders) have already been successfully attached to porpoises (e.g. Westgate et al. 1995). Major uncertainties about the ecological importance – and indeed the status – of porpoise populations remain due to the lack of good data on population size, especially in European waters. In the North Sea, the SCANS survey – based on a single month’s data collection in 1994 (Hammond et al. 1995) – still provides the most comprehensive picture of porpoise distribution and abundance. Good data on individual energy requirements are also required and could greatly affect present calculations of food consumption.

Acknowledgements MBS was supported by CEC Project No EVK3-CT-2000-00027 (BIOCET).

References Aarefjord, H., Bjørge, A., Kinze, C. C. & Lindstedt, I. 1995. Diet of the harbour porpoise, (Phocoena phocoena), in Scandinavian waters. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 211–222. Ackman, R. G. & Lamothe, F. 1989. Marine mammals. In Marine biogenic lipids, fats and oils. Vol. 2: Distribution of marine lipids, R. G. Ackmann (ed.). Boca Raton: CRC Press, 179–381. Addison, R. F. 1989. Organochlorines and marine mammal reproduction. Canadian Journal of Fisheries and Aquatic Sciences 46, 360–368. Aguilar, A. & Borrell, A. 1995. Pollution and harbour porpoises in the eastern North Atlantic: a review. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 231–242. Amano, M. & Miyazaki, N. 1992. Geographic variation in skulls of the harbour porpoise, Phocoena phocoena. Mammalia 56, 133–144. Amundin, M., Kallin, E. & Kallin, S. 1988. The study of the sound production apparatus in the harbour porpoise, Phocoena phocoena, and the jacobita, Cephalorhynchus commersoni, by means of serial cryo-microtome sectioning and 3-D computer graphics. In Animal sonar: processes and performance, P. Nachtigall & P. W. G. Moore (eds). New York: Plenum Press, 61–66. Andersen, L. W. 1993. The population structure of the harbour porpoise, Phocoena phocoena, in Danish waters and part of the North Atlantic. Marine Biology 116, 1–7. Andersen, S. 1965. L’alimentation du marsouin (Phocoena phocoena L.) en captivité. Vie Milieu, Series A, 16, 799–810. Andersen, S. & Clausen, B. 1983. Bycatches of the harbour porpoise, Phocoena phocoena, in Danish ﬁsheries 1980–1981, and evidence for overexploitation. International Whaling Commission, SC/35/SM14. Anonymous 2001. Report of the herring assessment working group for the area South of 62°N. International Council for the Exploration of the Sea, C. M. 2001/ACFM: 12, Copenhagen, Denmark. Anonymous 2002a. Report of the working group on the assessment of demersal stocks in the North Sea and Skagerrak. International Council for the Exploration of the Sea, C. M. 2002/ACFM: 1, Copenhagen, Denmark. Anonymous 2002b. Report of the herring assessment working group for the area South of 62°N. International Council for the Exploration of the Sea, C. M. 2002/ACFM: 12, Copenhagen, Denmark.

379

M. B. SANTOS & G. J. PIERCE

Anonymous 2002c. Report of the working group on the assessment of mackerel, horse mackerel, sardine and anchovy. International Council for the Exploration of the Sea, C. M. 2002/ACFM: 6, Copenhagen, Denmark. ASCOBANS 1994. Agreement on the conservation of small cetaceans of the Baltic and North Seas. Report from the First Meeting of the Parties, Stockholm, 26–28 September 1994. ASCOBANS Secretariat, Cambridge, UK. Baird, R. W. & Hooker, S. K. 2000. Ingestion of plastic and unusual prey by a juvenile harbour porpoise. Marine Pollution Bulletin 40, 719–720. Benke, H. & Siebert, U. 1996. The current status of harbour porpoises (Phocoena phocoena) in German waters. International Whaling Commission, SC/47/SM49, Cambridge, UK. Berggren, P. 1996. A preliminary assessment of the status of harbour porpoises (Phocoena phocoena) in the Swedish Skagerrak, Kattegat and Baltic Seas. International Whaling Commission, SC/47/SM50, Cambridge, UK. Berghahn, R. & Rösner, H.-U. 1992. A method to quantify feeding of seabirds on discard from the shrimp ﬁshery in the North Sea. Netherlands Journal of Sea Research 28, 347–350. Berrow, S. D., Long, S. C., McGarry, A. T., Pollard, D., Rogan, E. & Lockyer, C. 1998. Radionuclides (137Cs and 40K) in harbour porpoise Phocoena phocoena from British and Irish coastal waters. Marine Pollution Bulletin 36, 569–576. Berry, J. 1935. British mammals and birds as enemies of the Atlantic salmon (Salmo salar). Reports of the Avon Biological Research 1934–35, 31–64. Bjørge, A. & Donovan, G. P. (eds) 1995. Special Issue, 16: Biology of phocoenids. Cambridge: International Whaling Commission, 552 pp. Bjørge, A. & Øien, N. 1995. Distribution and abundance of harbour porpoise, Phocoena phocoena, in Norwegian waters. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 89–98. Boon, J. P., Reijnders, P. J. H., Dols, J., Wensvoort, P. & Hillebrand, M. T. J. 1987. The kinetics of individual polychlorinated biphenyl congeners in female harbour seals (Phoca vitulina), with evidence for structure-related metabolism. Aquatic Toxicology 10, 307–324. Börjesson, P. & Berggren, P. 1996. Seasonal variation in diet of harbour porpoises (Phocoena phocoena) from the Kattegat and Skagerrak Seas. In European research on cetaceans – 10, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 261 only. Börjesson, P. & Berggren, P. 1997. Morphometric comparisons of skulls of harbour porpoises (Phocoena phocoena) from the Baltic, Kattegat, and Skagerrak seas. Canadian Journal of Zoology 75, 280–287. Brodie, P. F. 1995. The Bay of Fundy/Gulf or Maine harbour porpoise (Phocoena phocoena): some considerations regarding species interactions, energetics, density dependence and bycatch. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 181–187. Brouwer, A., Reijnders, P. J. H. & Koeman, J. H. 1989. Polychlorinated biphenyl (PCB)-contaminated ﬁsh induces vitamin A and thyroid hormone deﬁciency in the common seal (Phoca vitulina). Aquatic Toxicology 15, 99–106. Burd, A. C. 1978. Long-term changes in North Sea herring stocks. Rapports et Procès-verbaux des Réunions, Conseil International pour l’Exploration de la Mer 172, 137–153. Camphuysen, C. J. 1994. The harbour porpoise, Phocoena phocoena in the southern North Sea. II: A come-back in Dutch coastal waters? Lutra 37, 54–61. Camphuysen, C. J. & Leopold, M. F. 1993. The harbour porpoise, Phocoena phocoena in the southern North Sea, particularly the Dutch sector. Lutra 36, 1–24. Clausen, B. & Andersen, S. 1988. Evaluation of bycatch and health status of the harbour porpoise (Phocoena phocoena) in Danish waters. Danish Review of Game Biology 13, 1–20. Corbet, G. B. & Harris, S. (eds) 1991. The handbook of British mammals. London: Blackwell, 3rd edn. Corten, A. 1990. Long-term trends in pelagic ﬁsh stocks of the North Sea and adjacent waters and their possible connection to hydrographic changes. Netherlands Journal of Sea Research 25, 227–235.

380

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Couperus, A. S. 1994. Killer whales (Orcinus orca) scavenging on discards of freezer trawlers north east of the Shetland islands. Aquatic Mammals 20, 47–51. Cushing, D. H. 1980. The decline of the herring stocks and the gadoid outburst. Journal du Conseil pour l’Exploration Internationale de la Mer 39, 70–81. Cushing, D. H. & Burd, A. C. 1957. On the herring of the southern North Sea. Fishery Investigations, London, Series 2 20, 1–31. Daan, N. (ed.) 1989. Data base report of the stomach sampling project 1981. ICES Cooperative Research Report, No. 164, Copenhagen, Denmark. Darling, F. F. 1947. Natural history in the Highlands and Islands. London: Collins. Desportes, G. 1985. La nutrition des odontocetes en Atlantique nord-est (cotes Francaises – iles Feroe). PhD thesis, Université de Poitiers, France. Dewhurst, H. W. 1834. The natural history of the order Cetacea. London: Dewhurst. Donovan, G. P. & Bjørge, A. 1995. Harbour porpoises in the North Atlantic: edited extract from the Report of the IWC Scientiﬁc Committee, Dublin 1995. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 3–25. Dudok van Heel, W. H. 1962. Experiments with Phocoena phocoena L. Netherlands Journal of Sea Research 1, 427–458. Duinker, J. C., Hillebrand, M. T. J., Zeinstra, T. & Boon, J. P. 1989. Individual chlorinated biphenyls and pesticides in tissues of some cetaceans species from the North Sea and the Atlantic Ocean; tissue distribution and biotransformation. Aquatic Mammals 15, 95–124. Dunnet, G. M. 1996. Aquatic predators and their prey: the take-home messages. In Aquatic predators and their prey, S. P. R. Greenstreet & M. L. Tasker (eds). Oxford: Fishing News Books, Blackwell Science, 184–186. Evans, P. G. H. 1980. Cetaceans in British waters. Mammal Review 10, 1–52. Evans, P. G. H. & Borges, L. 1995. Associations between porpoises, seabirds, and their prey in southeast Shetland, N. Scotland. In European research on cetaceans – 9, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 173–178. Evans, W. E. 1973. Echolocation by marine delphinids and one species of freshwater dolphin. Journal of the Acoustical Society of America 54, 191–199. Evans, W. E. 1975. Distribution, differentiation of populations, and other aspects of the natural history of Delphinus delphis Linnaeus in the northeast Paciﬁc Ocean. PhD thesis, University of California, Los Angeles, California. Fink, B. D. 1959. Observations of porpoise predation on a school of Paciﬁc sardines. California Fish and Game 45, 216–217. Fontaine, P. M., Hammill, P. O., Barrette, C. & Kingsley, M. C. 1994. Summer diet of the harbour porpoise (Phocoena phocoena) in the estuary and the northern Gulf of St Lawrence. Canadian Journal of Fisheries and Aquatic Sciences 51, 172–178. Fraser, F. C. 1946. Report on Cetacea stranded on the British coasts from 1933 to 1937. British Museum (Natural History) No. 12, London. Fuller, G. B. & Hobson, W. C. 1986. Effects of PCBs on reproduction in mammals. In Vol. 2: PCBs and the environment. Florida: CRC Press, 101–125. Furness, R. W. 1987. The skuas. Calton: T. & A. D. Poyser. Furness, R. W., Ensor, K. & Hudson, A. V. 1992. The use of ﬁshery waste by gull populations around the British Isles. Ardea 80, 105–113. Furness, R. W. & Hislop, J. R. G. 1981. Diets and feeding ecology of great skuas Catharacta skua during the breeding season in Shetland. Journal of Zoology, London 195, 1–23. Gannon, D. P., Craddock, J. E. & Read, A. J. 1998. Autumn food habits of harbour porpoises, Phocoena phocoena, in the Gulf of Maine. Fishery Bulletin 96, 428–437. Gaskin, D. E. 1977. Harbour porpoise Phocoena phocoena (L.) in the western approaches to the Bay of Fundy 1969–75. Reports of the International Whaling Commission 27, 487–492. Gaskin, D. E. 1984. The harbour porpoise, Phocoena phocoena (L.): regional populations, status, and

381

M. B. SANTOS & G. J. PIERCE

information on direct and indirect catches. Reports of the International Whaling Commission 34, 569–586. Gaskin, D. E., Arnold, P. W. & Blair, B. A. 1974. Phocoena phocoena. Mammalian Species 42, 1–8. Gaskin, D. E. & Blair, B. A. 1977. Age determination of harbour porpoise, Phocoena phocoena (L.), in the western North Atlantic. Canadian Journal of Zoology 55, 18–30. Gaskin, D. E. & Watson, A. P. 1985. The harbor porpoise, Phocoena phocoena, in Fish Harbor, New Brunswick, Canada: occupancy, distribution, and movements. Fishery Bulletin 83, 427–442. Gaskin, D. E., Yumamoto, S. & Kawamura, A. 1993. Harbour porpoise, Phocoena phocoena (L.), in the coastal waters of northern Japan. Fishery Bulletin 91, 440–454. Gearin, P. J., Melin, S. R., DeLong, R. L., Kajimura, H. & Johnson, M. A. 1994. Harbour porpoise interactions with a Chinook salmon set-net ﬁshery in Washington State. In Special Issue, 15: Gillnets and cetaceans, W. F. Perrin et al. (eds). Cambridge: International Whaling Commission, 427–438. Goodson, A. D. 1994. Bottom-set gillnets: problems of perception for dolphins and porpoises. In European research on cetaceans – 8, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 54–57. Goodson, A. D. & Sturtivant, C. R. 1996. Sonar characteristics of the harbour porpoise (Phocoena phocoena): source levels and spectrum. ICES Journal of Marine Science 53, 465–472. Gray, L. E., Ostby, J., Wolf, C., Lambright, C. & Kelce, W. R. 1998. The value of mechanistic studies in laboratory animals for the prediction of reproductive effects in wild life. Environmental Toxicology and Chemistry 17, 109–118. Hammond, P. S., Benke, H., Berggren, P., Borchers, D. L., Buckland, S. T., Collet, A., HeideJørgensen, M.-P., Heimlich-Boran, S., Hiby, A. R., Leopold, M. F. & Øien, N. 1995. Distribution and abundance of the harbour porpoise and other small cetaceans in the North Sea and adjacent waters. Final Report to the EU, Life 92-2/UK/027. Hammond, P. S., Hall, A. J. & Prime, J. H. 1994. The diet of grey seals around Orkney and other island and mainland sites in north-eastern Scotland. Journal of Applied Ecology 31, 340–350. Hammond, P. S. & Prime, J. H. 1990. The diet of British grey seals (Halichoerus grypus). Canadian Bulletin of Fisheries and Aquatic Sciences 222, 243–254. Hardy, A. C. 1959. The open sea: II. Fish and ﬁsheries. London: Collins. Härkönen, T. J. 1988. Food-habitat relationships of harbour seals and black cormorants in the Skagerrak and Kattegat. Journal of Zoology, London 214, 673–681. Härkönen, T. J. & Heide-Jørgensen, M.-P. 1991. The harbour seal Phoca vitulina as a predator in the Skagerrak. Ophelia 3, 191–207. Harmer, S. F. 1927. Report on Cetacea stranded on the British coasts from 1913 to 1926. British Museum (Natural History) No. 10, London. Harris, M. P. & Riddiford, N. J. 1989. The food of some young seabirds on Fair Isle in 1986–1988. Scottish Birds 15, 119–125. Harris, M. P. & Wanless, S. 1991. The importance of the lesser sandeel Ammodytes marinus in the diet of the shag Phalacrocorax aristotelis. Ornis Scandinavica 22, 375–382. Harwood, J. & Greenwood, J. J. D. 1985. Competition between British grey seals and ﬁsheries. In Marine mammals and ﬁsheries, J. R. Beddington et al. (eds). London: George Allen & Unwin, 153–169. Hatakeyama, Y. & Soeda, H. 1990. Studies on echolocation of porpoises taken in salmon gillnet ﬁsheries. In Sensory abilities of cetaceans, laboratory and ﬁeld evidence, J. A. Thomas & R. A. Kastelein (eds). New York: Plenum Press, 269–281. Heide-Jørgensen, M.-P., Mosbech, A., Teilmann, J., Benke, H. & Schultz, W. 1992. Harbour porpoise (Phocoena phocoena) densities obtained from aerial surveys north of Fyn in the bay of Kiel. Ophelia 35, 133–146. Heide-Jørgensen, M.-P., Teilmann, J., Benke, H. & Wulf, J. 1993. Abundance and distribution of harbour porpoises Phocoena phocoena in selected areas of the western Baltic and the North Sea. Helgoländer Meeresuntersuchungen 47, 335–346.

382

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Hislop, J. R. G. 1972. Some aspects of the biology of whiting, Merlangius merlangus (L.). PhD thesis, University of Aberdeen, Aberdeen, Scotland. Hislop, J. R. G. 1984. A comparison of the reproductive tactics and strategies of cod, haddock, whiting and Norway pout in the North Sea. In Fish reproduction: strategies and tactics, G. W. Potts & R. J. Wootton (eds). London: Academic Press, 311–329. Hislop, J. R. G. 1996. Changes in North Sea gadoid stocks. ICES Journal of Marine Science 53, 1146–1156. Hobson, K. A. 1990. Stable isotope analysis of marbled murrelets: evidence for freshwater feeding and determination of trophic level. Condor 92, 897–903. Hohn, A. A. & Brownell, R. L. 1990. Harbour porpoise in central Californian waters: life history and incidental catches. International Whaling Commission, SC/42/SM47, Cambridge, UK. Holden, A. V. & Marsden, K. 1967. Organochlorine pesticides in seals and porpoises. Nature 216, 1274–1276. Hooker, S. K., Iverson, S. J., Ostrom, P. & Smith, S. C. 2001. Diet of northern bottlenose whales inferred from fatty-acid and stable-isotope analyses of biopsy samples. Canadian Journal of Zoology 79, 1442–1454. Hudson, A. V. & Furness, R. W. 1989. The behaviour of seabirds foraging at ﬁshing boats around Shetland. Ibis 131, 225–237. Iverson, S. J., Frost, K. J. & Lowry, L. F. 1997. Fatty acid signatures reveal ﬁne scale structure of foraging distribution of harbor seals and their prey in Prince William Sound, Alaska. Marine Ecology Progress Series 151, 255–271. Iverson, S. J., Oftedal, O. T., Bowen, W. D., Boness, D. J. & Sampugna, J. 1995. Prenatal and postnatal transfer of fatty acids from mother to pup in the hooded seal. Journal of Comparative Physiology, B 165, 1–12. IWC 1991. Report of the scientiﬁc committee. Annex G. Report of the sub-committee on small cetaceans. Reports of the International Whaling Commission 41, 172–188. IWC 1992. Report of the scientiﬁc committee. Annex G. Report of the sub-committee on small cetaceans. Reports of the International Whaling Commission 42, 178–234. IWC 1994. Report of the workshop on mortality of cetaceans in passive ﬁshing nets and traps. In Special Issue, 15: Gillnets and cetaceans, W. F. Perrin et al. (eds). Cambridge: International Whaling Commission, 1–71. IWC 1995. Report of the scientiﬁc committee. Annex G. Report of the sub-committee on small cetaceans. Reports of the International Whaling Commission 45, 165–186. IWC 1996. Report of the scientiﬁc committee. Annex H. Report of the sub-committee on small cetaceans. Reports of the International Whaling Commission 46, 160–179. Jepson, P. D., Bennett, P. M., Allchin, C. R., Law, R. J., Kuiken, T., Baker, J. R., Rogan, E. & Kirkwood, J. K. 1999. Investigating potential associations between chronic exposure to polychlorinated biphenyls and infectious disease mortality in harbour porpoises from England and Wales. The Science of the Total Environment 243/244, 339–348. Jonsgåard, A. 1982. The food of minke whales (Balaenoptera acutorostrata) in the northern North Atlantic waters. Reports of the International Whaling Commission 32, 259–262. Kastelein, R. A., Au, W. W. L., Rippe, H. T. & Schooneman, N. M. 1999. Target detection by an echolocating harbour porpoise (Phocoena phocoena). Journal of the Acoustical Society of America 105, 2493–2498. Kastelein, R. A., Hardeman, J. & Boer, H. 1997b. Food consumption and body weight of harbour porpoises (Phocoena phocoena). In The biology of harbour porpoise, A. J. Read et al. (eds). Woerden, The Netherlands: De Spil Publishers, 217–233. Kastelein, R. A. & Lavaleije, M. S. S. 1992. Foreign bodies in the stomachs of a female harbour porpoise (Phocoena phocoena) from the North Sea. Aquatic Mammals 18, 40–46. Kastelein, R. A., Schooneman, N. M., Au, W. W. L., Verboom, W. C. & Vaughan, N. 1997a. The ability of a harbour porpoise (Phocoena phocoena) to discriminate between objects buried in the

383

M. B. SANTOS & G. J. PIERCE

sand. In The biology of harbour porpoise, A. J. Read et al. (eds). Woerden, The Netherlands: De Spil Publishers, 329–342. Kayes, R. J. 1985. The decline of porpoises and dolphins in the southern North Sea: a current status report. Greenpeace International, Political Ecology Research Group, Oxford. Kinze, C. C. 1985. Intraspeciﬁc variation in Baltic and North Sea harbour porpoises (Phocoena phocoena (L., 1758)). Videnskabelige Meddelelser fra Dansk Naturhistorie Forening, 146, 63–74. Kinze, C. C. 1989. On the reproduction, diet and parasitic burden of harbour porpoise (Phocoena phocoena) in the west Greenland waters. In European research on cetaceans – 3, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 91–95. Kinze, C. C. 1990a. Cetacean mortality in passive ﬁshing nets and traps in the Baltic Sea: a review. International Whaling Commission, SC/O90/G25, Cambridge, UK. Kinze, C. C. 1990b. Marsvineundersøgelser 1989, afsluttende rapport. Zoologisk Museum, Copenhagen, Denmark. Kinze, C. C. 1994. Incidental catch of harbour porpoises (Phocoena phocoena) in Danish waters, 1986–1989. In Special Issue, 15: Gillnets and cetaceans, W. F. Perrin et al. (eds). Cambridge: International Whaling Commission, 183–187. Kinze, C. C. 1995. Exploitation of harbour porpoises (Phocoena phocoena) in Danish waters: a historical review. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 141–153. Kirsch, P. E., Iverson, S. J. & Bowen, W. D. 2000. Effects of diet on body composition and blubber fatty acids in captive harp seals (Phoca groenlandica). Physiological and Biochemical Zoology 73, 45–59. Kirsch, P. E., Iverson, S. J., Bowen, W. D., Kerr, S. R. & Ackman, R. G. 1998. Dietary effects on the fatty acid signature of whole Atlantic cod (Gadus morhua). Canadian Journal of Fisheries and Aquatic Sciences 55, 1378–1386. Kleivane, L., Skaare, J. U., Bjorke, A., Ruiter, E. de & Reijnders, P. J. H. 1995. Organochlorine pesticide residue and PCBs in harbour porpoise (Phocoena phocoena) incidentally caught in Scandinavian waters. Environmental Pollution 89, 137–146. Kock, K. H. & Benke, H. 1996. On the by-catch of harbour porpoise (Phocoena phocoena) in German ﬁsheries in the Baltic and the North Sea. Archives of Fishery and Marine Research 44, 95–114. Koopman, H. N. 1998. Topographical distribution of the blubber of harbour porpoises (Phocoena phocoena). Journal of Mammalogy 79, 260–270. Koopman, H. N., Iverson, S. J. & Gaskin, D. E. 1996. Stratiﬁcation and age-related differences in blubber fatty acids of the male harbour porpoise (Phocoena phocoena). Journal of Comparative Physiology, B 165, 628–639. Kuiken, T. 1996. Review of the criteria for the diagnosis of by-catch in cetaceans. In Newsletter 26 (Special Issue): Diagnosis of by-catch in cetaceans, Proceedings of the Second ECS Workshop on Cetacean Pathology, T. Kuiken (ed.). Saskatoon, Canada: European Cetacean Society, 38–43. Kuiken, T., Bennett, P. M., Allchin, C. R., Kirkwood, J. K., Baker, J. R., Lockyer, C. H., Walton, M. J. & Sheldrick, M. C. 1994b. PCBs, cause of death and body condition in harbour porpoises (Phocoena phocoena) from British waters. Aquatic Toxicology 28, 13–28. Kuiken, T., Hoeﬂe, U., Bennett, P. M., Allchin, C. R., Kirkwood, J. K., Baker, J. R., Appleby, E. C., Lockyer, C. H., Walton, M. J. & Sheldrick, M. C. 1993. Adrenocortical hyperplasia, disease and chlorinated hydrocarbons in the harbour porpoise (Phocoena phocoena). Marine Pollution Bulletin 26, 440–446. Kuiken, T., Simpson, V. R., Allchin, C. R., Bennett, P. M., Codd, G. A., Harris, E. A., Howes, G. J., Kennedy, S., Kirkwood, J. K., Law, R. J., Merrett, N. R. & Phillips, S. 1994a. Mass mortality of common dolphins (Delphinus delphis) in south west England due to incidental capture in ﬁshing gear. Veterinary Record 134, 81–89. Langham, L. P. E. 1971. The distribution and abundance of larval sand-eels (Ammodytidae) in Scottish waters. Journal of the Marine Biological Association of the United Kingdom 51, 697–707.

384

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Lick, R. R. 1991a. Untersuchungen zu Lebenszyklus (Krebse – Fische – marine Säuger) und Gefrierresistentz anisakider Nematoden in Nord- und Ostsee [Investigations concerning the life cycle (crustaceans – ﬁsh – marine mammals) and freezing tolerance of anisakine nematodes in the North Sea and the Baltic Sea]. PhD thesis, University of Kiel, Kiel, Germany. Lick, R. R. 1991b. Parasites from the digestive tract and food analysis of harbour porpoise (Phocoena phocoena) from German waters. In European research on cetaceans – 5, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 65–68. Lindroth, A. 1962. Baltic salmon ﬂuctuations 2: porpoise and salmon. Reports of the Institute of Freshwater Research of Drottningholm 44, 105–112. Lockyer, C. 1995. Investigation of aspects of the life history of the harbour porpoise, Phocoena phocoena, in British waters. In Special Issue, 16: Biology of phocoenids, A. Bjørge & G. P. Donovan (eds). Cambridge: International Whaling Commission, 189–197. Lockyer, C. 1999. Application of a new method to investigate population structure in the harbour porpoise, Phocoena phocoena, with special reference to the North and Baltic Seas. Journal of Cetacean Research and Management 1, 297–304. Lockyer, C., Desportes, G., Anderson, K., Labberté, S. & Siebert, U. 2001. Monitoring growth of harbour porpoise (Phocoena phocoena) in human care. International Council for the Exploration of the Sea, C. M. 2001/J:29, Copenhagen, Denmark. Lowry, N. & Teilmann, J. 1994. Bycatch and bycatch reduction of the harbour porpoise (Phocoena phocoena) in Danish waters. In Special Issue, 15: Gillnets and cetaceans, W. F. Perrin et al. (eds). Cambridge: International Whaling Commission, 203–209. Macer, C. T. 1966. Sandeels (Ammodytidae) in the southwestern North Sea: their biology and ﬁshery. Fishery Investigations, London, Series 2 24, 1–55. Macintyre, D. 1934. Sea enemies of salmon. Salmon and Trout Magazine 74, 38–42. Malinga, M. & Kuklik, I. 1996. Food consumption of harbour porpoises (Phocoena phocoena) in Polish waters of the Baltic Sea. In European research on cetaceans – 10, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 260 only. Marchessaux, D. 1980. A review of the current knowledge of the cetaceans in the eastern Mediterranean Sea. Vie Marina 2, 59–64. Martin, A. R. 1996. The diet of harbour porpoises (Phocoena phocoena) in British waters. International Whaling Commission, SC/47/SM48, Cambridge, UK. Matthews, L. H. 1952. British mammals. London: Collins. Merson, M. H. & Kirkpatrick, R. L. 1976. Reproductive performance of captive white-footed mice fed a PCB. Bulletin of Environmental Contamination and Toxicology 16, 392–398. Millais, J. G. 1906. Mammals of Great Britain and Ireland. London: Longmans Green. Miyazaki, N., Amano, M. & Fujise, Y. 1987. Growth and skull morphology of the harbour porpoises in the Japanese waters. Memoirs of the Natural Science Museum of Tokyo 20, 137–146. Monaghan, P., Uttley, J. D., Burns, M. D., Thaine, C. & Blackwood, J. 1989. The relationship between food supply, reproductive effort and breeding success in Arctic terns Sterna paradisaea. Journal of Animal Ecology 58, 261–274. Murray, J. & Burt, J. R. 1977. The composition of ﬁsh. Torry Advisory Note No. 38, MAFF, Aberdeen. Northridge, S. P. & Lankester, K. 1992. Sightings of the harbour porpoise in the North Sea and some notes on interactions with the ﬁsheries. International Whaling Commission, SC/42/SM46, Cambridge, UK. Northridge, S. P., Tasker, M. L., Webb, A. & Williams, J. M. 1995. Distribution and relative abundance of harbour porpoises (Phocoena phocoena L.), white-beaked dolphins (Lagenorhynchus albirostris Gray) and minke whales (Balaenoptera acutorostrata Lacepède) around the British Isles. ICES Journal of Marine Science 52, 55–66. Otterlind, G. 1976. The harbour porpoise (Phocoena phocoena) endangered in Swedish waters. International Council for the Exploration of the Sea, C. M. 1976/N:16, Copenhagen, Denmark. Palka, D. L., Read, A. J., Westgate, A. J. & Johnston, D. W. 1996. Summary of current knowledge of

385

M. B. SANTOS & G. J. PIERCE

harbour porpoises in US and Canadian Atlantic waters. Reports of the International Whaling Commission 46, 559–565. Pauly, D., Trites, A. W., Capuli, E. & Christensen, V. 1998. Diet composition and trophic levels of marine mammals. ICES Journal of Marine Science 55, 467–481. Perkins, J. S., Bryant, P. J., Nichols, G. & Patten, D. R. 1982. Humpback whales (Megaptera novaeangliae) off the west coast of Greenland. Canadian Journal of Zoology 60, 2921–2930. Pierce, G. J. & Boyle, P. R. 1991. A review of methods for diet analysis in piscivorous marine mammals. Oceanography and Marine Biology: an Annual Review 29, 409–486. Pierce, G. J., Boyle, P. R., Diack., J. S. W. & Clark, I. 1990. Sandeels in the diets of seals: application of novel and conventional methods of analysis to faeces from seals in the Moray Firth area of Scotland. Journal of the Marine Biological Association of the United Kingdom 70, 829–840. Pierce, G. J., Diak, J. S. W. & Boyle, P. R. 1989. Digestive tract contents of seals in the Moray Firth area of Scotland. Journal of Fish Biology 35 (Suppl. A), 341–343. Pierce, G. J., Miller, A., Thompson, P. M. & Hislop, J. R. G. 1991a. Prey remains in grey seal (Halichoerus grypus) faeces from the Moray Firth, north-east Scotland. Journal of Zoology, London 224, 337–341. Pierce, G. J., Thompson, P. M., Miller, A., Diack, J. S. W., Miller, D. & Boyle, P. R. 1991b. Seasonal variation in the diet of common seals (Phoca vitulina) in the Moray Firth area of Scotland. Journal of Zoology, London 223, 641–652. Pierpoint, C., Earl, S. & Baines, M. 1994. Observations of harbour porpoise in West Wales, 1993. In European research on cetaceans – 8, P.G.H. Evans (ed.). Cambridge: European Cetacean Society, 75–78. Pierpoint, C. J. L., Baines, M. E., Earl, S. J., Harris, R. & Tregenza, N. 1999. Night-life of harbour porpoise. In European research on cetaceans – 13, P. G. H. Evans et al. (eds). Cambridge: European Cetacean Society, 70–71. Pinnegar, J. K., Jennings, S., O’Brien, C. M. & Polunin, N. V. C. 2002. Long-term changes in the trophic level of the Celtic Sea ﬁsh community and ﬁsh market place distribution. Journal of Applied Ecology 39, 377–390. Pulliam, H. R. 1974. On the theory of optimal diets. American Naturalist 108, 59–74. Rae, B. B. 1965. The food of the common porpoise (Phocoena phocoena). Journal of Zoology, London 146, 114–122. Rae, B. B. 1973. Additional notes on the food of the common porpoise (Phocoena phocoena). Journal of Zoology, London 169, 127–131. Read, A. J. 1999. The harbour porpoise – Phocoena phocoena (Linnaeus, 1758). In Handbook of marine mammals Vol. 6: The second book of dolphins and the porpoises, S. H. Ridgway & R. Harrison (eds). London: Academic Press, 323–355. Read, A. J. & Gaskin, D. E. 1988. Incidental catch of harbour porpoises by gill nets. Journal of Wildlife Management 52, 517–523. Read, A. J. & Gaskin, D. E. 1990. Changes in growth and reproduction of harbour porpoises, Phocoena phocoena, from the Bay of Fundy. Canadian Journal of Fisheries and Aquatic Sciences 47, 2158–2163. Read, A. J. & Hohn, A. A. 1995. Life in the fast lane: the life history of harbour porpoises from the Gulf of Maine. Marine Mammal Science 11, 423–440. Read, A. J. & Westgate, A. J. 1997. Monitoring the movements of harbour porpoises (Phocoena phocoena) with satellite telemetry. Marine Biology 130, 315–322. Read, A. J., Wiepkema, P. R. & Nachtigall, P. E. 1997. The harbour porpoise (Phocoena phocoena). In The biology of the harbour porpoise, A. J. Read et al. (eds). Woerden, The Netherlands: De Spil Publishers, 3–6. Reay, P. J. 1970. Synopsis of biological data on north Atlantic sandeels of the genus Ammodytes (A. tobianus, A. dubius, A. americanus and A. marinus). FAO Fisheries Synopsis 82. Recchia, C. A. & Read, A. J. 1989. Stomach contents of harbour porpoises, Phocoena phocoena (L.), from the Bay of Fundy. Canadian Journal of Zoology 67, 2140–2146.

386

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Reijnders, P. J. H. 1986. Reproductive failure in common seals feeding on ﬁsh from polluted coastal waters. Nature 324, 456–457. Rogan, E. & Berrow, S. D. 1996. A review of harbour porpoises, Phocoena phocoena, in Irish waters. Reports of the International Whaling Commission 46, 595–605. Rosel, P. E., Dizon, A. E. & Haygood, M. G. 1995. Variability of the mitochondrial control region in populations of the harbour porpoise, Phocoena phocoena, on interoceanic and regional scales. Canadian Journal of Fisheries and Aquatic Sciences 52, 1210–1219. Rosel, P. E., France, S. C., Wang, J. Y. & Kocher, T. D. 1999. Genetic structure of harbour porpoise Phocoena phocoena populations in the northwest Atlantic based on mitochondrial and nuclear markers. Molecular Ecology 8, S41–S54. Ross, G. J. B. 1984. The smaller cetaceans of the south east coast of southern Africa. Annals of the Cape Provincial Museums (Natural History) 15, 173–410. Ross, H. M. & Wilson, B. 1996. Violent interactions between bottlenose dolphins and harbour porpoises. Proceedings of the Royal Society of London, B 263, 283–286. Ross, P. S., De Swart, R. L., Timmerman, H. H., Reijnders, P. J. H., Vos, J. G., Van Loveren, H. & Osterhaus, A. D. M. E. 1996. Suppression of natural killer cell activity in harbour seals (Phoca vitulina) fed Baltic Sea herring. Aquatic Toxicology 34, 71–84. Santos, M. B. 1998. Feeding ecology of harbour porpoises, common and bottlenose dolphins and sperm whales in the northeast Atlantic. PhD thesis, University of Aberdeen, Aberdeen, Scotland. Santos, M. B., Clarke, M. R. & Pierce, G. J. 2001b. Assessing the importance of cephalopods in the diets of marine mammals and other top predators. Fisheries Research, 52, 121–139. Santos, M. B., Pierce, G. J., Reid, R. J., Patterson, I. A. P., Ross, H. M. & Mente, E. 2001c. Stomach contents of bottlenose dolphins (Tursiops truncatus) in Scottish waters. Journal of the Marine Biological Association of the United Kingdom 81, 873–878. Santos, M. B., Spencer, N., Pierce, G. J., Lockyer, C., Reid, R. J., Patterson, I. A. P. & Edwards, W. 2001a. Life history parameters of harbour porpoise (Phocoena phocoena) in Scottish waters. 15th Annual Conference of the European Cetacean Society, Rome, Italy. Scheffer, V. B. 1953. Measurements and stomachs contents of 11 delphinids from the northeast Paciﬁc. Murrelet 34, 27–30. Scott, T. 1903. Some further observations on the food of ﬁshes, with a note on the food observed in the stomach of a common porpoise. Report of the Fishery Board for Scotland 21, 218–227. Sekiguchi, K. 1987. Occurrence and behaviour of harbour porpoises (Phocoena phocoena) at Pajaro Dunes, Monterey Bay, California. MSc thesis, San Jose State University, California. Sekiguchi, K., Klages, N.T.W. & Best, P.B. 1992. Comparative analysis of the diets of smaller odontocete cetaceans along the coast of Southern Africa. South African Journal of Marine Science 12, 843–861. Sequeira, M. 1996. Harbour porpoises, Phocoena phocoena, in Portuguese waters. Reports of the International Whaling Commission 46, 583–586. Sergeant, D. E. 1969. Feeding rates of Cetacea. Fiskeridirektorates Skrifter, Serie Havundersøkelser 15, 246–258. Sergeant, D. E. & Fisher, H. D. 1957. The smaller Cetacea of eastern Canadian waters. Journal of the Fisheries Research Board of Canada 14, 83–115. Siebert, U., Benke, H., Frese, K., Pirro, F. & Lick, R. 1996. Postmortem examination of by-catches from German ﬁsheries and of suspected by-catches found on the coast of Germany. In Newsletter 26 (Special Issue): Diagnosis of by-catch in cetaceans, Proceedings of the Second ECS Workshop on Cetacean Pathology, T. Kuiken (ed.). Saskatoon, Canada: European Cetacean Society, 27–30. Silva, M. A., Sequeira, M., Prieto, R. and Alexandre, B. 1999. Observations of harbour porpoises (Phocoena phocoena) on the northern coast of Portugal. In European research on cetaceans – 13, P. G. H. Evans et al. (eds). Cambridge: European Cetacean Society, 267–269. Slijper, E. J. 1962. Whales. London: Hutchinson. Smeenk, C. 1987. The harbour porpoise Phocoena phocoena (L., 1758) in the Netherlands: stranding records and decline. Lutra 30, 77–90.

387

M. B. SANTOS & G. J. PIERCE

Smeenk, C., Leopold, M. F. & Addink, M. J. 1992. Note on the harbour porpoise Phocoena phocoena in Mauritania, West Africa. Lutra 35, 98–104. Smith, G. J. D. & Gaskin, D. E. 1974. The diet of harbour porpoises (Phocoena phocoena (L.)) in coastal waters of Eastern Canada, with special reference to the Bay of Fundy. Canadian Journal of Zoology 52, 777–782. Smith, R. J. & Read, A. J. 1992. Consumption of euphausiids by harbour porpoises (Phocoena phocoena) calves in the Bay of Fundy. Canadian Journal of Zoology 70, 1629–1632. Southwell, T. 1881. The seals and whales of the British seas. London: Jarrold. Sparholt, H. 1990. An estimate of the total biomass of ﬁsh in the North Sea. ICES Journal of Marine Science 46, 200–210. Stephen, A. C. 1926. Common porpoise stranded at Granton. Scottish Naturalist 1926, 46. Storey, G. 1993. The trophic ecology of the sandeel Ammodytes tobianus (L.), and other planktivorous ﬁsh species from Aberdeen Bay. PhD thesis, University of Aberdeen, Aberdeen. Sturtivant, C. R., Datta, S. & Goodson, D. S. 1994. A review of echolocation research on the harbour porpoise Phocoena phocoena and the common dolphin Delphinus delphis. In European research on cetaceans – 8, P. G. H. Evans (ed.). Cambridge: European Cetacean Society, 164–168. Subramanian, A. N., Tanabe, S., Tatsukawa, R., Saito, S. & Miyazaki, N. 1987. Reduction in the testosterone levels by PCBs and DDE in Dall’s porpoises of northwestern North Paciﬁc. Marine Pollution Bulletin 18, 643–646. Svärdson, G. 1955. Salmon stock ﬂuctuations in the Baltic Sea. Reports of the Institute of Freshwater Research of Drottningholm 36, 226–262. Swart, R. de L., Ross, P. S., Vedder, L. J., Timmerman, H. H., Heisterkamp, S., Loveren, H. V., Vos, J. G., Reijnders, P. J. H. & Osterhaus, A. D. M. E. 1994. Impairment of immune function in harbour seals (Phoca vitulina) feeding on ﬁsh from polluted waters. Ambio 23, 155–159. Tanabe, S., Watanabe, S., Kan, H. & Tatsukawa, R. 1988. Capacity and mode of PCB metabolism in small cetaceans. Marine Mammal Science 4, 103–124. Taylor, B. L. & Dawson, P. K. 1984. Seasonal changes in density and behaviour of harbour porpoise (Phocoena phocoena) affecting census methodology in Glacier Bay National Park, Alaska. Reports of the International Whaling Commission 34, 479–483. Teilmann, J. & Dietz, R. 1996. Status of the harbour porpoises (Phocoena phocoena) in Greenland. International Whaling Commission, SC/47/SM44, Cambridge, UK. Thompson, P. M., Pierce, G. J., Hislop, J. R. G., Miller, D. & Diack, J. S. W. 1991. Winter foraging by common seals (Phoca vitulina) in relation to food availability in the inner Moray Firth, N. E. Scotland. Journal of Animal Ecology 60, 283–294. Thompson, P. M., Tollit, D. J., Corpe, H. M., Reid, R. J. & Ross, H. M. 1997. Changes in haematological parameters in relation to prey switching in a wild population of harbour seals. Functional Ecology 11, 743–750. Tiedemann, R., Harder, J., Gmeiner, C. & Haase, E. 1996. Mitochondrial DNA sequence patterns of harbour porpoises (Phocoena phocoena) from the North and the Baltic Sea. Zeitschrift für Säugetierkunde 61, 104–111. Tiezen, L. L. 1978. Carbon isotope fractionation in biological material. Nature 276, 97–98. Tolley, K.A. & Heldal, H.E. 2002. Inferring ecological separation from regional differences in radioactive caesium in harbour porpoises (Phocoena phocoena). Marine Ecology Progress Series 228, 301–309. Tolley, K. A., Rosel, P. E., Walton, M., Bjørge, A. & Øien, N. 1999. Genetic population structure of harbour porpoises (Phocoena phocoena) in the North Sea and Norwegian waters. Journal of Cetacean Research and Management 1, 265–274. Tolley, K. A., Víkingsson, G. A. & Rosel, P. E. 2001. Mitochondrial DNA sequence variation and phylogeographic patterns in harbour porpoises Phocoena phocoena from the North Atlantic. Conservation Genetics 2, 349–361. Tollit, D. J. & Thompson, P. M. 1996. Seasonal and between-year variations in the diet of harbour seals in the Moray Firth, Scotland. Canadian Journal of Zoology 74, 1110–1121.

388

THE DIET OF HARBOUR PORPOISE IN THE NE ATLANTIC

Tomilin, A. G. 1957. Vol. IX: Cetacea. In Mammals of the U.S.S.R. and adjacent countries, V. G. Geptner (ed.). Moscow: Izdat. Akad. Nauk SSSR. (English Translation, 1967, Jerusalem: Israel Program for Scientiﬁc Translations). Tregenza, N. J. C., Berrow, S. D., Leaper, R. & Hammond, P. S. 1997. Harbour porpoise, Phocoena phocoena L., by-catch in set gill nets in the Celtic Sea. ICES Journal of Marine Science 54, 896–904. van Beneden, P. J. 1889. Histoire naturelle des delphinides des mers d’Europe. Mémoirs du Course de l’Academie Royal Belge 43, 1–253. van Bree, P. J. H. 1977. On former and recent strandings of cetaceans on the coast of the Netherlands. Zeitschrift fur Saugetierkunde 42, 101–107. van Utrech, W. L. 1978. Age and growth in Phocoena phocoena Linnaeus, 1758 (Cetacea, Odontoceti) from the North Sea. Bijdragen tot de Dierkunde 48, 16–28. Verwey, J. 1975. The cetaceans Phocoena phocoena and Tursiops truncatus in the Marsdiep area (Dutch Wadden Sea) in the years 1931–1973. Nederlands Instituut voor Onderzoek der Zee, Publicaties en verslagen, 1975 17, 1–153. Vinther, M. 1999. Bycatches of harbour porpoises (Phocoena phocoena L.) in Danish set-net ﬁsheries. Journal of Cetacean Research and Management 1, 123–135. Vos, J. G. & Luster, M. I. 1989. Immune alterations. In Halogenated biphenyls, terphenyls, naphthalenes, dibenzodioxins and related products, R. D. Kimbrough & A. A. Jensen (eds). Amsterdam: Elsevier, 295–322. Walton, M. J. 1997. The population structure of harbour porpoises Phocoena phocoena in the seas around the UK and adjacent waters. Proceedings of the Royal Society of London, B 264, 89–94. Wang, J. Y., Gaskin, D. E. & White, B. N. 1996. Mitochondrial DNA analysis of harbour porpoise, Phocoena phocoena, subpopulations in North American waters. Canadian Journal of Fisheries and Aquatic Sciences 53, 1632–1645. Wang, J. Y. & Berggren, P. 1997. Mitochondrial DNA analysis of harbour porpoises (Phocoena phocoena) in the Baltic Sea, the Kattegat–Skagerrak Seas and off the west coast of Norway. Marine Biology 127, 531–537. Wang, J. Y., Gaskin, D. E. & White, B. N. 1996. Mitochondrial DNA analysis of harbour porpoise, Phocoena phoconea, subpopulations in North American waters. Canadian Journal of Fisheries and Aquatic Sciences 5, 1632–1645. Waring, G. T., Gerrior, P., Payne, P. M., Parry, B. L. & Nicolas, J. R. 1990. Incidental take of marine mammals in foreign ﬁshery activities off the northeast United States, 1977–88. Fishery Bulletin 88, 347–360. Wassermann, M., Wassermann, D., Gershon, Z. & Zellermayer, L. 1979. Effects of organochlorine insecticides on body defense systems. In Biological effects of pesticides in mammalian systems, H. F. Kraybill (ed.). New York: Annals of the New York Academy of Sciences, 393–401. Watson, A. P. & Gaskin, D. E. 1983. Observations on the ventilation cycle of the harbour porpoise Phocoena phocoena in coastal waters of the Bay of Fundy. Canadian Journal of Zoology 61, 126–132. Watson, L. 1985. Whales of the world. London: Hutchinson. Wells, D. E. & McKenzie, C. 1994. Techniques for pattern recognition of organochlorine residues in sea mammals from Scottish coastal waters. International Council for the Exploration of the Sea, C. M. 1994/E N:10, Copenhagen, Denmark. Westgate, A. J., Read, A. J., Berggren, P., Koopman, H. N. & Gaskin, D. E. 1995. Diving behaviour of harbour porpoises, Phocoena phocoena. Canadian Journal of Fisheries and Aquatic Sciences 52, 1064–1073. Westgate, A. J. & Tolley, K. A. 1999. Geographical differences in organochlorine contaminants in harbour porpoises Phocoena phocoena from the western North Atlantic. Marine Ecology Progress Series 177, 255–268. Whitehead, P. J. P., Bauchot, M.-L., Hureau, J.-C., Nielsen, J. & Tortonese, E. (eds) 1989. Fishes of the north-eastern Atlantic and the Mediterranean. UNESCO, Paris.

389

M. B. SANTOS & G. J. PIERCE

Whymper, F. 1883. The ﬁsheries of the world. An illustrated and descriptive record of the international ﬁsheries exhibition. London: Cassell. Wijnsma, G., Pierce, G. J. & Santos, M. B. 1999. Assessment of errors in cetacean diet analysis: in vitro digestion of otoliths. Journal of the Marine Biological Association of the United Kingdom 79, 573–575. Wilke, F. & Kenyon, K. W. 1952. Notes on the food of fur seal, sea-lion, and harbour porpoise. Journal of Wildlife Management 16, 396–397. Winslade, P. 1974. Behavioural studies on the lesser sandeel Ammodytes marinus (Raitt). II. The effect of light intensity on activity. Journal of Fish Biology 6, 577–586. 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 & M. L. Tasker (eds). Oxford: Fishing News Books, Blackwell Science, 154–165. Yasui, W. C. & Gaskin, D. E. 1986. Energy budget of a small cetacean, the harbour porpoise, Phocoena phocoena (L.). Ophelia 25, 183–197. Yurick, D. B. 1977. Populations, subpopulations and zoogeography of the harbour porpoise, Phocoena phocoena (L.). MSc thesis, University of Guelph, Guelph, Ontario. Yurick, D. B. & Gaskin, D. E. 1987. Morphometric and meristic comparisons of skulls of harbour porpoise Phocoena phocoena (L.) from the North Atlantic and North Paciﬁc. Ophelia 27, 53–75.

390

AUTHOR INDEX

References to complete articles are given in bold type; references to bibliographical lists are given in italics; references to pages are given in normal type. Aarefjord, H., 360, 361, 363, 366, 367, 370, 379 Abele, D. See Fabricius, K.E., 305 Abele, L.G., 85, 97 Abell, D.H., 130, 147 Ackerman, J., 302, 303 Ackman, R.G., 359, 379; See Kirsch, P.E., 384 Adams, P.B., 23, 34 Adams, S.M., 118, 147 ADD Consortium, 56, 97 Addink, M.J. See Smeenk, C., 388 Addison, R.F., 376, 379 Addison, S. See Collins, M.A., 348; Jones, E.G., 350 Adorf, L. See Coleman, C.O., 101 Agatep, C.D., 73, 97 Agawin, N.S.R. See Vermaat, J.E., 236 Agnew, D.J., 131, 132, 147 Aguilar, A., 376, 379 Aguilar, N.M. See Graham, J.B., 38 Aharoni, J., 256; 272 Ahlgren, M.O. See Bowen, S.H., 304 Ahn, I.-Y., 70, 97 Aidar, E. See Lourenco, S.O., 306 Airoldi, L., 161–236, 164, 166, 167, 168, 170, 171, 173, 176, 177, 178, 179, 180, 184, 186, 187, 188, 189, 190, 191, 193, 194, 195, 196, 207, 211, 212, 216, 217, 218, 219, 220, 221, 222, 223, 224, 225, 226; See Benedetti-Cecchi, L., 226 Ajani, P. See Roberts, D.E., 234 Akazawa, Y. See Wada, H., 353 Alarcon-Castillo, J.C., 73, 97 Alber, M., 281, 288, 303; See D’Avanzo, C.D., 305 Albertelli, G. See Cattaneo-Vietti, R., 99 Albrecht, A.S., 185, 186, 187, 190, 196, 203, 226 Aleem, A.A., 165, 177, 226 Alexander, M. See Boaventura, D., 227 Alexandre, A. See Fontanier, C., 37 Alexandre, B. See Silva, M.A., 387 Alkemade, R., 124, 125, 147 Allchin, C.R. See Jepson, P.D., 383; See Kuiken, T., 384 Allcock, A.L., 68, 97 Alldredge, A., 288, 289, 303 Allen, G.R. See Randall, J.E., 308 Allen, J.A. See Rex, M.A., 111 Allen S.D. See Valiela, I., 158 Allison, P.A., 315, 317, 322, 323, 324, 333, 344, 348

Aloia, A. See Colombini, I., 150 Alongi, D.M., 280, 287, 303; See Hansen, J.A., 305 Alonzo, F., 286, 303 Alve, E., 15, 34; See Bernhard, J.M., 9, 35 Amann, R. See Dubilier, N., 37 Amano, M., 356, 379; See Miyazaki, N., 385 Ambrose Jr, W.G. See Renaud, P.E., 234 Amesbury, S.S., 162, 226 Amundin, M., 370, 379 Andersen, L.W., 356, 379 Andersen, S., 357, 359, 372, 375, 379; See Clausen, B., 368, 380 Anderson, D.M. See MacDonald, L.H., 232 Anderson, G.C., 5, 34 Anderson, G.S., 332, 348 Anderson, K. See Lockyer, C., 385 Anderson, R.Y. See Dean, W.E., 36 Anderson, T.W. See Williams, S.L., 236 Anderson, W.B., 137, 148; See Polis, G.A., 156 Andres, H.G. See Rauschert, M., 62, 110 Andriashek, D., 122, 148 Androsova, E.I., 72, 97 Anonymous, 374, 375, 376, 379, 380 Anschutz, P. See Fontanier, C., 37 Appleby, E.C. See Kuiken, T., 384 Arakawa, H., 196, 201, 206, 220, 226 Arias-Gonzalez, J.E., 280, 303 Arima, S. See Yoshida, G., 236 Arin, T. See White, A.T., 236 Armstrong, J.D. See Priede, I.G., 352 Arnaud, P.M., 65, 67, 73, 87, 93, 94, 97; See Cantera, J.R., 67, 99; See Warén, A., 113 Arnold, A.J. See Nienstedt, J.C., 14, 25, 41 Arnold, P.W. See Gaskin, D.E., 382 Arnould, J.P.Y. See Walker, T.R., 159 Arntz, W. See Gallardo, V.A., 38; See Tarazona, J., 44 Arntz, W.E., 8, 27, 28, 56, 78, 86, 87, 34, 35, 97; See Brey, T., 99; See Gutt, J., 104; See Klages, M., 350; See Neira, C., 41; See Rosenberg, R., 42; See Tarazona, J., 44 Arroyo, N.L. See Neira, C., 41 Arthur, M.A. See Glenn, C.R., 38 Arthur, W. See Hodge, S., 127, 128, 131, 134, 135, 153; See Hyslop, B.T., 230; See Leggett, M.C., 154; See Phillips, D.S., 127, 131, 134, 156 ASCOBANS, 357, 380

391

AUTHOR INDEX

Aseltine, D.A. See Palmer-Zwahlen, M.L., 187, 233 Ashjian, J. See Wishner, K.F, 44 Ashley, C.W., 333, 348 Astier, J.M. See Meinesz, A., 232 Au, W.W.L. See Kastelein, R.A., 383 Averintsev, V.G., 70, 97 Avery, D.M. See Avery, G., 148 Avery, G., 122, 148 Axe, L.M. See Choat, J.H., 298, 304 Ayling, A.M., 178, 187, 226 Azam, F., 288, 303; See Simon, M., 286, 308 Baas, J.H., 15, 35 Babin, R. See Fretey, J., 129, 151 Bach, H.K., 166, 167, 221, 226 Bachelet, G. See Goulletquer, P., 274 Bacher, C. See Gangnery, A., 274 Baco, A.R., 324, 326, 329, 330, 339, 340, 341, 343, 344, 348; See Distel, D.L., 349; See Feldman, R.A., 349; See Smith, C.R., 311–354, 314, 324, 352; See Williams, A.B., 354 Baco-Taylor, A.R., 317, 319, 324, 325, 326, 327, 328, 329, 330, 331, 339, 340, 342, 344, 348 Bagley, P.M. See Collins, M.A., 348; See Jones, E.G., 350; See Priede, I.G., 352 Bailey, G.W., 28, 35 Bailey, T.G., 290, 303 Baines, M. See Pierpoint, C., 386 Baird, R.W., 360, 380 Baker, E.T. See German, C.R., 103 Baker, J.R. See Jepson, P.D., 383; See Kuiken, T., 384 Bakun, A. See Weeks, S.J., 44 Bakus, G.J., 190, 226 Balduzzi, A. See Sarà, M., 111 Baldwin, C.T. See Rhoads, D.C., 42 Balestri, E. See Piazzi, L., 233 Ballesteros, E., 177, 226 Bally, R., 164, 165, 226; See Tarr, J.G., 158 Baltazar, M. See Gallardo, V.A., 37; See Gutiérrez, D., 38 Bamber, R.N., 65, 97 Bamford, M. See Schulz, M., 135, 157 Banse, K., 27, 35 Baptiste, E. See Hyland, J., 39 Barash, A., 254, 271, 272 Barbarino, E. See Lourenco, S.O., 306 Barber, R. See Smith, S.L., 44 Barberio, C., 133, 148 Barilotti, D.C., 142, 148 Barko, J.W., 162, 226 Barnard, J.L., 62, 98; See Coleman, C.O., 101 Barnes, D.K.A., 72, 89, 91, 98 Barnes, L.G. See Goedert, J.L., 349; See Squires,

R.L., 353 Barnes, R.S.K., 50, 51, 58, 61, 62, 63, 65, 69, 72, 74, 75, 79, 97 Barnett, A.M. See Raimondi, P.T., 234 Barrett, J.C. See Iannuzzi, T.J., 230 Barrette, C. See Fontaine, P.M., 381 Barry, J.P. See McHatton, S.C., 41; See Tamburri, M.N., 336, 353 Barsotti, G., 243, 272 Bartbieri, M. See Sarà, M., 111 Barthel, D., 59, 93, 98 Barwick, K.L. See Zmarzly, D.L., 354 Bassindale, R. See Lilly, S.J., 231 Battershill, C.N., 59, 98 Bauchot, M.-L. See Whitehead, P.J.P., 389 Bavay, A. See Tillier, L., 239, 255, 256, 277 Bavestrello, G., 165, 167, 173, 226; See Sarà, M., 111 Baxter, C.H. See Adams, P.B., 34 Bayer, F., 60, 98 Bayer, F.M., 60, 98 Baynes, T.W., 177, 226 Bazzocchi, P., 258, 272 Beaudreau, A., 24, 35 Beaulieu, S.E., 6, 35, 340, 348 Bedford, A.P., 126, 131, 148 Bedulli, D. See Sabelli, B., 277 Beer, N. See Weslawski, J.M., 159 Begeman, J. See Bruggemann, J.H., 304 Begon, M., 280, 281, 303 Behbehani, M.I., 119, 126, 127, 131, 148 Behl, R.J., 32, 35; See Cannariato, K.G., 36 B[ellet], D., 260, 272 Bellot, A. See Gabriele, M., 229 Belluscí de Miralles, D.A., 60, 98 Bellwood, D.R., 290, 296, 297, 298, 303, 304; See Ackerman, J., 302, 303; See Choat, J.H., 295, 297, 304; See Depczynski, M., 292, 294, 302, 305; See Purcell, S.W., 282, 283, 284, 295, 296, 301, 307; See Wilson, S.K., 279–309, 280, 281, 282, 284, 288, 292, 295, 296, 309 Benedetti-Cecchi, L., 193, 226 Benham, W.B., 70, 98 Benke, H., 361, 363, 365, 370, 380; See HeideJørgensen, M.-P., 382; See Hammond, P.S., 382; See Kock, K.H., 368, 384; See Siebert, U., 387 Bennett, B.A., 314, 315, 317, 323, 324, 325, 326, 331, 333, 339, 342, 348; See Allison, P.A., 348 Bennett, P.M. See Jepson, P.D., 383; See Kuiken, T., 384 Berelson, W. See Stott, L.D., 44 Berg, B.S., 140, 141, 148 Berger, W.H. See Herguera, J.C., 15, 38 Berggren, P., 361, 363, 370, 380; See Börjesson,

392

AUTHOR INDEX

P., 356, 360, 361, 364, 365, 366, 380; See Hammond, P.S., 382; See Wang, J.Y., 356, 389; See Westgate, A.J., 389 Berghahn, R., 374, 380 Bergman, K. See McClanahan, T.R., 306 Bergquist, P.R., 59, 98 Berlow, E.L. See Menge, B.A., 232 Bernasconi, I., 73, 98 Bernhard, J., 2, 8, 9, 12, 14, 15, 24, 35; See Reimers, C., 42 Bernhard, J.M., 7, 8, 9, 12, 14, 15, 20, 25, 26, 30, 35; See Alve, E., 15, 34; See Gooday, A.J., 38; See Müller, M.C., 41; See Pike, J., 42; See Todaro, M.A., 44 Bernier, P., 166, 227 Berrow, S.D., 356, 380; See Rogan, E., 360, 361, 363, 368, 370, 387; See Tregenza, N.J.C., 389 Berry, J., 373, 380 Bertin, G. See Chelazzi, L., 149 Bertness, M.D., 140, 148, 169, 172, 177, 227; See Brewer, J.S., 148 Bervas, J.Y. See Desbruyeres, D., 349 Best, P.B. See Sekiguchi, K., 387 Beyers, C.J.De.B. See Bailey, G.W., 35 Bianchi, C.N. See Morri, C., 275 Bianchi, T.S., 6, 35 Biddanda, B.A., 289, 304 Bigot, L., 126, 148 Bini, G., 243, 272 Birje, J., 165, 177, 194, 227 Bjørge, A., 357, 380; See Aarefjord, H., 379; See Donovan, G.P., 356, 357, 362, 381; See Tolley, K.A., 388 Bjorke, A. See Kleivane, L., 384 Bjorndal, K.A. See Thayer, G.W., 158 Black, M.B., 333, 348; See Feldman, R.A., 349 Blackwood, J. See Monaghan, P., 385 Blair, B.A. See Gaskin, D.E., 368, 382 Blake, G.A. See Pringle, C.M., 169, 233 Blanche, K.R., 141, 148 Blanchette, C.A. See Menge, B.A., 232 Blanco, O.M., 60, 98 Blazewicz, M., 64, 98 Bloesch, J., 173, 227 Blokker, P. See Kuypers, M.M.M., 40 Blomqvist, S., 173, 227 Bluteau, R. See Lambert, L., 275 Boaventura, D., 189, 227 Bockus, D. See Slattery, M., 165, 173, 189, 202, 206, 235 Bodkin, J.L., 129, 148 Boer, H. See Kastelein, R.A., 383 Bogi, C., 258, 273 Bolton, J.J. See Engledow, H.R., 165, 167, 172, 192, 194, 195, 229

Bombace, G., 239, 273 Boness, D.J. See Iverson, S.J., 383 Boney, A.D., 166, 187, 198, 227 Bonsdorf, E. See Karlson, K., 39 Boom, R.C.E. See Moodley, L., 41 Boon, J.P., 377, 380; See Duinker, J.C., 381 Borchers, D.L. See Hammond, P.S., 382 Borges, L. See Evans, P.G.H., 362, 381 Börjesson, P., 356, 360, 361, 364, 365, 366, 380 Borne, P.F. See Bernhard, J.M., 35 Borradaile, L.A., 65, 98 Borrell, A. See Aguilar, A., 376, 379 Bosma, E.M. See Bruggemann, J.H., 304 Bosman, A.L. See Branch, G.M., 227 Böttcher, M.E. See Bernier, P., 227 Bottjer, D.J. See Savrda, C.E., 28, 32, 43 Botton, M.L., 123, 148 Boucher, G., 50, 98 Bouchet, P., 84, 98, 243, 273; See Warén, A., 338, 340, 353 Boudouresque, C.F. See Francour, P., 274 Boudreau, B.P., 29, 35 Boudry, P. See Fabioux, C., 274 Boulle, D.P. See Jennings, S., 306 Bouslama, M.F. See Colombini, I., 150 Bowen, S.H., 280, 281, 283, 284, 285, 294, 295, 296, 304 Bowen, W.D. See Iverson, S.J., 383; See Kirsch, P.E., 384 Bowes, A. See Field, J.G., 151 Bowman, D., 145, 148 Bowser, S.S. See Bernhard, J.M., 9, 35; See Travis, J.L., 9, 44 Boyle, P.R. See Pierce, G.J., 358, 367, 377, 386 Bradley, D.W. See Bradley, R.A., 136, 148 Bradley, R.A., 136, 148 Bradshaw, J.S., 8, 35 Braﬁeld, A.E., 121, 148 Braham, H., 336, 348 Braham, H.W. See Rice, D.W., 352 Braine S. See Avery, G., 148 Branch, G.M., 165, 169, 172, 177, 188, 190, 227; See Bustamante, R.H., 117, 137, 149; See Pulfrich, A., 234 Brandon, M.A. See Trathan, P.N., 112 Brandt, A., 64, 84, 98, 99 Brattström, H., 177, 227 Breber, P., 261, 273 Breda, V.A. See Williams, S.L., 236 Breeman, A.M. See Bruggemann, J.H., 304 Breiwick, J. See Gosho, M., 349 Breiwick, J.M. See Mizroch, S.A., 351 Bressan, G. See Falace, A., 178, 194, 229 Brett, K. See Carefoot, T.H., 149 Brewer, J.S., 140, 148 Brewer, P.G. See Hinrichs, K.U., 38

393

AUTHOR INDEX

Brey, T., 85, 90, 93, 99; See Arntz, W.E., 97 Bridgeford, P., 122, 148 Bridges, T.S. See Levin, L.A., 253, 275 Briggs, D.E.G., 342, 348 Brinkhoff, T. See Schulz, H.N., 43 Britton, J.C., 122, 148, 149, 332, 348 Broch, H., 60, 99 Broch, H.J., 60, 99 Brodie, P.F., 371, 380 Broenkow, W.W. See Martin, J.H., 40 Bronsted, H.V., 58, 99 Brooks, J.M. See Fisher, C.R., 349 Brouwer, A., 376, 380 Brown, A.C., 116, 118, 121, 124, 125, 133, 138, 146, 149, 221, 227; See Bally, R., 226 Browne, E.T., 60, 99 Brownell, R.L. See Hohn, A.A., 356, 383 Bruggemann, J.H., 297, 304 Bruguier, N.I. See German, C.R., 103 Brunetti, R. See Gabriele, M., 229 Bruns, T. See Klages, M., 106 Bruun, A.F., 312, 315, 348 Bryant, P.J. See Perkins, J.S., 386 Buck, K.R. See Bernhard, J.M., 35 Buckland, S.T. See Hammond, P.S., 382 Buestel, D. See Gangnery, A., 274 Bugoni, L., 144, 149 Bulleri, F. See Benedetti-Cecchi, L., 226 Bullivant, J.S., 93, 99 Burd, A.C., 375, 380; See Cushing, D.H., 375, 381 Burger, A.E., 121, 149 Burger, J. See Clark, K.E., 150; See Tsipoura, N., 123, 158 Burger, J.F. See Lord, W.D., 121, 130, 154 Burlado, B. See Sarà, M., 111 Burnett, B.R. See Hessler, R.R., 350 Burnett, W. See Kim, K.H., 29, 39 Burnett, W.C. See Glenn, C.R., 38 Burns, K. See Wilson, S.K., 309 Burns, M.D. See Monaghan, P., 385 Burns, N.M. See Bloesch, J., 173, 227 Burrows, E.M., 189, 197, 227 Burt, J.R. See Murray, J., 376, 385 Burt, W.V. See Capuzzo, J.M., 227 Burton, M., 58, 59, 99 Bustamante, R.H., 117, 137, 139; See Pulfrich, A., 234 Bustin R.M. See Calvert, S.E., 35 Butler, I.B. See Pike, J., 42 Butler, J.L. See Adams, P.B., 34; See Hunter, J.R., 39 Butler, M.J. See Herrnkind, W.F., 230 Butlin, R.K., 134, 149; See Day, T.H., 150 Butman, C.A., 173, 227, 314, 315, 333, 336, 338, 345, 348

Buzzurro, G., 271, 273 Byrd, J.H., 332, 348 Cadien, D.B. See Montagne, D.E., 28, 41 Cahoon, L.B., 285, 304 Cailliet, G.M. See Josselyn, M.N., 153 Cairns, S.D., 60, 99 Cal Rodriguez, R.M. See Newell, S.Y., 155 Calcinai, B. See Pansini, M., 109 Callan, E. McC., 120, 149 Calvert, S.E., 6, 29, 35; See Cowie, G.L., 36 Cammidge, L. See Levin, L.A., 40 Campani, E. See Barsotti, G., 243, 272 Campbell, J. See Hyland, J., 39 Campbell, W.J. See Gloersen, P., 103 Camphuysen, C.J., 364, 375, 380 Cancela da Fonseca, L. See Boaventura, D., 227 Cañete, J.I. See Gallardo, V.A., 37 Canﬁeld, D.E. See Fossing, H., 37 Cannariato, K.G., 5, 32, 36 Cantera, J.R., 67, 99; See Warén, A., 113 Cantone, G., 70, 99 Capuli, E. See Pauly, D., 386 Capuzzo, J.M., 163, 167, 227 Carbajal, G. See Rosenberg, R., 42 Carballo, J.L., 177, 227; See Naranjo, S.A., 233 Carbonel, P. See Fontanier, C., 37 Carefoot, T.H., 132, 149; See Pennings, S.C., 156 Carleton, J.H. See Hamner, W.M., 305 Carlsson, B. See Olin, R., 155 Carlton, J.T., 238, 251, 261, 267, 271, 272, 273; See Butman, C.A., 348 Carpenter, R.C., 187, 227 Carpenter, S.R., 225, 227; See Barko, J.W., 226 Carrasco, F. See Gutiérrez, D., 38 Carrasco, F.D. See Gallardo, V.A., 37 Carreon, M. See Moore, S.L., 155 Carter, R.A. See Field, J.G., 151; See Jarman, N.G., 117, 153 Cary, S. See Hentschel, U., 38 Cary, S.C., 30, 31, 36 Castelli, A., 70, 99; See Gambi, M.C., 103 Castilla, J.C., 166, 193, 194, 227; See Fariña, J.M., 166, 172, 194, 229 Castner, J.L. See Byrd, J.H., 332, 348 Castric-Fey, A., 177, 227 Castro, G., 123, 149 Caswell, H. See Etter, R.J., 315, 349 Cattaneo-Vietti, R., 67, 260, 99, 273; See Bavestrello, G., 226; See Pansini, M., 109 Caussanel, C., 120, 149 Cavalieri, D.J. See Gloersen, P., 103 Cavanaugh, C. See Distel, D.L., 349 Cecalupo, A., 273 Cedhagen, T., 36 Cerrano, C. See Bavestrello, G., 226

394

AUTHOR INDEX

Cesari, P., 261, 273 Chalmers, A.G. See Fischer, J.M., 151 Chapelle, G., 14, 94, 36, 99 Chapman, A.S., 170, 204, 206, 227 Chapman, A.R.O., 117, 149 Chapman, V.J., 177, 227 Chardy, P. See Clavier, J., 304 Charpy, L., 289, 304 Chelazzi, G., 128, 149; See Messana, G., 155 Chelazzi, L., 119, 149; See Barberio, C., 148; See Colombini, I., 115–159, 120, 146, 150, See Messana, G., 155; See Ricci, S., 157 Chen, C. See Corliss, B.H., 15, 36 Cherbonnier, G., 73, 100 Cheung, S.G., 122, 149 Chevillon, C. See Clavier, J., 304 Chevin, H., 126, 128, 138, 149 Chiantore, M. See Cattaneo-Vietti, R., 99 Child, C.A., 65, 100 Childers, S.E. See Levin, L.A., 40 Childress, J.J., 9, 10, 11, 12, 36; See Fisher, C.R., 349; See Sanders, N.K., 11, 43 Chin, C.S., 71, 100; See Klinkhammer, G.P., 106 Chin, W., 288, 304 Chisholm, S.W. See Carpenter, S.R., 227 Choat, J.H., 280, 284, 290, 292, 293, 294, 295, 297, 298, 304; See Bellwood, D.R., 298, 304; See Crossman, D.J., 304; See Wilson, S.R., 279–309 Choi, T.S. See Kim, K.Y., 231 Chou, L.M., 162, 228; See Lim, G.S.Y., 191, 231 Chown, S.L., 128, 131, 133, 149; See Klok, C.J., 133, 153 Christensen, V. See Pauly, D., 386 Christiansen, J.P. See Codispoti, L.A., 3, 36 Chuang, E. See Distel, D.L., 349 Chuman de Flores, E. See Rosenberg, R., 42 Cianfanelli, A. See Chelazzi, L., 149 Cinelli, F. See Airoldi, L., 167, 173, 178, 180, 184, 186, 196, 207, 211, 216, 217, 218, 219, 220, 222, 224, 225, 226; See Benedetti-Cecchi, L., 226; See Piazzi, L., 188, 233 Cita, M.B. See Hsü, K.J., 275 Clanzig, S., 239, 261, 273 Claridge, M.F., 49, 100 Clark, A.H., 73, 100 Clark, A.M., 73, 100 Clark, I. See Pierce, G.J., 386 Clark, K.E., 123, 150 Clarke, A., 47–114, 53, 80, 81, 86, 88, 90, 91, 92, 94, 100, 101, 298, 304; See Brey, T., 93, 99; See Crame, J.A., 90, 101; See Eastman, J.T., 76, 77, 91, 102 Clarke, B.C. See Day, T.H., 150 Clarke, D.J. See McLachlan, A., 155 Clarke, K.R., 49, 101; See Warwick, R.M., 49,

113 Clarke, M.R. See Santos, M.B., 387 Clarke, W.D., 177, 228 Clausen, B., 368, 380; See Andersen, S., 357, 375, 379 Clavier, J., 289, 304 Clayton, M.N., 58, 101, 283, 304; See Wiencke, C., 58, 113 Clements, K.D. See Choat, J.H., 284, 294, 297, 304; See Crossman, D.J., 304 Codd, G.A. See Kuiken, T., 384 Codi, S. See Wilson, S.K., 309 Codispoti, L. See Smith, S.L., 44 Codispoti, L.A., 3, 36; See Morrison, J.M., 41; See Piper, D.Z., 6, 42 Coleman, C.O., 62, 101 Coles, S.L. See Johannes, R.E., 306 Collet, A. See Hammond, P.S., 382 Collins, M.A., 336, 348; See Jones, E.G., 350 Collins, P.M. See Butlin, R.K., 149 Colombini, I., 115–159, 119, 120, 126, 127, 128, 131, 138, 146, 150; See Barberio, C., 148; See Chelazzi, L., 149; See Ricci, S., 157 Comiso, J.C. See Gloersen, P., 103 Committee on Biological Diversity in Marine Systems, 314, 348 Connell, J.H., 127, 129, 150, 331, 332, 349 Connell, S.D. See Irving, A.D., 186, 189, 207, 214, 215, 216, 217, 221, 222, 230; See Melville, A.J., 169, 178, 232 Connelly, D.P. See German, C.R., 103 Connor, J. See Josselyn, M.N., 153 Cook, A.A., 16, 21, 36 Coon, D.A. See Neushul, M., 233 Corbet, G.B., 372, 380 Corliss, B.H., 15, 36 Corpe, H.M. See Thompson, P.M., 388 Corten, A., 375, 380 Cortés, J., 162, 166, 228 Costa, A.S. See Vetter, R.D., 44 Costello, M.J., 57, 101 Cotton, A.D., 165, 176, 181, 197, 228 Couperus, A.S., 374, 381 Cowen, R. See Josselyn, M.N., 153 Cowie, G.L., 6, 36; See Keil, R.G., 39 Craddock, C., 344, 349 Craddock, J.E. See Gannon, D.P., 381 Crafford, J.E., 133, 150 Craig, P.C., 131, 150 Craigie, J.S. See Chapman, A.R.O., 117, 149 Craik, G.J.S., 173, 228 Crame, J.A., 68, 84, 90, 91, 101; See Clarke, A., 86, 90, 91, 92, 100; See Feldmann, R.M., 65, 102 Cranford, P.J. See Hargrave, B.T., 350 Creasey, S., 18, 26, 32, 36

395

AUTHOR INDEX

Creasey, S.S., 26, 36 Cremer, M. See Smith, C.R., 43 Croker, R.A. See Behbehami, M.I., 119, 126, 127, 131, 148 Cronk, Q.C.B., 92, 101 Crooks, J. See Levin, L.A., 40 Crooks, J.A., 267, 273 Crossman, D.J., 281, 282, 283, 284, 286, 296, 297, 304 Crothers, J.H., 194, 228 Crowther, P.R. See Briggs, D.E.G., 342, 348 Croxall, J.P. See Walker, T.R., 159 Cruickshank, A.R.I. See Martill, D.M., 351 Cullen, M. See Frost, A., 145, 151 Cullen, S.J., 128, 134, 150 Cummings, V.J. See Turner, S.J., 235 Cunningham, A.P. See German, C.R., 103 Currie, B. See Weeks, S.J., 44 Currie, B.R., 288, 304 Currie, R.I. See Hart, T.J., 17, 38 Curry, W.B. See Haake, B., 38 Cushing, D.H., 375, 381 Cutler, E.B. See Saiz-Salinas, J.I., 111 Daan, N., 364, 373, 381 Daehlmann, A. See Klinkhammer, G.P., 106 D’Agostino, M.M. See Bernasconi, I., 73, 98 Dahl, A.L., 181, 189, 197, 228 Dahlin, K.A. See Adams, P.B., 34 Dahm, C., See Brey, T., 99 Dahms, H.U. See Herman, R.L., 70, 105 D’Alaffoley, D. See Davidson, N.S., 150 Daly, M.A., 165, 167, 168, 170, 172, 173, 176, 177, 182, 186, 194, 196, 221, 228 Damste, J.S.S. See Kuypers, M.M.M., 40 Danin, Z. See Barash, A., 254, 271, 272 Danko, J.P. See Pennings, S.C., 156 Danovaro, R., 26, 36; See Bavestrello, G., 226 D’Antonio, C.M., 165, 167, 170, 173, 177, 180, 183, 191, 192, 199, 206, 221, 222, 228 Darling, F.F., 359, 381 Darlington, P.J., 88, 101 Datta, S. See Sturtivant, C.R., 388 D’Avanzo, C.D., 281, 305 Davidson, N.S., 143, 150 Davies, M.S. See Hyslop, B.T., 194, 203, 206, 230 Davis, A.R. See Roberts, D.E., 234 Davis, C.L. See Fielding, P.J., 117, 151 Dawah, H.A. See Claridge, M.F., 100 Dawe, C. See Day, T.H., 150 Dawson, E.W., 67, 101 Dawson, P.K. See Taylor, B.L., 364, 388 Day, T.H., 134, 150; See Butlin, R.K., 134, 149; See Cullen, S.J., 150; See Leggett, M.C., 154; See Phillips, D.S., 156 Dayton, P.K., 80, 101, 117, 150, 190, 196, 228,

317, 330, 332, 337, 349; See Hansen, J.A., 305; See Robilliard, G.A., 61, 111; See Seymour, R.J., 234 Dean, H.K. See Saiz-Salinas, J.I., 111 Dean, T.A. See Deysher, L.E., 173, 190, 228 Dean, W.E., 3, 5, 28, 36; See Glenn, C.R., 38 Dearborn, J.H. See Bullivant, J.S., 93, 99; See Snelgrove, P.V.R., 50, 53, 112, 315, 338, 353; See Speel, J.A., 73, 112 de Beer, D. See Dubilier, N., 37 De Broyer, C., 62, 86, 87, 89, 94, 101, 102; See Gutt, J., 104; See Jazdzewski, K., 62, 106 De Buffrénil, V. See Robineau, D., 319, 322, 352 Decraemer, W. See Neira, C., 41 De Grave, S. See Barnes, D.K.A., 72, 89, 91, 98 Dekker, H., 269, 271, 273 de Koeyer, P. See Hemminga, M.A., 152 Del Campo, J.A. See Vargas, M.A., 308 DeLaca, T.E. See Stockton, W.L., 312, 315, 353; See Lipps, J.H., 107; See Moe, R.L., 58, 87, 88, 107 Delamotte, M., 271, 273 de Langlais-Jeannin, I. See Jolivet, E., 153 Delesalle, B. See Arias-Gonzalez, J.E., 303 Dell, R.K., 56, 60, 67, 68, 89, 102, 312, 313, 314, 325, 337, 338, 341, 347, 349 Dell’Anno, A. See Danovaro, R., 36 Della Santina, P. See Boaventura, D., 227 DeLong, E.F. See Hinrichs, K.U., 38 DeLong, R.L. See Gearin, P.J., 382 Demaison, G.J., 6, 36 DeMaster, D.J. See Smith, C.R., 43 DeMaster, D.P. See Perry, S.L., 352 De Min, R., 239, 273 Deming, J., 314, 322, 323, 324, 326, 327, 328, 331, 344, 349 Deming, J.W. See Allison, P.A., 348; See Smith, C.R., 352 Demopoulos, A.W.J. See Smith, C.R., 312, 352 den Dulk, M., 25, 36; See van der Zwaan, G.I., 44 den Hartog, C., 118, 150; See van Katwijk, M.M., 236 DeNiro, M.J., 326, 328, 349 Dennis, J.V., 121, 150; See Gunn, C.R., 120, 152 Depczynski, M., 292, 294, 302, 305 De Pina, G.M.A., 62, 102 De Robertis, A., 11, 36 de Rougemont, G., 126, 150 De Roy Moore, T., 139, 150 de Ruiter, E. See Kleivane, L., 384 Desbruyeres, D., 338, 349 Desportes, G., 360, 361, 381; See Lockyer, C., 385 de Stigter, H.C. See Jorissen, F.J., 39 de Swart, R.L. See Ross, P.S., 387 Dethier, M.N., 178, 228

396

AUTHOR INDEX

Devinny, J.S., 170, 194, 196, 198, 205, 206, 207, 220, 228 Devol, A.H., 6, 36 Dewhurst, H.W., 359, 381 Deysher, L.E., 173, 190, 228 Dhonncha, E.N. See Guiry, M.D., 163, 230 Diack, J.S.W. See Pierce, G.J., 386; See Thompson, P.M., 388 Dial, K.P., 92, 102 Diaz, R.J., 2, 3, 12, 13, 14, 23, 36 Di Cello, F. See Barberio, C., 148 Dieckmann, G.S. See Field, J.G., 151 Dietrich, W.E. See MacDonald, L.H., 232 Dietz, R. See Teilmann, J., 362, 388 Díez, I. See Gorostiaga, J.M., 166, 167, 193, 229, 230 Di Geronimo, I., 256, 273 di Monterosato, T.A., 254, 256, 272, 273 Di Natale, A., 256, 273 Di Pietro, N. See Cantone, G., 70, 99 Distel, D.L., 325, 344, 347, 349 Ditazato, H. See Wada, H., 353 Ditmar, W., 7, 37 Dixon, J.D. See Schroeter, S.C., 234 Dixon, T.J. See Dixon, T.R., 144, 150 Dixon, T.R., 144, 150 Dizon, A.E. See Rosel, P.E., 387 Dobson, T., 127, 128, 131, 134, 135, 150; See Day, T.H., 150 Dodge, R.E., 162, 228 Dols, J. See Boon, J.P., 380 Dommisse, M., 288, 289, 305 Doneddu, M. See Giuseppetti, G., 274 Donn, T.E., 140, 151 Donovan, G.P., 356, 357, 362, 381; See Bjørge, A., 357, 380 Doody, J.P. See Davidson, N.S., 150 Doty, M.S., 176, 228 Douglas, R. See Stott, L.D., 44 Douglas, R.G., 24, 37 Dower, K.M. See McQuaid, C.D., 168, 195, 224, 232 Drake, C.M. See Davidson, N.S., 150 Drews, M. See Schmaljohann, R., 43 Duarte, C.M. See Gacia, E., 169, 229; See Vermaat, J.E., 236 Dubilier, N., 31, 37 Ducklow, H.W., 284, 287, 288, 289, 305 Dudley, R. See Graham, J.B., 38 Dudok van Heel, W.H., 376, 381 Due, A.D., 137, 151 Duedall, I.W. See Capuzzo, J.M., 227 Duff, K.L. See Davidson, N.S., 150 Duffy, J.E. See Paul, V.J., 156 Dugan, J.E., 126, 127, 136, 143, 151 Duggan, R.M. See Zmarzly, D.L., 354

Duggins, D.O., 117, 124, 137, 151, 207, 209, 216, 228; See Eckman, J.E., 210, 216, 228 Duijnstee, I.A.P. See van der Zwaan, G.I., 44 Duinker, J.C., 377, 381 Dunn, D.F., 60, 79, 102 Dunnet, G.M., 362, 381 Dye, A.H. See McLachlan, A., 155 Eagle, R.A., 166, 167, 228 Earl, S. See Pierpoint, C., 386 Earl, S.J. See Pierpoint, C., 386 Earll, R.C., 145, 151 Eastman, J.T., 76, 77, 91, 102 Ebeling, A.W., 191, 228; See Reed, D.C., 234 Ebling, F.J. See Lilly, S.J., 131; See Muntz, L., 233; See Round, F.E., 234 Eckman, J.A. See Jumars, P.A., 26, 39 Eckman, J.E., 169, 173, 210, 216, 228; See Duggins, D.O., 228 Edesa, S. See Levin, L.A., 11, 40 Edwards, W. See Santos, M.B., 387 Eekhout, S. See Branch, G.M., 227; See Bustamante, R.H., 149 Egglishaw, H.J., 134, 135, 151 Egorova, E.N., 67, 102 Eiane, M.K. See De Robertis, A., 36 Eikmeier, N. See Coleman, C.O., 101 Ekau, W., 93, 102; See Gutt, J., 104 Ekdale A.A., 32, 37 Ekman, S., 263, 273 El Beshbeeshy, M., 60, 102 El Sayed, S., 257, 274 Eldredge, L.G., 77, 83, 84, 102 Elfwing, T. See McClanahan, T.R., 307 ElGtari, M. See Colombini, I., 150 Eliot, I.G. See McLachlan, A., 155 Ellis, D.V., 166, 193, 228 Ellison, A.M. See Bertness, M.D., 140, 148 Ellison, J.C., 162, 228 Elmgren, R. See Bianchi, T.S., 35 Elton, C.S., 238, 274 Emblow, C.S. See Costello, M.J., 101 Emeis, K.-C., 28, 37 Emerson, S.E., 172, 177, 187, 191, 229 Engel, A., 289, 305 Engledow, H.R., 165, 167, 172, 192, 194, 195, 229 Enríquez, S. See Vermaat, J.E., 236 Enriquez-Briones, S. See Gallardo, V.A., 37 Ensor, K. See Furness, R.W., 381 Epstein, S. See DeNiro, M.J., 326, 328, 349 Erasmus, T. See McLachlan, A., 155 Erbacher, J. See Kuypers, M.M.M., 40 Ercolini, A. See Messana, G., 155 Erftemeijer, P.L.A. See Ochieng, C.A., 118, 119, 125, 126, 155

397

AUTHOR INDEX

Eriksson, B.K., 176, 193, 229; See Johansson, G., 230 Ernst, S.R. See van der Zwaan, G.I., 44 Erseus, C. See Dubilier, N., 37 Escofet, A. See Lopez-Uriate, E., 154; See Pineda, J., 165, 180, 182, 200, 233 Espinoza, J., 183, 189, 229 Estes, J.A., 169, 229; See Duggins, D.O., 151 Etter, R.J., 315, 349; See Levin, L.A., 40 Evans, C.W. See Hunter, C.L., 162, 230 Evans, P.G.H., 356, 362, 375, 381 Evans, P.R. See Percival, S.M., 136, 156 Evans, R.A., 165, 229 Evans, S.M., 165, 177, 194, 229; See Uneputty, P.A., 145, 158 Evans, W.E., 362, 370, 381 Fabiano, M. See Airoldi, L., 226; See Bavestrello, G., 226; See Danovaro, R., 36 Fabioux, C., 260, 274 Fabricius, K.E., 289, 305 Faideau, F. See Lambert, L., 275 Falace, A., 178, 194, 229 Fallaci, M. See Barberio, C., 148; See Chelazzi, L., 149; See Colombini, I., 150; See Ricci, S., 157 Fallon, R.D. See Harvey, H.R., 38; See Newell, S.Y., 155 Fani, R. See Barberio, C., 148 Fariduddin, M. See Loubere, P., 15, 40 Fariña, J.M., 166, 172, 194, 229 Farmer, M.A. See Bernhard, J.M., 35 Farrant, P.A. See King, R.J., 184, 231 Farrow, G.E., 177, 229 Fauchald, K., 70, 102 Felbeck, H. See Cary, S.C., 36; See Hentschel, U., 38 Feldman, R.A., 324, 349 Feldmann, R.M., 65, 91, 102 Fenchel, T., 6, 37; See Azam, F., 303 Fenical, W. See Hay, M.E., 132, 152; See Paul, V.J., 156 Féral, J.-P. See Poulin, E., 84, 92, 110 Ferdelman, T. See Dubilier, N., 37 Ferdelman, T.G. See Schulz, H.N., 43 Ferrara, F. See Chelazzi, G., 128, 149; See Messana, G., 155 Ferrero, T.J. See Lambshead, P.J.D., 106 Field, J.G., 117, 151; See Azam, F., 303; See Koop, K., 118, 119, 131, 132; See Newell, R.C., 155; See Stuart, V., 158; See Velimirov, B., 158 Field, M.E. See Storlazzi, C.D., 165, 235 Fielding, P.J., 117, 151 Finger, G. See Rosenberg, R., 42 Fink, B.D., 360, 369, 370, 381

Finlay, B.J. See Fenchel, T., 6, 37 Finogenova, N., 24, 39 Fischer, J.M., 140, 151 Fischer, P., 238, 274 Fisher, C.R., 328, 349 Fisher, H.D. See Sergeant, D.E., 360, 387 Fisher, J.S. See Fonseca, M.S., 141, 151, 169, 229 Fisher, W.K., 73, 103 Fitter, A. See Williamson, M., 267, 277 Flagg, C. See Morrison, J.M., 41; See Smith, S.L., 44 Fletcher, R.L. See Chapman, A.S., 170, 204, 206, 227 Flores, L. See Arntz, W.E., 34 Flores, L.A. See Rosenberg, R., 42 Fogg, G.E., 80, 103 Fonck, E. See Santelices, B., 234 Fonseca M.S., 141, 151, 169, 229 Fontaine, P.M., 361, 367, 370, 381 Fontanier, C., 20, 37 Foreman, R.E. See Smith, B.D., 125, 157 Fornos, J.J., 165, 229 Forster, S. See Fossing, H., 37 Fortes, M.D. See Vermaat, J.E., 236 Forteza, V. See Fornos, J.J., 229 Fortin, M.J. See Meekan M.G., 307 Fossing, H., 7, 37; See Schulz, H., 43 Foster, M.S., 194, 229; See Neushul, M., 233 Foster, M.W., 71, 103 Fowler, C.W., 144, 151 Fowler, M.R. See Tunnicliffe, V., 338, 353 Fralick, R.A. See Mathieson, A.C., 176, 232 France, S.C. See Rosel, P.E., 387 Francis, H.C. See Moore, P.G., 131, 155 Francour, P., 239, 274 Frankenberg, D., 7, 22, 23, 37 Fraser, F.C., 359, 381 French, P.W., 165, 166, 169, 229 Frese, K. See Siebert, U., 387 Fretey, J., 129, 151 Fris, M.B., 283, 305; See Horn, M.H., 306 Frost, A., 145, 151 Frost, K.J. See Iverson, S.J., 383 Fry, B., 118, 151, 326, 349; See Van Dover, C.L., 329, 353 Fry, W.G., 65, 66, 103 Fujioka, K., 314, 349; See Naganuma, R., 351; See Wada, H., 353 Fujise, Y. See Miyazaki, N., 385 Fukuchi, M. See Numanami, H., 108 Fuller, G.B., 377, 381 Furnas, M.J. See Wilson, S.K., 279–309 Furness, R.W., 364, 373, 374, 381; See Hudson, A.V., 374, 383 Furstenburg J.P. See McGwynne, L.E., 154 Fusco, G. See Minelli, A., 107

398

AUTHOR INDEX

Gabriele, M., 177, 190, 229 Gacia, E., 169, 229 Gad, G. See Neira, C., 41 Gaffney, P.M. See Ó Foighil, D., 276 Gage, J. See Creasey, S., 36; See Levin, L.A., 40 Gage, J.D., 7, 37, 315, 349; See Lamont, P.A., 9, 40; See Levin, L.A., 6, 23, 25, 33, 40; See Smallwood, B.J., 43; See Smith, C.R., 43 Gaglini, A., 254, 274 Gaines, S.D. See Robertson, D.R., 291, 295, 308; See Sousa, W.P., 235 Galil, B., 266, 274; See Tom, M., 271, 277 Gallagher, E.D. See Jumars, P.A., 22, 39, 317, 350 Gallardo, V. See Levin, L.A., 40 Gallardo, V.A., 7, 8, 22, 27, 28, 30, 37, 38; See Arntz, W.E., 34, 97; See Fossing, H., 37; See Gutiérrez, D., 38; See Jørgensen, B.B., 7, 39; See Neira, C., 41 Gallotti, D. See Gabriele, M., 229 Galzin, R. See Arias-Gonzalez, J.E., 303; See Polunin, N.V.C., 307 Gambi, M.C., 70, 103 Gangnery, A., 261, 274 Gannon, D.P., 361, 364, 366, 367, 370, 381 Gans, C. See Graham, J.B., 38 Garavelli, C.L., 266, 274 Garbary, D.J. See Kim, K.Y., 231 Garcia Carrascosa, A.M. See Peña Cantero, A.L., 60, 109 García-Alvarez, O., 68, 103 García-Gómez, J.C. See Carballo, J.L., 227; See Naranjo, S.A., 233 Gardner, J.V. See Dean, W.E., 36 Gardner, R.H., 225, 229 Gardner, W.D., 173, 229; See Morrison, J.M., 41 Garland, C.D. See Volkman, J.K., 308 Garza, M. See Vetter, R.D., 44 Gaskin, D.E., 333, 336, 349, 356, 360, 361, 364, 366, 367, 368, 369, 370, 375, 381, 382; See Koopman, H.N., 384; See Read, A.J., 356, 357, 377, 386; See Smith, G.J.D., 360, 364, 366, 367, 370, 388; See Wang, J.Y., 389; See Watson, A.P., 369, 389; See Westgate, A.J., 389; See Yasui, W.C., 372, 390; See Yurick, D.B., 356, 390 Gaston, K.J., 48, 95, 103; See Spicer, J.I., 14, 44 Gaurin, S. See Morrison, J.M., 41 Gazey, N.J. See Hyslop, B.T., 230 Gearin, P.J., 360, 361, 365, 367, 382 Gebelein, C.D. See Neumann, A.C., 233 Gelfman, C. See Wishner, K.F., 44, 45 Geller, J.B. See Carlton, J.T., 238, 273 Gerber, R.P., 289, 302, 305 Gerking, S.D., 280, 290, 295, 305; See

Montgomery, W.L., 307 German, C.R., 71, 103; See Tyler, P.A., 353; See Van Dover, C.L., 353 Gerrior, P. See Waring, G.T., 389 Gershon, Z. See Wassermann, M., 389 Giannuzzi-Savelli, R. See Sabelli, B., 227 Gibbons, M.J., 177, 194, 196, 229 Gibbs, P.E., 312, 313, 338, 341, 349 Gibson, R., 67, 103 Giere, O., 16, 31, 38; See Dubilier, N., 37 Giesen, W.B.J.T. See van Katwijk, M.M., 236 Gilat-Gottlieb, E., 271, 274 Gill, M.E. See Evans, S.M., 229 Gillanders, B.M., 147, 151, 166, 170, 229 Gilson, H.C. See Kitching, J.A., 231 Giuseppetti, G., 266, 274 Glaser, B. See Bennet, B.A., 348 Glenn, C.R., 5, 38; See Resig, J.M., 20, 42 Gloersen, P., 53, 103 Glover, A. See Smith, C.R., 352 Glover, H.E. See Ward, B.B., 44 Glud, R.N. See Fossing, H., 37 Glynn, P.W., 302, 305 Gmeiner, C. See Tiedemann, R., 388 Goedert, J.L., 314, 343, 349; See Squires, R.L., 353 Gofas, S., 237–277, 263, 274; See Bouchet, P., 243, 273; See Rueda, J.L., 266, 277; See Zenetos, A., 277 Golani, D., 239, 266, 274 Golik, A. See Bowman, D., 148 Gollash, S. See Reise, K., 277 Gon, O., 76, 77, 103 Gonzalez, R.R., 12, 13, 38 González, S. See Lopez-Uriate, E., 154 Gooday, A.J., 8, 9, 11, 15, 20, 21, 25, 38; See Levin, L.A., 40 Goodson, A.D., 369, 370, 382 Goodson, D.S. See Sturtivant, C.R., 388 Gordon, J. See Davidson, N.S., 150 Gordon, M.S., 13, 38 Gorny, M. See Brey, T., 99; See Gutt, J., 104 Gorny, M.E. See Gutt, J., 104 Gorostiaga, J.M., 166, 167, 176, 178, 229, 230 Gorsline, D. See Stott, L.D., 44 Gorsline, D.S. See Savrda, C.E., 43 Gosho, M., 336, 349 Gotelli, N.J., 85, 103, 167, 173, 207, 208, 216, 217, 230 Gottfried, M., 289, 305 Goulletquer, P., 267, 274 Gowing, M. See Wishner, K., 45 Gowing, M.M., 20, 31, 38; See Smith, S.L., 44; See Wishner, K.F., 44, 45 Graco, M. See Gallardo, V.A., 38 Graham, J.B., 14, 38; See Yang, T.H., 45

399

AUTHOR INDEX

Grahame, J., 177, 188, 196, 230 Grant, G.S., 131, 151 Grant, W.D. See Butman, C.A., 227 Grassle, J.F., 50, 103, 315, 338, 349; See Snelgrove, P.V.R., 353 Graus, R.R., 94, 104 Graves, G.R. See Gotelli, N.J., 85, 103 Gray, J.S., 2, 12, 13, 14, 33, 38, 90, 95, 104; See Azam, F., 303 Gray, L.E., 377, 382 Green, G., 332, 350 Green, P.T., 121, 151 Greenwood, J.J.D. See Harwood, J., 373, 382 Gregorio, D. See Moore, S.L., 155 Greig, J.A., 73, 104 Greppi, E. See Buzzurro, G., 271, 273 Grieshaber, M.K., 12, 13, 38 Grifﬁths, C.L., 117, 119, 126, 127, 128, 135, 137, 151; See Field, J.G., 151; See Koop, K., 123, 124, 153; See Newell, R.C., 155; See StentonDozey, J.M.E., 118, 119, 123, 126, 128, 131, 158; See Tarr, J.G., 158; See Velimirov, B., 158 Grifﬁths, R.J. See Field, J.G., 151 Grime, J.P., 171, 221, 230 Groene L.C. See Newell, S.Y., 155 Gross, O., 28, 38 Grottanelli, V.L., 142, 143, 151 Grover, P.B., 136, 151 Gruvel, A., 65, 256, 104, 274 Grygier, M., 65, 104 Grygier, M.J., 65, 104 Gualberto, E. See Schramm, W., 308 Guerrero, M.G. See Vargas, M.A., 308 Guidetti, P., 166, 230 Guidi, J.-B. See Bernier, P., 227 Guille, A. See Cherbonnier, G., 73, 100 Guiry, M.D., 163, 230 Guitart, R. See Tomás, J., 158 Guizzardi, M. See Gambi, M.C., 103 Gulliksen, B., 165, 173, 230; See Evans, R.A., 229; See Jørgensen, L.L., 164, 167, 195, 230 Gundersen, J.K. See Fossing, H., 37 Gundersen, J.S. See Morrison, J.M., 41 Gunn, C.R., 120, 152; See Dennis, J.V., 121, 150 Gunnison, D. See Barko, J.W., 226 Gustafson, K. See Paul, V.J., 156 Gustafson, R.G. See Craddock, C., 349; See Peek, A.S., 352 Gutiérrez, D., 5, 16, 22, 27, 28, 30, 38; See Neira, C., 41 Gutschick, R. See Rhoads, D.C., 42 Gutt, J., 59, 67, 73, 87, 93, 104; See Arntz, W.E., 97; See Barthel, D., 93, 98; See Brey, T., 99; See Ekau, W., 93, 102; See Klages, M., 106; See Piepenburg, D., 110

Gwada, P. See Hemminga, M.A., 152 Haake, B., 6, 38 Haas, G., 243, 256, 274 Haase, E. See Tiedemann, R., 388 Haedrich, R.L., 32, 38 Hain, S., 67, 104; See Brey, T., 99 Håkanson, L. See Blomqvist, S. 173, 227 Hall, A.J. See Hammond, P.S., 382 Hamilton, S.C. See Smith, C.R., 18, 43 Hammill, P.O. See Fontaine, P.M., 381 Hammond, A.L., 115, 152 Hammond, P.S., 357, 358, 372, 373, 379, 382; See Tregenza, N.J.C., 389 Hamner, W.M., 288, 289, 305 Hanna, F.S. See Grahame, J., 177, 188, 196, 230 Hansen, H.J. See Leutenegger, S., 9, 40 Hansen, J.A., 119, 152, 280, 286, 305; See Lenanton, R.C.J., 154; See Robertson, A.I., 117, 137, 157 Hanski, I. See Koskela, H., 129, 153 Hardeman, J. See Kastelein, R.A., 383 Harder, J. See Tiedemann, R., 388 Hardewig, I. See Grieshaber, M.K., 38 Hardiman, P.A. See Eagle, R.A., 228 Harding, J.M. See Mann, R., 275 Hardy, A.C., 359, 382 Hardy, F.G. See Evans, S.M., 229 Hardy, T. See Crossman, D.J., 304 Harger, B.W.W. See Neushul, M., 233 Hargrave, B.T., 336, 350 Härkönen, T.J., 372, 373, 382 Harmelin, J.G. See Francour, P., 274 Harmelin-Vivien, M. See Polunin, N.V.C., 307, 296, 305; See Francour, P., 274 Harmer, S.F., 357, 359, 382 Harper, D.E. See Rabalais, N.N., 42 Harper, J.L. See Begon, M., 303 Harries, H.C., 138, 152 Harris, E.A. See Kuiken, T., 384 Harris, L.H. See Levin, L.A., 351 Harris, M.P., 364, 373, 382 Harris, R. See Pierpoint, C., 386 Harris, S. See Corbet, G.B., 372, 380 Harrison, P.G., 125, 126, 132, 152 Harrold, C., 347, 350 Hart, H., 17, 38, 346, 350 Hart, M.B. See Koutsoukos, E.A.M., 40 Hart, T.J., 17, 38 Hartman, O., 70, 104 Hartman, W.D., 59, 104 Hartmann, G.C. See Abell, D.H., 147 Hartmann-Schröder, G., 70, 104, 106 Hartnett, H.E. See Devol, A.H., 6, 36 Hartwig, E.O., 116, 152 Harvey, H.R., 6, 38

400

AUTHOR INDEX

Harwood, J., 373, 382 Hasan, A.K., 260, 274 Hashimoto, J. See Craddock, C., 349 Haslam, S.F.I., 139, 141, 152 Haszprunar, G., 337, 350; See McLean, J.H., 340, 351 Hatakeyama, Y., 370, 382 Hatcher, B.G., 280, 286, 287, 298, 305 Hauri I.R. See Hamner, W.M., 305 Hawes, S. See Josselyn, M.N., 153 Hawkins, J.P. See Roberts, C.M., 272, 277 Hawkins, L.E. See Cook, A.A., 36 Hawkins, S.J. See Boaventura, D., 227 Hawksworth, D.L., 48, 105 Hay, M.E., 132, 152; See Paul, V.J., 156 Hayden, B.P. See Ray, G.C., 115, 156 Hayes, J.M. See Hinrichs, K.U., 38, 39 Hayes, W.B., 119, 131, 132, 133, 152 Haygood, M.G. See Rosel, P.E., 387 Haynes, D., 145, 152 Hayns, S. See Crothers, J.H., 194, 228 Hayward, P.J., 72, 83, 84, 104 Headland, R.K., 80, 105 Heard, S.B., 127, 131, 152 Hecker, B., 344, 350 Hedgpeth, J.W., 89, 105; See Fry, W.G., 65, 66, 103 Heemstra, P.C. See Gon, O., 76, 77, 103 Heide-Jørgensen, M.-P., 357, 382; See Hammond, P.S., 382; See Härkönen, T.J., 373, 382 Heimlich-Boran, S. See Hammond, P.S., 382 Heip, C., 126, 152 Heisterkamp, S. See Swart, R. de L., 388 Held, C., 84, 105 Heldal, H.E. See Tolley, K.A., 356, 388 Helfman, G.S. See Meyer, J.L., 307 Hellemaa, P., 140, 152 Hemminga, M.A., 118, 125, 140, 152 Henley, W.F., 162, 230 Hentschel, U. 30, 38 Herbert T.D. See Sarmiento, J.L., 43 Herguera, J.C., 15, 38 Herman, P.M.J. See Moodley, L., 41 Herman, R.L., 70, 105 Hernandez, M. See Schulz, H.N., 43 Heros, V. See Bouchet, P., 98 Herrnkind, W.F., 173, 187, 200, 230 Hersteinsson, P., 122, 152 Hesp, P.A., 139, 140, 152 Hess, C. See Moodley, L., 15, 41 Hessler, R.R., 317, 330, 350; See Dayton, P.K., 317, 330, 332, 337, 349; See Levin, L.A., 40; See Rex, M.A., 111; See Smith, C.R., 315, 352 Heurtebise, S. See Fabioux, C., 274 Hewitt, J.E. See Turner, S.J., 235 Hewson, R., 332, 350

Heywood, V.H., 48, 105 Hiatt, R.W., 290, 305 Hiby, A.R. See Hammond, P.S., 382 Hickman, C.S. See Lipps, J.H., 91, 107 Hidalgo, J.G., 263, 274 Hilbish, T.J. See Ó Foighil, D., 276 Hillebrand, M.T.J. See Boon, J.P., 380; See Duinker, J.C., 381 Hillier, P.C. See Day, T.H., 150 Hinrichs, K.U., 8, 38, 39 Hiscock, K., 164, 167, 230; See Norton, T.A., 233 Hiscocks, K., 122, 152 Hislop, J.R.G., 362, 371, 383; See Furness, R.W., 364, 373, 381; See Pierce, G.J., 386; See Thompson, P.M., 388 Hobbie, J.E. See Valiela, I., 158 Hobday, A.J., 119, 153 Hobson, K.A., 359, 383 Hobson, W.C. See Fuller, G.B., 377, 381 Hodge, S., 127, 128, 131, 134, 135, 153 Hoeﬂe, U. See Kuiken, T., 384 Hoeh, W.R. See Craddock, C., 349 Hoek, P.P.C., 65, 105 Hoenselaar, H.J., 261, 274 Hoenselaar, J. See Hoenselaar, H.J., 261, 274 Hoffmann-Kobert, B. See Coleman, C.O., 101 Hogler, J.A., 342, 343, 346, 350 Hohn, A.A., 356, 383; See Read, A.J., 356, 386 Holden, A.V., 377, 383 Holmer, M., 166, 230 Holroyd, S. See Hyslop, B.T., 230 Holt, R.D. See Polis, G.A., 156 Hong, Y., 288, 305 Hooker, S.K., 359, 383; See Baird, R.W., 360, 380 Hoover, D. See Smallwood, B.J., 43 Hoover, D.J. See Smith, C.R., 43 Hopkins, D.W. See Haslam, S.F.I., 139, 141, 152 Hopkins, T.L., 70, 105 Hoppe, K., 254, 274 Horn, M.H., 280, 290, 297, 306; See Fris, M.B., 283, 305 Houart, R., 254, 275 Houston, D.C., 332, 350 Howes, G.J. See Kuiken, T., 384 Howson, C.M., 83, 84, 105 Hoyt, W.D., 176, 230 Hsü, K.J., 238, 275 Huang, L.Q. See Ren, X.Q., 65, 110 Hubbard, D.M. See Dugan, J.E., 126, 127, 136, 143, 151 Hubbell, S.P., 48, 85, 105 Hubold, G., 85, 105 Hudson, A.V., 374, 383; See Furness, R.W., 381 Huggett, C.L. See Levin, L.A., 40, 351

401

AUTHOR INDEX

Huh, S.H. See Kim, K.Y., 231 Huitric, M. See McClanahan, T.R., 306 Hummon, W.D. See Todaro, M.A., 44 Hunt, J.N. See Prince, W.A., 156 Hunt, W.G. See Moriarty, D.J.W., 307 Hunter, C.L., 162, 230 Hunter, J.R., 12, 23, 39; See Jacobson, L.D., 23, 24, 39 Hurd, S.D. See Polis, G.A., 121, 136, 137, 138, 156 Hureau, J.-C. See Whitehead, P.J.P., 389 Hurlbert, S.H., 85, 105 Hurley, A.C. See Josselyn, M.N., 153 Huston, M.A., 48, 105 Hüttel, M. See Fossing, H., 37 Huvet, A. See Fabioux, C., 274 Hyland, J., 16, 39 Hyslop, B.T., 166, 167, 172, 193, 194, 203, 206, 230 Iannuzzi, T.J., 166, 172, 230 Igarashi, A. See Numanami, H., 108 Ijima, I., 59, 105 Imhoff, J.F. See Schmaljohann, R., 43 Inglis, G., 116, 123, 125, 126, 127, 128, 153 Ingram, C.L. See Hessler, R.R., 350 Irelan, C. See Horn, M.H., 306 Irimura, S. See Seno, J., 73, 111 Irving, A.D., 186, 189, 207, 214, 215, 216, 217, 221, 222, 230 Isaacs, J.D., 317, 332, 337, 350 Issel, A., 238, 275 Ittekkott, V. See Haake, B., 38 Iverson, S.J., 359, 378, 383; See Hooker, S.K., 383; See Kirsch, P.E., 384; See Koopman, H.N., 384 Iwami, T. See Numanami, H., 108 IWC, 356, 357, 362, 368, 375, 376, 383 Jablonski, D., 253, 275; See Lutz, R., 351; See Roy, K., 111 Jackson, J.B.C., 186, 230 Jackson, N.L. See Thornton, L., 145, 158 Jacobs, D.K., 5, 32, 33, 39 Jacobson, L.D., 23, 24, 39 Jacobson, T.R. See Botton, M.L., 148 Jacques, G. See Tréguer, P., 53, 112 Jahnke, R.A. See Glenn, C.R., 38 James, D.W. See Levin, L.A., 351 Jameson, R.J. See Bodkin, J.L., 129, 148 Jannasch, H.W. See McHatton, S.C., 41 Jannik, N.T. See van der Zwaan, G.I., 44 Janowska, E. See Sicinski, J., 70, 87, 111 Janson, K., 254, 275 Janssen, H.H. See Brandt, A., 64, 99 Jarman, N. See Field, J.G., 151

Jarman, N.G., 117, 153 Jarms, G., 60, 105 Jaume, C. See Fornos, J.J., 229 Jaya Sree, V., 65, 105; See Sreepada, R.A., 112 Jazdzewski, K., 62, 87, 106; See Arnaud, P.M., 97; See Blazewicz, M., 64, 97; See De Broyer, C., 62, 86, 87, 89, 94, 102 Jeffrey, S.W. See Volkman, J.K., 308 Jelmert, A., 336, 338, 350 Jennings, S., 297, 306; See Pinnegar, J.K., 386 Jensen, A. See Lund-Hansen, L.C., 231 Jensen, P. See Lambshead, P.J.D., 106 Jepson, P.D., 377, 383 Jipa, D. See Manheim, F., 40 Johannes, R.E., 162, 230, 288, 289, 302, 306 Johannesson, K., 243, 275 Johansson, B. See Bianchi, T.S., 35 Johansson, G., 177, 230; See Eriksson, B.K., 229 John, D.D., 73, 106 Johns, R.B. See Currie, B.R., 288, 304; See Shaw, P.M., 287, 308 Johnson, J., 336, 350 Johnson, M.A. See Gearin, P.J., 382 Johnson, S. See Vadas, R.L., 236 Johnston, D.W. See Palka, D.L., 385 Johnston, I.A. See Clarke, A., 91, 100 Johnston, N.M. See Clarke, A., 47–114, 100, 298, 304 Johnstone, R.W., 286, 302, 306 Jolivet, E., 141, 142, 153 Jollivet, D. See Creasey, S., 36 Jones, E.G., 319, 320, 330, 337, 350 Jones, M.L., 327, 350 Jones, M.S. See Hamner, W.M., 305 Jonsgåard, A., 364, 373, 383 Jørgensen, B.B., 7, 39; See Schulz, H.N., 43; See Fossing, H., 37 Jørgensen, L.L., 164, 167, 195, 230 Jorissen, F., 15, 21, 39 Jorissen, F.J., 21, 39; See Fontanier, C., 37 Josselyn, M.N., 118, 153 Jouin-Toulmond, C. See Müller, M.C., 41 Jumars, P.A., 22, 23, 26, 39, 317, 350; See Smith, C.R., 352 Juniper, S.K. See Sarrazin, J., 329, 352; See Tunnicliffe, V., 338, 343, 353 Jurasz, W. See Jazdzewski, K., 106 Kahler, A. See Pfannkuche, O., 42 Kaiho, K., 15, 39 Kajimura, H. See Gearin, P.J., 382 Kallin, E. See Amundin, M., 379 Kallin, S. See Amundin, M., 379 Kamermans, P., 190, 204, 230 Kamykowski, D., 2, 3, 5, 39 Kan, H. See Tanabe, S., 388

402

AUTHOR INDEX

Kang, Y.-C. See Ahn, I.-Y., 70, 97 Kann, L. See Wishner, K.F., 44 Kaplan, I.R. See Rau, G.H., 352 Karaman, G.S. See Barnard, J.L., 98 Karl, D.M., 20, 39; See Martin, J.H., 40 Karlson, K., 2, 12, 13, 14, 33, 39 Kastelein, R.A., 360, 370, 372, 383 Kastendiek, J. See Schroeter, S.C., 234 Katona, S., 333, 350 Kawaguchi, T. See Matsunaga, K., 232 Kawamura, A. See Gaskin, D.E., 382 Kawamura, K. See Wada, H., 353 Kayes, R.J., 375, 384 Kazungu, J. See Hemminga, M.A., 152 Keesing, J.K. See Wells, F.E., 138, 159 Keil, R.G., 6, 39 Kelce, W.R. See Gray, L.E., 382 Keller, C., 239, 255, 275 Keller, N.B., 60, 106 Keller, R.A. See Klinkhammer, G.P., 106 Kemp, P.F., 6, 39 Kemp, W.M. See Gardner, R.H., 229 Kempers, A.J. See Moodley, L., 41 Kempers, L. See Moodley, L., 41 Kendrick, G.A., 170, 173, 178, 179, 180, 188, 207, 210, 217, 221, 231; See Kirkman, H., 117, 118, 135, 140, 141, 153; See Walker, D.I., 163, 236 Kennedy, J. See Hyland, J., 39 Kennedy, S. See Kuiken, T., 384 Kennedy, V.S. See Gardner, R.H., 229 Kennelly, S.J., 169, 173, 178, 187, 191, 194, 231 Kennett, J.P. See Behl, R.J., 32, 35; See Cannariato, K.G., 5, 36 Kensley, B., 128, 131, 132, 133, 153 Kenyon, K.W. See Wilke, F., 360, 370, 390 Kerr, S.R. See Kirsch, P.E., 384 Kester, D.R. See Capuzzo, J.M., 227 Key, R. See Davidson, N.S., 150 Khairallah, N.H. See Bogi, C., 258, 273 Khripounoff, A. See Desbruyeres, D., 349 Kikuchi, T., 118, 153 Kiliaan P.H.L. See Andriashek, D., 148 Kim, K.H., 29, 39 Kim, K.Y., 166, 177, 193, 194, 231 Kimbrell, C. See Hunter, J.R., 39 King, R.J., 184, 231; See Clayton, M.N., 283, 304 Kingsford, M.J. See Gillanders, B.M., 147, 151, 166, 170, 229 Kingsley, M.C. See Fontaine, P.M., 381 Kinkel, H. See Kuypers, M.M.M., 40 Kinne, O., 163, 231 Kinze, C.C., 356, 357, 360, 361, 366, 368, 375, 377, 384; See Aarefjord, H., 379 Kirby, P., 143, 153 Kirkman, H., 117, 118, 135, 140, 141, 153

Kirkpatrick, R.L. See Merson, M.H., 377, 385 Kirkwood, J.K. See Jepson, P.D., 383; See Kuiken, T., 384 Kirkwood, J.M., 65, 106 Kirsch, P.E., 359, 384 Kitazato, H., 330, 333, 343, 350 Kitching, J.A., 176, 231; See Lilly, S.J., 231; See Muntz, L., 233; See Norton, T.A., 233; See Round, F.E., 234 Kittel, W. See Jazdzewski, K., 106 Klages, M., 65, 106, 336, 350; See Arntz, W.E., 97; See Brey, T., 99 Klages, N.T.W. See Sekiguchi, K., 387 Klein, M. See Dubilier, N., 37 Kleivane, L., 375, 384 Klingelhoeffer, E. See Gallardo, V.A., 38 Klinkhammer, G.P., 71, 106; See Chin, C.S., 100 Klok, C.J., 133, 153 Klöser, H. See Clayton, M.N., 101 Klug, J.L. See Fischer, J.M., 151 Klumpp, D.W., 281, 287, 290, 296, 297, 302, 306; See Hansen, J.A., 305; See Polunin, N.V.C., 290, 297, 307 Knauer, G.A. See Karl, D.M., 20, 39; See Martin, J.H., 40 Knollenberg, W.G. See Wilson, D.S., 129, 159 Knopf, F.L. See Grover, P.B., 136, 151 Knowlton, N., 50, 106 Knox, G.A., 70, 79, 106 Kocher, T.D. See Rosel, P.E., 387 Kock, K.H., 368, 384 Koehler, R., 73, 106 Koeman, J.H. See Brouwer, A., 380 Kohda, M. See Sakai, Y., 291, 308 Koltun, V.M., 58, 106; See Gutt, J., 59, 104 Kompfrner, H., 134, 153 Konar, B., 165, 176, 178, 193, 231 Koop, K., 117, 118, 119, 123, 124, 125, 131, 132, 153, 289, 302, 306; See Grifﬁths, C.L., 151; See Johnstone, R.W., 306 Koopman, H.N., 359, 384; See Westgate, A.J., 389 Koskela, H., 129, 153 Kott, P., 75, 106 Koumjian, L. See Valiela, I., 158 Koutsoukos, E.A.M., 25, 40 Kouwenhoen, T.J. See van der Zwaan, G.I., 44 Kramp, P.L., 60, 106 Krause, L. See Bugoni, L., 149 Krause, P.R. See Raimondi, P.T., 234 Krebs, C.J. See Carpenter, S.R., 227 Krebs, W.N. See Lipps, J.H., 107 Kreutzer, U. See Grieshaber, M.K., 38 Krieger, J. See Giere, O., 31, 38; See Dubilier, N., 37 Krogh, A., 312, 315, 350

403

AUTHOR INDEX

Kropp, R. See Hyland, J., 39 Kudenov, J.D. 70, 106 Kuenzel, N.T. See Johannes, R.E., 306 Kuiken, T., 367, 368, 377, 384; See Jepson, P.D., 383 Kukert, H., 315, 351; See Allison, P.A., 348; See Smith, C.R., 352 Kuklik, I. See Malinga, M., 360, 361, 385 Kulbicki, M., 300, 306 Kunzmann, K., 59, 62, 64, 66, 70, 106 Kuo, J., 118, 153; See Kirkman, H., 118, 153 Kuo, S., 290, 306 Küver, J. See Fossing, H., 37 Kuypers, M.M.M., 8, 40 Kuznetsova, I.A., 65, 106 Labberté, S. See Lockyer, C., 385 Labelle, M., 298, 306 Lai, N.C. See Yang, T.H., 45 Laidig, T.E. See Adams, P.B., 34 Laist, D.W., 116, 144, 154 Lakaseru, B.O. See Willoughby, N.G., 159 Lamarck, [J.B. de], 260, 275 Lambert, L., 260, 275 Lambright, C. See Gray, L.E., 382 Lambshead, P.J. See Cook, A.A., 36 Lambshead, P.J.D., 50, 106; See Boucher, G., 50, 98 Lamont, P. See Levin, L.A., 40 Lamont, P.A., 9, 40; See Levin, L.A., 40 Lamothe, F. See Ackamn, R.G., 359, 379 Lange, C. See Reimers, C., 42 Langham, L.P.E., 364, 371, 384 Lankester, K. See Northridge, S.P., 357, 377, 385 Lapègue, S. See Fabioux, C., 274 Lapointe, B.E. See Littler, M.M., 231 Larcombe, P., 162, 131 Larkum, A.W.D., 286, 287, 306; See Johnstone, R.W., 306; See Koop, K., 289, 302, 306 Larrain, A.P., 73, 106 Lasiak, T. See Madzena, A., 145, 154 Lasiak, T.A. See McLachlan, A., 155 Lassus, P. See Holmer, M., 230 Latham, R.E. See Petraitis, P.S., 233 Laubier, L. See Desbruyeres, D., 338, 349 Laubitz, D.R., 62, 106 Laughlin, J. See Thompson, B., 44 Laur, D.R. See Ebeling, A.W., 228; See Reed, D.C., 234 Lavaleije, M.S.S. See Kastelein, R.A., 360, 383 Lavie, B., 267, 275 Lavoie, D., 126, 127, 128, 138, 154 Law, R.J. See Jepson, P.D., 383; See Kuiken, T., 384 Lawson, G.W. See Towsend, C., 165, 187, 235 Lawver, L. See Chin, C.S., 100

Lazzari, G., 261, 275 Leach, J.H., 116, 154 Leaper, R. See Tregenza, N.J.C., 389 Leary, P.N. See Koutsoukos, E.A.M., 40 Leatherwood, S. See Jones, M.L., 350 Lee, C., 6, 40 Lee, D.J. See Turner, S.J., 235 Lee, M.S.Y., 49, 106 Leecaster, M.K. See Moore, S.L., 155 Lefevre, J.R. See Meinesz, A., 232 Leggett, M. See Phillips, D.S., 156 Leggett, M.C., 134, 154 Legner, E.F. See Moore, I., 120, 126, 127, 128, 155 Leistikow, A. See Coleman, C.O., 101 Lemly, A.D. See Henley, W.F., 230 Lenanton, R.C.J., 118, 138, 154; See Robertson, A.I., 118, 138, 157 Lenihan, H.S., 190, 231 Lentz, L.F. See Kamermans, P., 230 Leopold, M.F. See Camphuysen, C.J., 364, 380; See Hammond, P.S., 382; See Smeenk, C., 388 Leta, A.C. See Young, P.S., 65, 113 Leutenegger, S., 9, 40 Levenstein, R.Y., 70, 107 Lévi, C., 59, 107 Levin, L. See Wishner, K., 45 Levin, L.A., 1–45, 2, 5, 6, 7, 11, 14, 15, 16, 17, 18, 19, 20, 21, 22, 23, 24, 25, 26, 27, 29, 30, 31, 32, 33, 40, 253, 275, 317, 322, 328, 338, 351; See Cook, A.A., 36; See Gage, J.D., 37; See Gooday, A.J., 38; See Neira, C., 41; See Smith, C.R., 43, 352; See Wishner, K.F., 44 Levin, S.A., 223, 231 Levine, J.M. See Brewer, J.S., 148 Levinton, J.S. See Lopez, G.R., 154 Lewis, J.R., 177, 231 Lewis, M.S. See Taylor J.D., 118, 158 Licari, L. See Fontanier, C., 37 Lick, R. See Siebert, U., 387 Lick, R.R., 360, 361, 363, 365, 370, 385 Lidgard, S. See Clarke, A., 81, 90, 101 Light, K. See Harrold, C., 350 Lilly, S.J., 167, 170, 177, 231 Lim, G.S.Y., 191, 231 Lindberg, D.R. See Jacobs, D.K., 5, 32, 33, 39; See Ponder, W.F., 67, 110 Lindroth, A., 357, 360, 373, 377, 385 Lindstedt, I. See Aarefjord, H., 379 Linke, P., 9, 40; See Schmaljohann, R., 43 Linley, E. See Field, J.G., 151 Linse, K. See Steiner, G., 68, 112 Lipps, J.H., 53, 91, 107 Lipschitz, S.R. See Bailey, G.W., 35 Lipschultz, F. See Ward, B.B., 44 Lisin, S. See Harrold, C., 350

404

AUTHOR INDEX

Little, C., 172, 177, 194, 231 Littler, D.S. See Littler, M.M., 231 Littler, M.M., 165, 167, 168, 170, 172, 173, 176, 177, 178, 180, 182, 186, 193, 194, 196, 221, 222, 224, 231; See Seapy, R.R., 166, 176, 177, 189, 190, 193, 194, 195, 234; See Taylor, P.R., 170, 180, 182, 194, 222, 235 Livermore, R.A. See German, C.R., 103 Livingstone, D.R., 12, 40 Llewellyn, M.J. See Braﬁeld, A.E., 121, 148 Llewellyn, P.J., 140, 143, 144, 154 Locard, A., 266, 275 Lockyer, C., 312, 334, 351, 356, 372, 385; See Berrow, S.D., 380; See Santos, M.B., 387 Lockyer, C.H. See Kuiken, T., 384 Long, S.C. See Berrow, S.D., 380 Longhurst, A., 53, 107 Lonsdale, P., 313, 351 Lopez, G.R., 131, 154 López, C.M., See Arnaud, P.M., 97 López Gappa, J., 91, 107 Lopez-Uriate, E., 136, 154 Lord, W.D., 121, 130, 154 Lorenti, M. See Gambi, M.C., 103 Lou D.C. See Choat, J.H., 304 Loubere, P., 15, 40 Lourenco, S.O., 285, 286, 306 Lourens, L.J. See Reichart, G.L., 42 Loutit, R. See Avery, G., 148 Løvås, S.M., 140, 154 Loveland, R.E. See Botton, M.L. 148 Loveren, H.V. See Swart, R. de L., 388 Lowry, J.K. See Knox, G.A., 70, 79, 106 Lowry, L.F. See Iverson, S.J., 383 Lowry, N., 357, 377, 385 Lozouet, P. See Bouchet, P., 98 Lucarelli, E. See Chelazzi, L., 149 Lucas, M.I. See Koop, K., 153 Lucas J.S. See Robertson, A.I., 127, 132, 138, 157 Lumb, C.M., 166, 231 Lund-Hansen, L.C., 173, 231 Lupton, J.E. See Chin, C.S., 100 Luster, M.I. See Vos, J.G., 376, 389 Lutz, E.V. See Bowen, S.H., 304 Lutz, R., 333, 351 Lutz, R.A. See Black, M.B., 348; See Craddock, C., 349; See Jablonski, D., 253, 275; See Peek, A.S., 352 Lutze, G.F. See Linke, P., 9, 40 Lyngby, J.E., 201, 231 Lynn, E.A. See Hunter, J.R., 39; See Vetter, R.D., 44 Maas, P.A.Y. See Black, M.B., 348 MacAlister, H.E. See Moore, P.G., 155; See

Pugh, P.J.A., 123, 133, 156 Macan, T.T. See Kitching, J.A., 231 MacAndrew, R., 238, 275 MacArthur, R.H., 48, 49, 85, 107 MacDonald, D.W. See Hersteinsson, P., 122, 152 MacDonald, L.H., 166, 232 Macer, C.T., 364, 371, 385 Machain-Castillo, M.L. See Perez-Cruz, L.L., 42; See Sen Gupta, B.K., 43 Macintyre, D., 373, 385 Macintyre, I.G. See Steneck, R.S., 235 Maciolek, N.J. See Grassle, J.F., 50, 103 Macko, S.A. See Deming, J., 349; See Fisher, C.R., 349 Madsen, F.J., 73, 107 Madzena, A., 145, 154 Maestrati, P. See Bouchet, P., 98 Magaard, L. See Smith, C.R., 43 Magri, M. See Piazzi, L., 233 Magurran, A.E., 48, 95, 107 Malinga, M., 360, 361, 385 Malmgren, B.A. See Naidu, P.D., 15, 41 Malta, E.-J. See Kamermans, P., 230 Manahan, D.T. See Marsh, A.G., 351 Manghnani, V. See Morrison, J.M., 41 Mangi, S. See McClanahan, T.R., 307 Manheim, F., 6, 40 Mann, K.H., 117, 137, 154, 288, 306; See Harrison, P.G., 125, 152; See Robertson, A.I., 125, 126, 131, 132, 157 Mann, R., 267, 275 Manor-Samsonov, N. See Bowman, D., 148 Manunza, B. See Giuseppetti, G., 274 Marbà, N. See Vermaat, J.E., 236 Marchessaux, D., 356, 385 Marcomini, A. See Sfriso, A., 189, 235 Margulis, L., 51, 58, 107 Markham, J.W., 165, 167, 172, 180, 181, 182, 186, 197, 198, 232 Marnane, M.J., 290, 306 Marquez, U.M.L. See Lourenco, S.O., 306 Marra, J. See Smith, S.L., 44 Marsden, I.D., 118, 119, 128, 131, 154 Marsden, K. See Holden, A.V., 377, 383 Marsh, A.G., 333, 351 Marshall, B.A., 312, 313, 314, 325, 337, 340, 341, 343, 351 Marshall, D.J., 170, 184, 186, 192, 200, 232 Marshall, N., 289, 302, 306; See Gerber, R.P., 289, 302, 305 Martill, D.M., 338, 343, 351 Martin, A.R., 361, 362, 363, 366, 370, 385 Martin, C. See Levin, L.A., 40 Martin, C.M. See Levin, L.A., 351 Martin, J.H., 6, 39, 40; See Karl, D.M., 39

405

AUTHOR INDEX

Martinez-Taberner, A. See Fornos, J.J., 229 Martini, F.H., 318, 351 Marty, J.C. See Viso, A.C., 285, 308 Martz, D.R. See Littler, M.M., 231 Marzluff, J.M. See Dial, K.P., 92, 102 Mascagni, A. See Chelazzi, L., 149 Mascarenhas, A. See Paropkari, A.L., 42 Masini, R.J. See Walker, D.I., 159 Mason, T.R. See Ekdale A.A., 32, 37 Massuti, E. See Golani, D., 274 Mateo, R. See Tomás, J., 158 Mathew, K.J. See Thomas, P.A., 60, 112 Mathieson, A.C., 165, 167, 170, 172, 176, 181, 194, 232; See Daly, M.A., 165, 167, 168, 170, 172, 173, 176, 177, 182, 186, 194, 196, 221, 228 Matsuike, K. See Arakawa, H., 196, 201, 206, 220, 226 Matsunaga, K., 178, 232 Matthews, L.H., 359, 385 Mattson, D.J. See Green, G., 350 Maughan, B.C., 173, 178, 207, 214, 217, 232 May, R.M., 50, 107 Maybaum, H.L. See Bennett, B.A., 348 Mayer, L.M. See Shull, D.H., 29, 43 Mayor, S. See Gutiérrez, D., 38 Mayzaud, P. See Alonzo, F., 303 Mazzella, L. See Gambi, M.C., 103 McArthur, A.G. See Tunnicliffe, V., 353 McCall, P.L. See Rhoads, D.C., 42, 352 McClain, C.R., 14, 41 McClanahan, T.R., 162, 166, 232, 281, 297, 306, 307 McComb, A.J. See Kuo, J., 118, 154 McConochie, J. See Crossman, D.J., 304 McCook, L.J. See Umar, M.J., 235 McDougall, K. See Mullins, H.T., 41 McField, M. See McClanahan, T.R., 306 McGarry, A.T. See Berrow, S.D., 380 McGuinness, K.A., 173, 191, 207, 208, 216, 217, 232 McGwynne, L.E., 121, 124, 154; See McLachlan, A., 118, 124, 155 McHatton, S.C., 7, 41 McHugh, D. See Tunnicliffe, V., 353 McIntyre, A.D., 124, 125, 154 McKenzie, C. See Wells, D.E., 377, 389 McKenzie, J.A., 238, 275 McKinnerney, M., 129, 130, 154 McKinnon, A.D. See Klumpp, D.W., 281, 287, 302, 306 McKnight, D.G., 73, 107 McLachlan, A., 118, 123, 124, 125, 126, 127, 139, 154, 155; See Brown, A.C., 116, 118, 121, 124, 125, 138, 146, 149; See McGwynne, L.E., 154; See Soares, A.G., 157 McLean, J.H., 314, 340, 341, 344, 351

McMurtry, G. See Smith, C.R., 43 McQuaid, C.D., 168, 195, 224, 232; See Bally, R., 164, 165, 226; See Marshall, D.J., 170, 184, 186, 192, 200, 232 Meadows, A., 28, 29, 41; See Murray, J.M.H., 41 Meadows, P.S. See Meadows, A., 41; See Murray, J.M.H., 41 Mearns, A.J. See Rau, G.H., 352 Meekan, M.G., 299, 307 Meier, D. See Coleman, C.O., 101 Meier, N.F. See van Katwijk, M.M., 236 Meinesz, A., 166, 232 Melin, S.R. See Gearin, P.J., 382 Melone, N. See Garavelli, C.L., 266, 274 Melville, A.J., 169, 178, 232 Memon, G.M. See den Dulk, M., 36 Menge, B.A., 165, 173, 189, 192, 232 Mente, E. See Santos, M.B., 387 Menzel, R.W., 260, 275 Menzies, R.J. See Frankenberg, D., 7, 22, 23, 37 Mercer, S. See Moss, B., 233 Merello, S.E. See Relini, G., 234 Merrett, N.R. See Kuiken, T., 384; See Priede, I.G., 352 Merritt, R.W., 332, 351 Merson, M.H., 377, 386 Mesías, J. See Gallardo, V.A., 37 Messana, G., 119, 155 Messeri, P. See Messana, G., 155 Mettam, C., 172, 177, 232 Meyer, J.L., 290, 307 Meyer-Reil, L.A. See Azam, F., 303 Michener, A.E. See Levin, L.A., 351 Mienis, H.K., 254, 258, 275 Miles, S. See Day, T.H., 150 Millais, J.G., 359, 385 Miller, A. See Pierce, G.J., 386 Miller, D. See Pierce, G.J., 386; See Thompson, P.M., 388 Miller, M.W. See Sluka, R.D., 298, 308 Miller, R.G., 77, 107 Miller, R.J., 191, 232 Miller, S.E. See Eldredge, L.G., 102 Minelli, A., 50, 51, 65, 92, 107 Minigawa, M., 326, 328, 351 Mironov, A.N., 73, 107 Mitchell, N. See Cook, A.A., 36 Mitchell, P. See Hodge, S., 153 Mitchell, R. See Davidson, N.S., 150; See Ducklow, H.W., 287, 289, 305 Mitchell, R.E. See Clarke, A., 100 Miyagi, V.K. See Wakabara, Y., 113 Miyazaki, N., 356, 385; See Amano, M., 356, 379; See Subramanian, A.N., 388 Mizroch, S.A., 336, 351 Mizzan, L., 261, 275

406

AUTHOR INDEX

Moazzo, G. See Gruvel, A., 256, 274 Moazzo, P.G., 239, 254, 255, 256, 275 Mocogni, M. See Velander, K.A., 145, 158 Moe, R.L., 58, 87, 88, 107 Moggi, G., 119, 120, 155 Molander, A.R., 60, 107 Moloney, C.L. See Ryan, P.G., 116, 144, 157 Monaghan, P., 364, 373, 375, 385 Monniot, C., 75, 107, 108 Monniot, F. See Monniot, C., 75, 107, 108 Monro, C.C.A., 70, 108 Montagne, D.E., 28, 41 Montgomery, W.L., 286, 307 Moodley, L., 15, 21, 41 Moore, G.T. See Demaison, G.J., 6, 36 Moore, I., 120, 126, 127, 128, 155 Moore, P.G., 122, 131, 132, 155, 163, 164, 165, 167, 169, 173, 177, 190, 232, 233; See Agnew, D.J., 131, 132, 147; See Bedford, A.P., 126, 131, 148 Moore, S.L., 145, 155 Moran, P.J., 191, 193, 233 Moreno, R.J. See Vargas, M.A., 308 Moreton, S.G. See Pike, J., 42 Moriarty, C.M. See Moriarty, D.J.W., 307 Moriarty, D.J.W., 284, 286, 297, 307 Morinaga, T. See Arakawa, H., 201, 220, 226 Moring, J.R., 165, 172, 178, 190, 233 Morot-Gaudry, J.F. See Jolivet, E., 153 Morri, C., 239, 275 Morril, W. See Distel, D.L., 349 Morris, P. See German, C.R., 103 Morrison, J.M., 4, 20, 41 Morrison, J.P.E., 262, 276 Morritt, D., 121, 155 Morse, J.W. See Rhoads, D.C., 2, 21, 42 Morse-Porteous, L.S. See Grassle, J.F., 315, 338, 349 Mortensen, S.M. See Lyngby, J.E., 201, 231 Mortensen, T.H., 73, 108 Morton, B., 122, 123, 155, 267, 276; See Britton, J.C., 122, 148, 149, 332, 348 Mosbech, A. See Heide-Jørgensen, M.-P., 382 Moschella, P.S. See Benedetti-Cecchi, L., 226 Moss, B., 196, 198, 233 Mou Tham, G. See Kulbicki, M., 306 Moyano, G.H.I., 72, 91, 108 Mühlenhardt-Siegel, U., 67, 108 Mulders, C. See Dubilier, N., 37 Müller, M.C., 12, 16, 24, 41 Mullineaux, L. See Wishner, K., 45; See Wishner, K.F., 44 Mullineaux, L.S., 340, 351; See Marsh, A.G., 351; See Smith, C.R., 352 Mullins, H.T., 16, 17, 20, 22, 41; See Thompson, J.B., 44

Mulssow, S.G. See Rhoads, D.C., 42 Munilla León, T., 65, 66, 79, 108 Munkejord, A.A. See Berg, B.S., 140, 141, 148 Munoz, P. See Levin, L.A., 40 Muntz, L., 217, 233 Murina, V.V., 62, 108 Murphy, E.J. See Trathan, P.N., 112 Murray, J., 376, 385 Murray, J.M.H., 10, 29, 41; See Meadows, A., 41 Murray, J.W., 20, 41 Murray, S.N. See Horn, M.H., 306; See Littler, M.M., 193, 231 Muthiga, N.A. See McClanahan, T.R., 307 Muyakshin, S. See Klages, M., 350 Myers, J.P. See Castro, G., 123, 149 Myers, R.F., 293, 307 Nachtigall, P.E. See Read, A.J., 386 Naganuma, R., 317, 351 Naganuma, T. See Wada, H., 353 Naidu, P.D., 15, 41 Nair, R.R. See Haake, B., 38 Napolitano, G.E., 286, 307 Naqvi, S.W.A. See Morrison, J.M., 41 Naranjo, S.A., 194, 233; See Carballo, J.L., 227 Navarrete, S.A. See Menge, B.A., 232 Nealler, E. See Castilla, J.C., 166, 193, 194, 227 Neira, C., 6, 10, 15, 16, 21, 24, 25, 26, 27, 41; See Gutiérrez, D., 38; See Levin, L.A., 40 Neira, F. See Retamal, M.A., 110 Nelson, D.C. See McHatton, S.C., 41 Nelson, S.G., 295, 307 Neumann, A.C., 187, 233 Neurova, A.M. See Kuznetsova, I.A., 65, 106 Neushul, M., 207, 208, 216, 233; See Clarke, W.D., 177, 228 Neves, R.J. See Henley, W.F., 230 Nevo, E. See Lavie, B., 267, 275 Newell, R., 284, 307 Newell, R.C., 117, 123, 124, 137, 155; See Koop, K., 153; See Seiderer, L.J., 117, 157; See Stuart, V., 158 Newell, S.Y., 125, 155 Newman, W.A., 65, 108 Newroth, P.R. See Markham, J.W., 167, 180, 182, 186, 197, 198, 232 Newton, C.R. See Thompson, J.B., 44 Nichols, G. See Perkins, J.S., 386 Nichols, P.D. See Volkman, J.K., 308 Nicol, D., 50, 67, 94, 108 Nicolas, J.R. See Waring, G.T., 389 Nicolay, K., 239, 258, 262, 276 Nielsen, C., 51, 65, 71, 108 Nielsen, J. See Whitehead, P.J.P., 389 Nielsen, L.P. See Fossing, H., 37 Nienstedt, J.C., 14, 25, 41

407

AUTHOR INDEX

Niesen, T.M. See Josselyn, M.N., 153 Niesenbaum, R.A. See Petraitis, P.S., 233 Nieuwenhuize, J. See Hemminga, M.A., 118, 125, 140, 152 Nigi, G. See Matsunaga, K., 232 Niles, L.J. See Clark, K.E., 150 Nilsson-Cantell, C.A., 65, 108 Nishihira, M. See Tsuchiya, M., 196, 235 Noel, P. See Goulletquer, P., 274 Nordemar, I. See McClanahan, T.R., 306 Norse, E.A. See Carlton, J.T., 273 Northridge, S.P., 357, 364, 377, 385 Norton, M.G. See Eagle, R.A., 228 Norton, T.A., 177, 196, 199, 206, 233; See Muntz, L., 233; See Vadas, R.L., 236 Nose, Y. See Sano, M., 308 Nowlin, W.D. See Orsi, A.H., 109 Numanami, H., 67, 108 Nunny, R.S. See Eagle, R.A., 228 Nursall, J.R. See Labelle, M., 298, 306 Nyden, B.B. See Williams, S.L., 236 Nyström, M. See McClanahan, T.R., 306 O’Brien, C.M. See Pinnegar, J.K., 386 Obura, D. See McClanahan, T.R., 162, 166, 232 Obura, D.O., 118, 155 Occhipinti, A. See Mann, R., 275 Occhipinti-Ambrogi, A., 261, 276 Ochieng, C.A., 118, 119, 125, 126, 155 Odendaal, F.J. See Brown, A.C., 133, 149 O’Donoghue, C.H., 239, 256, 276 Ó Foighil, D., 260, 276 Oftedal, O.T. See Iverson, S.J., 383 Ogden, J.C. See Thayer, G.W., 158 Ohme, U. See Santelices, B., 234 Øien, N. See Bjørge, A., 357, 380; See Hammond, P.S., 382; See Tolley, K.A., 388 Okano, H. See Fujioka, K., 349 Okutani, T. See Numanami, H., 108 Olaso, I. See Arnaud, P.M., 97 Olin, R., 145, 155 Olivares, H. See Vargas, M.A., 308 Oliver, P.G., 10, 24, 67, 41, 108 Oliverio, M., 258, 269, 276 Olivi, G., 253, 255, 276 Olson, R.J. See Rau, G.H., 352 Olu, K. See Sibuet, M., 313, 329, 340, 352 Oppen-Bernsten, D.O. See Jelmert, A., 336, 338, 350 Or, Y.Y. See Gray, J.S., 38 Orellana, M.V. See Chin, W., 304 Oren, O.H., 257, 276 Orensanz, J.M., 70, 109 Orlin, Z. See Dekker, H., 269, 271, 273 Orosco, C. See Schramm, W., 308

Orphan, V.J.H. See Hinrichs, K.U., 38 Orsi, A.H., 55, 56, 109 Orsi-Relini, L. See Golani, D., 274 Ostby, J. See Gray, L.E., 382 Osterhaus, A.D.M.E. See Ross, P.S., 387; See Swart, R. de L., 388 Ostrom, P. See Hooker, S.K., 383 Ostrovskii, A.N., 72, 109 O’Sullivan, D., 60, 109 Otsuki, A., 288, 307 Otterlind, G., 359, 375, 385 Owen, R.W., 20, 42 Ozturk, B. See Zaitsev, Y., 267, 277 Pacheco-Ruíz, I., 141, 142, 156 Palacios E. See Lopez-Uriate, E., 154 Paling, E.I. See Walker, D.I., 159 Palka, D.L., 364, 385 Pallary, P., 238, 254, 256, 263, 276 Palma, A.T. See Poulin, E., 110 Palmer-Zwahlen, M.L., 187, 233 Palmisano, J.F. See Estes, J.A., 169, 229 Palumbi, S.R. See Butman, C.A., 348 Pancost, R.D. See Kuypers, M.M.M., 40 Pannacciulli, F. See Benedetti-Cecchi, L., 226 Pansini, M., 59, 109 Parapar, J. See San Martin, G., 70, 111 Pardi, L. See Messana, G., 155 Park, P.K. See Capuzzo, J.M., 227 Parkin, D.T. See Day, T.H., 150 Parkins, C.A. See Pulfrich, A., 234 Parkinson, C.L. See Gloersen, P., 103 Parks, D.S. See Shaffer, J.A., 165, 177, 178, 224, 235 Parnell, P.E. See Seymour, R.J., 234 Paropkari, A.L., 6, 42 Parry, B.L. See Waring, G.T., 389 Parson, L.M. See Van Dover, C.L., 353 Parulekar, A.H. See Jaya Sree, V., 105; See Sreepada, R.A., 112 Pashtan, A. See Ritte, U., 267, 277 Pasko, D. See Zmarzly, D.L., 354 Passow, U., 288, 289, 307 Pasternak, F.A., 60, 109 Patience, A. See Levin, L.A., 40; See Smallwood, B.J., 43 Patten, D.R. See Perkins, J.S., 386 Patterson, I.A.P. See Santos, M.B., 387 Patterson, M.A. See Henley, W.F., 230 Patton, J.S. See Harvey, H.R., 38 Paul, V.J., 132, 156 Paulay, G., 77, 109 Pauly, D., 359, 386 Pawson, D. See Levin, L.A., 40 Pawson, D.L., 73, 109 Payne, P.M. See Waring, G.T., 389

408

AUTHOR INDEX

Pearson, T.H., 28, 42, 163, 233, 315, 317, 322, 338, 351; See Tyson, R.V., 5, 44 Peck, L.S. See Chapelle, G., 14, 94, 36, 99 Pedersén, M., 177, 233; See See Johansson, G., 230 Pederson, T.F., 6, 42; See Calvert, S.E., 35; See Cowie, G.L., 36; Peek, A.S., 343, 352; See Baco, A.R., 348 Peek, J.M. See Green, G., 350 Peirano, A. See Morri, C., 275 Pejrup, M. See Lund-Hansen, L.C., 231 Pellizzato, M. See Cesari, P., 261, 273 Pelorce, J., 242, 276 Peña Cantero, A.L., 60, 109, 110 Pennings, S.C., 131, 133, 139, 156 Percival, S.M., 136, 156 Pérès, J.M., 177, 233 Perez-Cruz, L.L., 8, 25, 42 Perkins, J.S., 364, 373, 386 Pernthaler, A. See Dubilier, N., 37 Perrin, M.R. See Hiscocks, K., 122, 152 Perry, S.L., 333, 352 Persson, L.E., 131, 156 Peter, F.M. See Wilson, E.O., 48, 113 Petersen, J.E. See Gardner, R.H., 229 Petersen, S. See Klinkhammer, G.P., 106 Petit de la Saussaye, S., 242, 276 Petraitis, P.S., 196, 223, 233 Petrecca, R.F. See Snelgrove, P.V.R., 353 Petry, M.V. See Bugoni, L., 149 Pettibone, M.H., 314, 338, 341, 352 Petzold, D. See Coleman, C.O., 101 Pezzoli, G. See Colombini, I., 150 Pfannkuche, O., 6, 42 Philipp, R., 116, 156 Phillips, D.S., 127, 131, 134, 156; See Leggett, M.C., 154 Phillips, G.A. See Hargrave, B.T., 350 Phillips, S. See Kuiken, T., 384 Phleger, P.B., 8, 9, 24, 25, 42 Piatkowski, U., See Allcock, A.L., 68, 97 Piatt, J.F. See van Pelt, T.I., 122, 158 Piazzi, L., 182, 186, 188, 233 Picard, J. See Pérès, J.M., 177, 233 Picken, G.B., 94, 95, 110; See Oliver, P.G., 67, 108 Picton, B.E. See Costello, M.J., 101; See Howson, C.M., 83, 84, 105 Pielou, E.C., 95, 110, 335, 352 Pienkowski, M.W. See Davidson, N.S., 150 Piepenburg, D., 73, 110; See Gutt, J., 93, 104 Pierce, G.J., 358, 364, 367, 372, 373, 377, 386; See Santos, M.B., 355–390, 387; See Thompson, P.M., 388; See Wijnsma, G., 390 Pierpoint, C., 369, 386 Pike, J., 29, 42

Pilkington, M.D. See Day, T.H., 150 Pineda, J., 165, 180, 182, 200, 233; See Levin, L.A., 40 Pineda, V. See Gallardo, V.A., 37 Pinnegar, J.K., 375, 386 Piper, D.Z., 6, 42 Pirez, A.M.S., 64, 110 Pirro, F. See Siebert, U., 387 Plaia, G.R. See Levin, L.A., 40, 351 Platnick, N.I., 85, 110 Plaziat, J.C., 238, 276 Polis, G.A., 115, 119, 121, 136, 137, 138, 156; See Anderson, W.B., 137, 148; See Due, A.D., 137, 151; See Rose, M.D., 122, 136, 137, 157 Pollard, D. See Berrow, S.D., 380 Pollard, P.C. See Moriarty, D.J.W., 307 Polloni, P.T. See Haedrich, R.L., 38 Polunin, N.V.C., 280, 290, 295, 296, 297, 307; See Jennings, S., 306; See Klumpp, D.W.,290, 296, 297, 306 ; See Pinnegar, J.K., 386 Pond, K. See Philipp, R., 156; See Rees, G., 145, 156 Ponder, W.F., 67, 110 Poore, G.C.B., 50, 90, 110 Pope, R.H. See Smith, C.R., 43 Por, F.D., 239, 252, 258, 276 Portner, H.O. See Grieshaber, M.K., 38 Poulin, E., 84, 92, 110 Powell, A.W.B., 67, 110 Power, M.E., 169, 233 Poydenot, F. See Birje, J., 227 Prakash Babu, C. See Paropkari, A.L., 42 Presler, E. See Jazdzewski, K., 106 Presler, P. See Arnaud, P.M., 97; See Jazdzewski, K., 106 Preston, F.W., 94, 110 Price, I.R. See Umar, M.J., 235 Price, N.B. See Pederson, T.F., 42 Priede, I.G., 336, 352; See Collins, M.A., 348; Jones, E.G., 350 Prieto, R. See Silva, M.A., 387 Prime, J.H. See Hammond, P.S., 373, 382 Prince, W.A., 140, 156 Pringle, C.M., 169, 233 Prosperie, L. See Morrison, J.M., 41 Prouse, N.J. See Hargrave, B.T., 350 Prusova, I. See Smith, S.L., 44 Pugh, P.J.A., 65, 110, 123, 133, 156 Pulfrich, A., 166, 177, 178, 191, 192, 194, 234 Pulliainen, E., 332, 352 Pulliam, H.R., 362, 386 Purcell, S.W., 173, 234, 282, 283, 284, 295, 296, 301, 307 Pushkin, A.F., 65, 110 Pybus, C. See Burrows, E.M., 189, 227

409

AUTHOR INDEX

Qasim, S.Q., 289, 307 Qasim, S.Z. See Sankaranarayanan, V.N., 27, 43 Quadri, P. See Cecalupo, A., 271, 273 Quignard, J.P. See Francour, P., 274; See Golani, D., 274 Quilty, P.G. See Feldmann, R.M., 65, 102 Quiñones, R. See Gallardo, V.A., 12, 13, 37 Quiñones, R.A. See Gonzalez, R.R., 38 Quintara, R. See Retamal, M.A., 110 Rabalais, N.N., 2, 33, 42 Rae, B.B., 360, 365, 369, 373, 374, 386 Raga, J.A. See Tomás, J., 158 Raimbault, R., 260, 276 Raimondi, P.T., 166, 202, 205, 233 Ramaswamy, V. See Haake, B., 38 Ramil, F. See Arnaud, P.M., 97 Ramirez-Llodra, E. See Tyler, P.A., 353 Ramos, A. See Arnaud, P.M., 97 Ramos-Esplá, A.A. See Arnaud, P.M., 97 Ramsing, N.B. See Fossing, H., 37 Ramsing, R. See Schulz, H., 43 Randall, J.E., 291, 298, 300, 308 Rasmussen, E.K. See Bach, H.K., 226 Rassmann, K., 139, 156 Rathburn, A.E. See Gooday, A.J., 21, 38; See Levin, L.A., 40, 351 Rau, G. See De Robertis, A., 36 Rau, G.H., 326, 329, 352 Rauschert, M., 62, 110 Ray, G.C., 115, 156 Razouls, S. See Alonzo, F., 303 Ré, P. See Boaventura, D., 227 Read, A.J., 355, 356, 357, 364, 369, 371, 376, 377, 386; See Gannon, D.P., 381; See Palka, D.L., 385; See Recchia, C.A., 361, 364, 365, 370, 386; See Smith, R.J., 360, 361, 364, 365, 388; See Westgate, A.J., 389 Read, I.L. See Butlin, R.K., 149 Reaka-Kudla, M.L., 50, 110; See Carlton, J.T., 273 Reay, P.J., 364, 371, 386 Recchia, C.A., 361, 364, 365, 370, 386 Reed, D.C., 190, 234 Reed-Anderson, T. See Fischer, J.M., 151 Rees, G., 145, 156; See Philipp, R., 156 Rees, T.K., 176, 234 Rees, W.J., 254, 276 Reichart, G.J. See den Dulk, M., 36 Reichart, G.L., 5, 42 Reid, D.G., 253, 276 Reid, K. See Walker, T.R., 159 Reid, R.J. See Santos, M.B., 387; See Thompson, P.M., 388 Reid, R.P. See Steneck, R.S., 235 Reid, W. See Schoenly, D., 330, 352; See Schoenly, K., 127, 129, 157

Reijnders, P.J.H., 377, 387; See Boon, J.P., 380; See Brouwer, A., 380; See Kleivane, L., 384; See Ross, P.S., 387; See Swart, R. de L., 388 Reimers, C., 5, 42 ; See Bernhard, J., 14, 15, 24, 35 Reise, K., 254, 267, 277 Reitsma, C.S. See Valiela, I., 140, 158 Relini, G., 207, 213, 217, 234; See BenedettiCecchi, L., 226 Remmert, K. See Coleman, C.O., 101 Ren, X.Q., 65, 110 Renaud, P.E., 165, 172, 173, 176, 190, 234 Rennet, N. See Thomson, J.A., 60, 112 Resig, J.M., 20, 42; See Glenn, C.R., 38 Retamal, M.A., 73, 110 Rex, M.A., 32, 42, 90, 111; See Levin, L.A., 40; See McClain, C.R., 14, 41 Reysenbach, A.L. See Deming, J., 349 Rhoads, D.C., 2, 21, 28, 32, 42, 315, 352 Riber, H.H. See Bach, H.K., 226 Ribic, C.A., 145, 157 Ricci, S., 122, 157 Rice, D. See Braham, H., 336, 348; See Gosho, M., 349 Rice, D.W., 334, 336, 352; See Mizroch, S.A., 351 Richards, C.L. See Pennings, S.C., 139, 156 Richardson, M.D., 87, 111 Richmond, R.H., 162, 234 Riddiford, N.J. See Harris, M.P., 364, 373, 382 Riddle, M.J. See Hansen, J.A., 305 Riggs, S.R. See Renaud, P.E., 234 Rinaldi, E. See Lazzari, G., 261, 275 Rindi, F. See Airoldi, L., 226 Rippe, H.T. See Kastelein, R.A., 383 Risk, M.J. See Cortés, J., 162, 166, 228 Ritte, U., 267, 277 Rivas, J. See Vargas, M.A., 308 Rivas, M. See Gutiérrez, D., 38 Rixen, T. See Haake, B., 38 Roa, R. See Gallardo, V.A., 37 Robba, E., 238, 277 Robbins, W.D. See Choat, J.H., 304 Robecchi Bricchetti, L., 142, 157 Roberts, C. See Konar, B., 165, 176, 178, 193, 231 Roberts, C.M., 272, 277 Roberts, D.E., 193, 234 Robertson, A.I., 117, 118, 125, 126, 127, 131, 132, 137, 138, 157; See Lenanton, R.C.J., 154 Robertson, D.R., 290, 291, 294, 295, 308; See Bailey, T.G., 290, 303 Robilliard, G.A., 61, 111 Robineau, D., 319, 322, 352 Robles, C., 173, 177, 190, 192, 194, 234 Roderick, G.K. See Baco, A.R., 348

410

AUTHOR INDEX

Rodhouse, P.G.K. See Allcock, A.L., 68, 97 Rodriguez, H. See Espinoza, J., 183, 189, 229; See Vargas, M.A., 308 Roelofs, E.M.B. See den Dulk, M., 36 Rogan, E., 360, 361, 363, 368, 370, 387; See Berrow, S.D., 380; See Jepson, P.D., 383 Rogers, A. See Creasey, S., 36 Rogers, A.D., 2, 5, 32, 42; See Creasey, S., 36 Rogers, C.S., 162, 166, 207, 220, 223, 234 Rogers, G.I. See Volkman, J.K., 308 Rolfe, M.S. See Eagle, R.A., 228 Romagna-Manoja, E. See Nicolay, K., 258, 276 Roman, M. See Smith, S.L., 44 Roman, M.R. See Gottfried, M., 289, 305 Ronan, T.E. See Lipps, J.H., 107 Ronconi, L. See Colombini, I., 150 Rose, M.D., 122, 136, 137, 157 Rosel, P.E., 356, 387 Rosel, P.E. See Tolley, K.A., 388 Rosen, B.R. See Crame, J.A., 90, 101 Rosenberg, R., 2, 3, 6, 12, 13, 14, 16, 18, 23, 33, 42, See Diaz, R.J., 36; See Karlson, K., 39; See Pearson, T.H., 28, 42, 163, 233, 315, 317, 322, 338, 351 Rosenfeldt, P. See Hartmann-Schröder, G., 70, 104, 105 Rosengarten, F., 121, 157 Rosenzweig, M.L., 48, 85, 86, 111 Rösner, H.-U. See Berghahn, R., 374, 380 Ross, A. See Newman, W.A., 65, 108 Ross, G.J.B., 357, 387 Ross, H.M., 357, 368, 373, 387; See Santos, M.B., 387; See Thompson, P.M., 388 Ross, P.S., 376, 387; See Swart, R. de L., 388 Rosso, A., 72, 111 Rossouw, G. See McLachlan, A., 155 Round, F.E., 190, 234 Rouse, G. See Fauchald, K., 70, 102 Rouse, I.P. See German, C.R., 103 Rowe, G.T., 7, 16, 17, 18, 20, 43; See Haedrich, R.L., 38; See Manheim, F., 40 Rowley, R.J. See Ebeling, A.W., 228 Roy, K., 49, 90, 111 Rueda, J.L., 266, 277 Rugh, D.J. See Shelden, K.E.W., 334, 336, 352 Russ, G.R., 297, 308 Russo, G. See Zenetos, A., 277 Russo, G.F. See Gambi, M.C., 103 Rutten, G.M.W. See Moodley, L., 41 Ryan, P.A., 162, 234 Ryan, P.G., 116, 143, 144, 157 Ryan, W.B.F. See Hsü, K.J., 275 Ryland, J.S., See Hayward, P.J., 83, 84, 105 Sabelli, B., 238, 277 Sahling, H. See Klinkhammer, G.P., 106

Saito, S. See Subramanian, A.N., 388 Saiz-Salinas, J.I., 68, 69, 111, 166, 173, 177, 178, 186, 193, 194, 234 Sakai, Y., 291, 308 Sala, E. See McClanahan, T.R., 306 Salamanca, M. See Levin, L.A., 40 Saltzman, J. See Wishner, K.F., 44 Salvat, B. See Arias-Gonzalez, J.E., 303 Salzwedel, H. See Arntz, W.E., 34; See Tarazona, J., 44 Sampugna, J. See Iverson, S.J., 383 San Martin, G., 70, 111 Sanders, H., 7, 14, 16, 17, 20, 25, 43 Sanders, H.L. See Rex, M.A., 111 Sanders, N.K., 11, 43 Sandnes, O.K. See Evans, R.A., 229 Sanﬁlippo, R. See Cantone, G., 70, 99 Sangkoyo, H. See Willoughby, N.G., 159 Sankaranarayanan, V.N., 27, 43; See Qasim, S.Q., 289, 307 Sano, M., 291, 292, 308 Santelices, B., 183, 190, 199, 234 Santolaria, A. See Gorostiaga, J.M., 230 Santos, M.B., 355–390, 356, 360, 361, 362, 363, 364, 365, 366, 368, 370, 372, 373, 374, 387; See Wijnsma, G., 390 Santos, R., 177, 190, 234 Sarà, M., 58, 59, 111; See Pansini, M., 109 Sarmiento, J.L., 3, 43 Sarrazin, J., 329, 352 Sarthou, G. See Klinkhammer, G.P., 106 Sartori, S. See Minelli, A., 107 Sauriau, P.G. See Goulletquer, P., 274 Savrda, C.E., 28, 29, 32, 43 Scapini, F. See Colombini, I., 150 Schafer, H.A. See Rau, G.H., 352 Schartau M. See Engel, A., 289, 305 Schaub, B.E.M. See Moodley, L., 41 Scheffer, V.B., 360, 370, 387 Schiaparelli, S. See Cattaneo-Vietti, R., 99 Schickan, T. See Gutt, J., 93, 104 Schiebling R.E., 138, 157 Schierl, F. See Coleman, C.O., 101 Schindler, D.W. See Carpenter, S.R., 227 Schlacher, T.A. See Soares, A.G., 157 Schmaljohann, R., 7, 43 Schmid, K. See Renaud, P.E., 234 Schoenly, D., 330, 352 Schoenly, K., 127, 129, 157 Scholtz, C.H. See Crafford, J.E., 133, 150 Schönfeld, J. See Baas, J.H., 35 Schooneman, N.M. See Kastelein, R.A., 383 Schouten, S. See Kuypers, M.M.M., 40 Schramm, W., 288, 297, 308; See Vogt, H., 165, 176, 190, 193, 236 Schrijvers, L. See Kamermans, P., 230

411

AUTHOR INDEX

Schroeter, S.C., 166, 173, 191, 192, 193, 234; See Sousa, W.P., 235 Schuetz, R.D. See Hart, H., 346, 350 Schulte, M.C. See Coleman, C.O., 101 Schultz, E.T. See Meyer, J.L., 307 Schultz, W. See Heide-Jørgensen, M.-P., 382 Schulz, H. See Cowie, G.L., 36; See Fossing, H., 37; See von Rad, U., 44 Schulz, H.N., 7, 43 Schulz, M., 135, 157 Schwartz, K.V. See Margulis, L., 107 Schwartzlose, R.A. See Isaacs, J.D., 317, 332, 337, 350 Scipione, M.B. See Gambi, M.C., 102 Scoccianti, C. See Ricci, S., 157 Scofﬁn, T.P., 168, 169, 181, 186, 187, 234; See Neumann, A.C., 233 Scott, T., 359, 387 Seapy, R.R., 166, 176, 177, 189, 190, 193, 194, 195, 234 Secilla, A. See Gorostiaga, J.M., 230 Seibel, B.A. See Childress, J.J., 9, 10, 11, 12, 36 Seiderer, L.J., 117, 157 Sekiguchi, K., 360, 361, 367, 369, 387 Seku, F.O.K. See Evans, S.M., 229 Sellanes, J., 16, 43; See Gutiérrez, D., 38; See Levin, L.A., 40; See Neira, C., 41 Selli, G., 254, 277 Sellner, K.G. See Iannuzzi, T.J., 230 Sen Gupta, B.K., 8, 15, 24, 43; See Bernhard, J., 2, 8, 9, 12, 15, 35; See Bernhard, J.M., 15, 20, 24, 25, 35 Seno, J., 73, 111 Sequeira, M., 356, 387 Sequeira, M. See Silva, M.A., 387 Sergeant, D.E., 360, 372, 387 Seurat, L.G., 256, 269, 272, 277 Sewell, A.T. See Duggins, D.O., 228; See Eckman, J.E., 228 Seymour, R.J., 190, 234 Sfriso, A., 189, 235 Shackley, S.E. See Llewellyn, P.J., 140, 143, 144, 154 Shaffer, J.A., 165, 177, 178, 224, 335 Shank, T.M. See Feldman, R.A., 349 Shao, See Kuo, S., 290, 306 Shaw, P.M., 287, 308 Sheader, A. See Moss, B., 233 Shelden, K.E.W., 334, 336, 352 Sheldon, J.W., 122, 157 Sheldrick, M.C. See Kuiken, T., 384 Sherr, E.B. See Fry, B., 326, 349 Shibata, Y., 126, 157 Shimizu, M. See Sano, M., 308 Shimmield, G.B. See Pederson, T.F., 42 Shirayama, Y. See Kitazato, H., 330, 333, 343, 350

Shull, D.H., 29, 43 Sibuet, M., 313, 329, 340, 352 Sicinski, J., 70, 87, 111; See Arnaud, P.M., 97; See Jazdzewski, K., 106 Siebensohn, E. See Coleman, C.O., 101 Siebert, U., 368, 387; See Benke, H., 361, 363, 365, 370, 380; See Lockyer, C., 385 Sieg, J., 64, 71, 111, 112 Siewert, A. See Weslawski, J.M., 159 Silber, G.K. See Perry, S.L., 352 Silesu, M. See Giuseppetti, G., 274 Silva, M.A., 369, 387 Silver, M.W. See Alldredge, A., 288, 289, 303 Simenstad, C.A. See Duggins, D.O., 151 Simmons, S.L. See Earll, R.C., 151 Simon, M., 286, 308 Simpson, K.W., 134, 157 Simpson, V.R. See Kuiken, T., 384 Sirenko, B.I., See Gutt, J., 104 Skaare, J.U. See Kleivane, L., 384 Slattery, M., 165, 173, 189, 202, 206, 235 Slatyer, R.O. See Connell, J.H., 127, 129, 150, 331, 332, 349 Slijper, E.J., 359, 387 Slim, F.J. See Hemminga, M.A., 152 Sloane, J.F. See Lilly, S.J., 231; See Round, F.E., 234 Slobodkin, L.B. See Lopez, G.R., 154 Sluka, R.D., 298, 308 Smallwood, B.J., 6, 18, 43 Smeenk, C., 356, 375, 376, 387, 388 Smirnov, I.S. See Gutt, J., 104 Smirnov, R.V., 71, 112 Smith, A. See Collins, M.A., 348; See Roberts, D.E., 234 Smith, B.D., 125, 157 Smith, C.R., 11, 18, 26, 28, 29, 30, 43, 311–354, 312, 313, 314, 315, 317, 318, 319, 320, 321, 322, 323, 324, 325, 327, 330, 331, 332, 333, 338, 339, 340, 342, 347, 352; See Allison, P.A., 348; See Baco, A.R., 329, 330, 348; See Bennett, B.A., 348; See Deming, J., 349; See Distel, D.L., 349; See Feldman, R.A., 349; See Kukert, H., 315, 350; See Levin, L.A., 40, 351; See Smallwood, B.J., 43; See Snelgrove, P.V.R., 50, 112; See Williams, A.B., 354 Smith, F., 165, 189, 194, 235; See Klinkhammer, G.P., 106 Smith, G.J.D., 360, 364, 366, 367, 370, 388 Smith, K.L. See Priede, I.G., 352; See Wakeﬁeld, W.W., 23, 44 Smith, L.P. See Little, C., 172, 177, 194, 231 Smith, N.D. See Boaventura, D., 227 Smith, R.J., 360, 361, 364, 365, 388 Smith, R.O. See Schroeter, S.C., 234 Smith, S.C. See Hooker, S.K., 383

412

AUTHOR INDEX

Smith, S.L., 11, 43, 44; See Morrison, J.M., 41 Smith Jr, W.O. See Hong, Y., 305 Snelgrove, P.V.R., 50, 53, 112, 315, 338, 353 Snoeijs, P. See Eriksson, B.K., 229; See Pedersén, M., 177, 233; See Johansson, G., 230 Snyder, L.M. See Willey, P., 332, 353 Snyder, S.W. See Renaud, P.E., 234 Soares, A.G., 137, 139, 140, 157 Sobarzo, M. See Gallardo, V.A., 37 Soeda H. See Hatakeyama, Y., 370, 382 Soltwedel, T. See Klages, M., 350 Somero, G.N. See Yang, T.H., 45 Sommer, S. See Pfannkuche, O., 42 Sonne 90 Scientiﬁc Party. See von Rad, U., 44 Sorokin, Y.I., 284, 289, 308 Soto, A. See Gutiérrez, D., 38; See Neira, C., 41 Sousa, W.P., 187, 190, 193, 196, 235 Soutar, A. See Phleger, P.B., 8, 9, 24, 25, 42 Southwell, T., 359, 388 Sparholt, H., 371, 388 Speel, J.A., 73, 112 Speer, K.G. See Van Dover, C.L., 353 Spencer, N. See Santos, M.B., 387 Spicer, J.I., 14, 44 Spiess, F., 7, 44 Squires, R.L., 314, 338, 343, 353; See Goedert, J.L., 349 Sreepada, R.A., 75, 112; See Jaya Sree, V., 105 Stahre, B. See Olin, R., 155 Stanek, A. See Weslawski, J.M., 159 Starmans, A. See Gutt, J., 93, 104; See Klages, M., 106 Statham, P.J. See German, C.R., 103 Stavridis, T. See Coleman, C.O., 101 Stebbins, T.D. See Zmarzly, D.L., 354 Steene, R.C. See Randall, J.E., 308 Stefani, J. See Bayer, F.M., 60, 98 Steiner, G., 68, 112 Steneck, R.S., 179, 190, 191, 235, 297, 308 Stenton-Dozey, J. See Grifﬁths, C.L., 117, 119, 126, 127, 128, 151 Stenton-Dozey, J.M.E., 118, 119, 123, 126, 127, 128, 131, 158 Stephen, A.C., 359, 388 Stephenson, T.A., 165, 177, 235 Steuer, A., 256, 277 Steven, A.D.L. See Meekan M.G., 307 Stewart, J.E. See Holmer, M., 230 Stewart, J.G., 165, 167, 168, 172, 177, 179, 183, 187, 188, 191, 235 Stiboy-Risch, C., 65, 112 Stiller, M., 70, 112; See Brey, T., 99 Stockton, W.L., 312, 315, 353 Stolz, J.F. See Rhoads, D.C., 42 Stolzenbach, K.D. See Butman, C.A., 227

Storey, G., 364, 371, 373, 388 Stork, N.E., 50, 112 Storlazzi, C.D., 165, 235 Stott, L.D., 5, 44 Strasburg, D.W. See Hiatt, R.W., 290, 305 Stuart, C.T. See Levin, L.A., 40; See Rex, M.A., 111 Stuart, V., 117, 158 Sturtivant, C.R., 370, 388; See Goodson, A.D., 370, 382 Subramanian, A.N., 376, 388 Suhr, S. See Gooday, A.J., 38 Sumida, P.Y.G. See Pirez, A.M.S., 64, 110 Summons, R.E. See Hinrichs, K.U., 39 Suzuki, Y. See Matsunaga, K., 232 Svärdson, G., 373, 388 Svoboda, A. See Peña Cantero, A.L., 109, 110 Swain, T. See Valiela, I., 158 Swanepoel, D. See Ryan, P.G., 143, 157 Swart, R. de L., 376, 388 Swartz, S.L. See Jones, M.L., 350 Sweet, M.H. See Whorff, J.S., 236 Sylva, S.P. See Hinrichs, K.U., 38, 39 Syster, D.A. See Renaud, P.E., 234 Syvitski, J.P.M. See Farrow, G.E., 229 Szmant, A.M., 288, 308 Szmant-Froelich, A. See Dodge, R.E., 162, 228 Tabak, M. See Reimers, C., 42 Taborsky, M. See Zoufal, R., 297, 309 Taggart, S.J. See Zabel, C.J., 122, 159 Takeuchi, I. See Numanami, H., 108 Tamburri, M.N., 336, 353 Tanabe, S., 377, 388; See Subramanian, A.N., 388 Tankersley, R.A. See Herrnkind, W.F., 230 Tarafa, M. See Emeis, K.-C., 37 Tararam, A. See Wakabara, Y., 113 Tarazona, J., 5, 22, 27, 28, 44; See Arntz, W.E., 34; See Rosenberg, R., 42 Tarr, J.G., 135, 158 Tarr, P.W. See Tarr, J.G., 135, 158 Tasker, M.L. See Northridge, S.P., 385 Tatsukawa, R. See Subramanian, A.N., 388; See Tanabe, S., 388 Taviani, M., 243, 277 Taylor, A.C. See Moore, P.G., 155 Taylor, B.E. See Carefoot, T.H., 149 Taylor, B.L., 364, 388 Taylor, J.D., 118, 158 Taylor, M.A. See Martill, D.M., 351 Taylor, M.K. See Andriashek, D., 148 Taylor, P.D. See Ostrovskii, A.N., 72, 109 Taylor, P.R., 170, 180, 182, 194, 222, 235 Teal, J.M. See Valiela, I., 158 Tebble, N., 312, 314, 325, 337, 353

413

AUTHOR INDEX

Tegner, M.J. See Seymour, R.J., 234 Teilmann, J., 362, 388; See Heide-Jørgensen, M.-P., 382; See Lowry, N., 385 Temnikow, N.K. See Lipps, J.H., 107 Templado, J. See Zenetos, A., 277 Tendal, O. See Barthel, D., 59, 98 Tendal, O.S. See Barthel, D., 98 Terawaki, T. See Yoshida, G., 236 Teske, A. See Fossing, H., 37; See Schulz, H.N., 43 Thaine, C. See Monaghan, P., 385 Thamdrup, B. See Fossing, H., 37 Thayer, G.W., 118, 158 Thiel, H., 2, 44 Thiermans F. See Dubilier, N., 37 Thingstad, F. See Azam, F., 303 Thollot, P. See Kulbicki, M., 306 Thom, R.M., 193, 235 Thomas, C.L. See Levin, L.A., 40 Thomas, P.A., 59, 60, 112 Thompson, B., 21, 28, 44 Thompson, J.B., 18, 19, 20, 21, 28, 29, 30, 44; See Mullins, H.T., 41 Thompson, P.M., 364, 372, 376, 388; See Pierce, G.J., 386; See Tollit, D.J., 372, 388 Thomson, J.A., 60, 112 Thornton, L., 145, 158 Thorpe, J.P. See Allcock, A.L., 68, 97 Thorpe, S.E. See Trathan, P.N., 112 Thorson, G., 270, 277 Thrush, S.F. See Turner, S.J., 235 Thurston, M.H., 62, 94, 95, 112; See Watling, L., 84, 113 Tibbetts, I.R. See Townsend, K.A., 298, 308 Tiedemann, R., 356, 388 Tiefenbacher, L., 65, 112 Tiemann, H. See Jarms, G., 60, 105 Tietjen, J. See Lambshead, P.J.D., 106 Tiezen, L.L., 359, 388 Tillier, L., 239, 255, 256, 277 Timmerman, H.H. See Ross, P.S., 387; See Swart, R. de L., 388 Tinker, S.W., 142, 158 Todaro, M.A., 16, 24, 44 Toggweiler J.R. See Sarmiento, J.L., 43 Tolley, K.A., 356, 388; See Westgate, A.J., 356, 389 Tollit, D.J., 372, 388; See Thompson, P.M., 388 Tom, M., 271, 277 Tomás, J., 144, 158 Tomilin, A.G., 356, 358, 359, 360, 364, 366, 389 Tomlin, J.R. le B., 239, 277 Tomlinson, K.W. See Prince, W.A., 156 Tomlinson, P.B., 118, 158 Torchia, G. See Relini, G., 234 Tortonese, E. See Whitehead, P.J.P., 389 Tørum, A. See Løvås, S.M., 140, 154

Townsend, C.R. See Begon, M., 303 Townsend, K.A., 298, 308 Towsend, C., 165, 187, 235 Trathan, P.N., 55, 56, 112 Travis, J.L., 9, 44 Tregenza, N. See Pierpoint, C., 386 Tregenza, N.J.C., 357, 377, 389 Tréguer, P., 53, 112 Tringali, L., 269, 277 Trites, A.W. See Pauly, D., 386 Trivedi, A. See Black, M.B., 348 Troncoso, J.S. See Arnaud, P.M., 97 Trowbridge, C.D., 164, 165, 167, 168, 171, 173, 180, 184, 186, 191, 192, 201, 224, 235 Tshudy, D.M. See Feldmann, R.M., 65, 91, 102 Tsipoura, N., 123, 158 Tsuchiya, M., 196, 235 Tsuchiya, Y. See Numanami, H., 108 Tsukada, D. See Thompson, B., 44 Tudor, D.T. See Earll, R.C., 151; See Williams, A.T., 145, 159 Tunnicliffe, V., 329, 338, 340, 343, 353; See Farrow, G.E., 229 Turner, R. See Lutz, R., 351 Turner, R.D., 317, 337, 353 Turner, R.E. See Rabalais, N.N., 2, 42 Turner, S.J., 166, 193, 235 Turner, T., 170, 180, 181, 235 Turpaeva, E.P., 65, 112 Turpaeva, Y.P., 65, 112 Tyler, P. See Creasey, S., 36 Tyler, P.A., 347, 353; See Gage, J.D., 315, 349; See German, C.R., 103 Tyson, R.V., 5, 44 Ulloa, O. See Fossing, H., 37; See Gallardo, V.A., 37 Umar, M.J., 185, 191, 207, 213, 217, 235 Underwood, A.J., 173, 190, 235 Uneputty, P.A., 145, 158 United Nations Environmental Programme, 162, 236 Urdangarin, I.I. See Saiz-Salinas, J.I., 177, 178, 186, 193, 194, 234 Urgorri, V. See García-Alvarez, O., 103 Uri, J.S. See Vermaat, J.E., 236 Urkiaga-Alberdi, J. See Saiz-Salinas, J.I., 166, 173, 177, 193, 194, 234 Utinomi, H., 65, 112 Uttley, J.D. See Monaghan, P., 385 Vadas, R.L., 189, 236 Vaillant, E., 238, 277 Valdivia E. See Arntz, W.E., 35 Valentine, J.W. See Roy, K., 111 Valeur, J. See Lund-Hansen, L.C., 231

414

AUTHOR INDEX

Valiela, I., 123, 129, 132, 140, 158; See Alber, M., 281, 288, 303; See D’Avanzo, C.D., 281, 305 Van Beneden, P.J., 359, 389 Van Bree, P.J.H., 375, 389 van Breugel, P. See Moodley, L., 41 van der Horst, G. See McLachlan, A., 155 van der Land, J., 62, 112 Van der Velde, G. See van Katwijk, M.M., 236 van der Zwaan, G.J., 15, 44; See den Dulk, M., 36; See Moodley, L., 41 van Dijken, G.L. See El Sayed, S., 257, 274 Van Dover, C.L., 329, 333, 345, 346, 353; See Tyler, P.A., 353 van Hove, E.M. See van Katwijk, M.M., 236 van Katwijk, M.M., 162, 166, 236 van Loon, R. See van Katwijk, M.M., 236 van Loveren, H. See Ross, P.S., 387 van Mol, J.J. See Arnaud, P.M., 97 van Pelt, T.I., 122, 158 van Rijswijk, P. See Alkemade, R., 124, 125, 147 van Utrech, W.L. 357, 377, 389 van Vierssen, W. See Vermaat, J.E., 236 Vannini, M. See Messana, G., 155 Vardala-Theodorou, E. See Delamotte, M., 271, 273 Vargas, M.A., 286, 308 Vásquez, C. See Gutiérrez, D., 38 Vásquez, J. See Santelices, B., 234 Vaughan, N. See Kastelein, R.A., 383 Vazquez, E. See Young, C.M., 45 Vedder, L.J. See Swart, R. de L., 388 Velander, K.A., 145, 158 Velásquez, C.R. See Pulfrich, A., 234 Velimirov, B., 117, 158; See Field, J.G., 151 Venables, B.J., 131, 158 Verboom, W.C. See Kastelein, R.A., 383 Verburg, P. See Bruggemann, J.H., 304 Vercoutere, T. See Thompson, J.B., 44 Vercoutere, T.L. See Mullins, H.T., 41 Verdugo, P. See Chin, W., 304 Verhecken, A., 258, 277 Verlaque, M. See Birje, J., 227 Vermaat, J.E., 162, 236 Vermeij, G.J., 238, 239, 270, 277 Verschuure, J.M. See Kamermans, P., 230 Vervoort, W. See Peña Cantero, A.L., 60, 109, 110 Verwey, J., 375, 389 Vetter, E.W., 347, 353 Vetter, R.D., 10, 44; See Cary, S.C., 36; See Jacobson, L.D., 23, 24, 39 Víkingsson, G.A. See Tolley, K.A., 388 Villa, R. See Tringali, L., 269, 277 Vincx, M. See Heip, C., 152 Vinogradova, N.G., 70, 112

Vinther, M., 357, 389 Vio, E. See De Min, R., 239, 273 Virgilio, M. See Airoldi, L., 226 Virnstein, R.W. See Fry, B., 118, 151 Viso, A.C., 285, 308 Vogt, H., 165, 176, 190, 193, 236 Vogt, H.P. See White, A.T., 236 Vokes, E.H. See Houart, R., 254, 275 Volkel, S. See Grieshaber, M.K., 13, 38 Volkman, J.K., 285, 308 Volse, L.A. See Devinny, J.S., 170, 194, 196, 198, 205, 206, 207, 220, 228 von Rad, U., 5, 44; See Cowie, G.L., 36; See Schmaljohann, R., 43 von Salvini-Plawen, L., 68, 113; See GarcíaAlvarez, O., 103 Vos, J.G., 376, 389; See Ross, P.S., 387; See Swart, R. de L., 388 Voss, J., 67, 73, 113; See Piepenburg, D., 110 Vranken, G. See Heip, C., 152 Vrijenhoek, R.C., 333, 353; See Baco, A.R., 348; See Black, M.B., 348; See Craddock, C., 349; See Feldman, R.A., 349; See Peek, A.S., 352; See Van Dover, C.L., 353 Wada, E. See Minigawa, M., 326, 328, 351 Wada, H., 314, 325, 353; See Fujioka, K., 349; See Naganuma, R., 351 Wägele, J.W., 64, 113; See Sieg, J., 71, 112; See Brey, T., 99 Wagner, K. See Coleman, C.O., 101 Wagner, M. See Dubilier, N., 37 Wakabara, Y., 62, 113 Wakeﬁeld, W.W., 23, 44; See Adams, P.B., 34 Walker, D.I., 117, 159, 163, 236 Walker, T.R., 145, 159 Wallace, J.R. See Merritt, R.W., 332, 351 Walsh, J.J., 53, 113 Walter, S. See Schmaljohann, R., 43 Walters, S.M., 92, 113 Walton, M. See Tolley, K.A., 388 Walton, M.J., 356, 389; See Kuiken, T., 384 Wang, J.Y., 356, 389; See Rosel, P.E., 387 Wanless, S. See Harris, M.P., 364, 373, 382 Wantiez, L. See Kulbicki, M., 306 Ward, B.B., 3, 31, 44; See Karl, D.M., 39 Warén, A., 67, 113, 313, 325, 337, 338, 340, 341, 343, 353 Waring, G.T., 367, 389 Warwick, R.M., 49, 113; See Clarke, K.R., 49, 101 Wassenberg, T.J. See Moriarty, D.J.W., 307 Wassermann, D. See Wassermann, M., 389 Wassermann, M., 376, 389 Wassman, P. See Passow, U., 288, 289, 307 Wasti, S.S. See Abell, D.H., 147

415

AUTHOR INDEX

Watanabe, S. See Tanabe, S., 388 Watling, L., 84, 113 Watson, A.P., 369, 389; See Gaskin, D.E., 364, 382 Watson, L., 356, 389 Way, L.S. See Davidson, N.S., 150 Webb, A. See Northridge, S.P., 385 Webb, C.O., 49, 113 Weeks, S.J., 28, 44 Weinstein, M.P. See Iannuzzi, T.J., 230 Weisberg, S.B. See Moore, S.L., 155 Wells, D.E., 377, 389 Wells, F.E., 138, 159 Wensvoort, P. See Boon, J.P., 380 Weslawski, J.M., 144, 159; See Jazdzewski, K., 106 West, F.J.C. See Meadows, A., 41; See Murray, J.M.H., 41 Westgate, A.J., 356, 369, 379, 389; See Palka, D.L., 385; See Read, A.J., 364, 386 Weston, D.P., 315, 322, 353 Wetzel, R.G. See Otsuki, A., 288, 307 Wheatcroft, R.A., 32, 44; See Smith, C.R., 352 Whelan, J.K. See Emeis, K.-C., 37 White, A. See Hong, Y., 305 White, A.T., 162, 236 White, B.N., 32, 44; See Wang, J.Y., 389 White, K.M. See O’Donoghue, C.H., 276 Whitehead, H. See Katona, S., 333, 350 Whitehead, P.J.P., 371, 374, 389 Whittaker, R.H., 49, 113 Whitworth, T. See Orsi, A.H., 109 Whorff, J.S., 167, 172, 188, 236 Whorff, L.L. See Whorff, J.S., 236 Whymper, F., 356, 390 Widdowson, T.B. See Thom, R.M., 193, 235 Widmark, J.G.V. See Jorissen, F.J., 39 Wiencke, C., 58, 113; See Clayton, M.N., 101 Wiepkema, P.R. See Read, A.J., 386 Wijnsma, G., 374, 390 Wilcockson, R. See Phillips, D.S., 156 Wilcockson, R.W. See Leggett, M.C., 154 Wild C. See Fabricius, K.E., 305 Wildish, D.J. See Holmer, M., 230 Wilke, F., 360, 370, 390 Wilkins, S.DeC. See Nelson, S.G., 295, 307 Wilkinson, C.R., 286, 308 Wilkinson, M.R. See Turner, S.J., 235 Willan, R.C., 261, 267, 277 Willey, P., 332, 353 Williams, A.B., 341, 353 Williams, A.T., 145, 159; See Earll, R.C., 151 Williams, D.McB. See Hamner, W.M., 305 Williams, J.M. See Northridge, S.P., 385 Williams, R.B., 60, 113 Williams, S. See Hyland, J., 39

Williams, S.L., 181, 186, 236; See Carpenter, R.C., 187, 227; See Thayer, G.W., 158 Williamson, M., 267, 277 Williamson, R.B. See Turner, S.J., 235 Willis, J.C., 59, 92, 113 Willoughby, N.G., 145, 159 Wilson, B. See Ross, H.M., 357, 368, 373, 387 Wilson, C. See Chin, C.S., 100; See Klinkhammer, G.P., 106 Wilson, D.S., 129, 159 Wilson, E.O., 48, 50, 51, 113; See MacArthur, R.H., 48, 107 Wilson, G.D.F. See Poore, G.C.B., 50, 90, 110; See Rex, M.A., 111 Wilson, M.R. See Claridge, M.F., 100 Wilson, M.T. See Feldmann, R.M., 65, 102 Wilson, S.K., 279–309, 280, 281, 282, 284, 285, 286, 287, 288, 292, 294, 295, 296, 298, 299, 300, 308, 309 Winslade, P., 364, 371, 390 Winston, J.E., 72, 78, 113 Wishner, K., 45; See Levin, L.A., 40; See Morrison, J.M., 41; See Smith, S.L., 44 Wishner, K.F., 4, 11, 14, 18, 20, 21, 23, 24, 31, 44, 45; See Gowing, M.M., 20, 31, 38 Witman, J.D. See Smith, F., 165, 189, 194, 235 Woessner, J.W. See Neushul, M., 233 Wolanski, E. See Fabricius, K.E., 305 Wolf, C. See Gray, L.E., 382 Wolff, G.A. See Gage, J.D., 37; See Smallwood, B.J., 6, 43 Wolff, T., 337, 347, 354 Wolff, W.J. See Reise, K., 277 Wolman, A. See Johnson, J., 336, 350 Wolman, A.A. See Rice, D.W., 352 Wood, D.M. See Murray, J.M.H., 41 Wooldridge, T. See McLachlan, A., 155 Woolfe, K.J. See Larcombe, P., 162, 231 World Conservation Monitoring Centre, 48, 113 Wotton, R.S., 163, 236 Wright, P.J., 364, 390 Wright, R.F. See Carpenter, S.R., 227 Wu, B., 62, 113 Wu, R.S. See Gray, J.S., 38 Wulf, J. See Heide-Jørgensen, M.-P., 382 Wyrtki, K., 2, 3, 4, 7, 45 Yamada, S.B. See Menge, B.A., 232 Yang, T.H., 12, 45 Yasui, W.C., 372, 390 Yayanos, A.A. See Hessler, R.R., 350 Yingst, J.Y. See Rhoads, D.C., 42, 352 Yoon, C.K., 77, 113 Yoshida, G., 191, 236 Yoshikawa, K. See Yoshida, G., 236 Young, A.M. See Cullen, S.J., 150

416

AUTHOR INDEX

Young, C. See Creasey, S., 36 Young, C.M., 10, 24, 45; See Marsh, A.G., 351 Young, D.R. See Rau, G.H., 352 Young, P.S., 65, 113 Yuen, W.Y. See Morton, B., 122, 123, 155 Yule, G.U. See Willis, J.C., 92, 113 Yumamoto, S. See Gaskin, D.E., 382 Yurick, D.B., 356, 390 Zabel, C.J., 122, 159 Zachariasse, W.J. See den Dulk, M., 36; See Reichart, G.L., 42 Zahn, R. See Baas, J.H., 35 Zaitsev, Y., 267, 277 Zeballos, J. See Arntz, W.E., 35 Zedler, J.B. See Emerson, S.E., 172, 177, 187, 191, 229 Zeidler, W., 62, 113 Zeinstra, T. See Duinker, J.C., 381 Zeitzschel, B. See Owen, R.W., 20, 42

Zellermayer, L. See Wassermann, M., 389 Zenetos, A., 237–277, 239, 257, 277; See Galil, B., 266, 274 Zentara, S.J. See Kamykowski, D., 2, 3, 5, 39 Zertuche-González, J.A. See Barilotti, D.C., 142, 148; See Pacheco-Ruíz, I., 141, 142, 156 Zevina, G.B., 65, 114 Zhao, J. See Wu, B., 62, 113 Zhender, A., 14, 45 Zibrowius, H., 239, 277 Ziegler, A. See Pennings, S.C., 156 Zieman, J.C. See Thayer, G.W., 158 Zimmer, M. See Pennings, S.C., 156 Zimmerman, A. See Brey, T., 99 Zmarzly, D.L., 315, 322, 354 Zoli, C., 142, 159 Zoufal, R., 297, 309 Zoutendyk, P. See Field, J.G., 151; See Velimirov, B., 158 Zwalley, H.J. See Gloersen, P., 103

417

SYSTEMATIC INDEX

References to complete articles are given in bold type; references to sections of articles are given in italics; references to pages are given in normal type. Abramis brama, 360 Acanthocardia tuberculata, 269 Acanthocephala, 51 Acanthuridae, 291, 292, 293, 297, 299, 300 Acanthurus bahianus, 291 chirurgus, 291 coeruleus, 291 lineatus, 282, 290 nigricauda, 291, 292 olivaceus, 291, 292 tennenti, 291 Acar plicata, 248 Acarina, 78 Achiropsettidae, 76 Acrochaetium, 187 Acteocina mucronata, 246, 258 Actinopteryx fucicola, 128 Actitis macularia, 136 Adalactaeon amoenus, 246, 269 fulvus, 246 Adipicola, 314, 338, 343 (Idas) arcuatilis, 341 osseocola, 337, 341 pelagica, 312, 341 Adula (Adipicola) simpsoni, 312 Aeolidiella indica, 248 Afrocardium richardi, 250 Aglaja taila, 241 Agnatha, 76, 78 Ahnfeltia concinna, 182 plicata, 176, 182 Ahnfeltiopsis linearis, 182, 186, 196, 197, 202, 205 Alca torda, 373 Alcyonaria, 60, 61 Alcyonidium, 72 Alcyonium paessleri, 189, 202, 203 Aldrichetta forsteri, 138 Allorchestes compressa, 127, 132, 138 Alopex lagopus, 122 Alvania consociella, 254 discors, 254 dorbignyi, 240, 244, 251, 254 lanciae, 254 Amblygobius nocturnis, 292 Amblyrhynchus cristatus, 138 Ammodytes, 360 marinus, 371, 374 tobianus, 371

Ammodytidae, 355, 360 Ammothea, 66 Ammotheidae, 66 Ampelisca, 10, 23 Ampharetidae, 71, 341 Amphibolis, 118 Amphicteis, 23 Amphipod sp. D, 327 Amphipoda, 62, 63, 78, 87, 126 Amphiura, 73 Amphiuridae, 74 Amygdalum agglutinans, 11, 265 anoxicolum, 11, 21, 23, 24, 26 Anachis selasphora, 269 Anadara corbuloides, 265 demiri, 248 inaequivalvis, 237, 248, 261, 267, 268, 272 natalensis, 248, 256 notabilis, 241 Anaplopoma ﬁmbria, 23 Anarthruridae, 63, 64 Angiola punctostriata, 244 Angiospermae, 117 Anguilla anguilla, 359 Annelida, 21, 52, 69, 78, 87, 341 Anobothrus sp. nov., 341 Anomia ephippium, 214 Anomura, 341 Anopla, 67 Anoplopoma ﬁmbria, 337 Antedonidae, 74 Anthomedusae, 60 Anthopleura elegantissima, 176, 182, 200 Anthozoa, 59, 60, 61, 78, 87 Antigona lamellaris, 250, 258 Antrops truncipennis, 128, 133 Aphelochaeta, 26 Aphya minuta, 360 Aplacophora, 68, 341 Aplidium, 75 Aplysia juliana, 241 Apterosessinia peringueyi, 120 Arachnida, 78 Arachnopusidae, 72 Archaea, 7, 49 Archaegastropoda, 341 Arcidae, 269 Arctinula groenlandica, 241 Arcturidae, 63, 64

419

SYSTEMATIC INDEX

Brachidontes pharaonis, 237, 248, 255, 256, 257 Brachiopoda, 52, 71, 78, 87 Brachyodontes pharaonis, 262, 267 Branta bernicla, 136 Brissopsis paciﬁca, 18, 19 Bruciella, 340 laevigata, 341 pruinosa, 341 Bryozoa, 52, 72, 73, 78, 87 Buccinidae, 69 Buddenbrockia plumatellae, 52 Bugulidae, 72 Bulimina, 20 Bulla ampulla, 246 Bullia miran, 264 Burnupena, 192 Bursa marginata, 240 Bursatella leachi, 237, 246, 256, 258, 259, 266, 270

Arenaria interpres, 123, 135, 136 melanocephala, 136 Argentina, 363 Argopecten purpuratus, 28 Aricidea, 23 Arripis georgianus, 138 Artedidraconidae, 76 Arthropoda, 52, 78, 340, 341 Asabellides sp. nov., 341 Ascidiacea, 75 Ascophyllum nodosum, 176 Aspella anceps, 254 Asteriidae, 74 Asteroidea, 74, 78, 87 Asterropteryx semipunctatus, 292 Astyris (Mitrella) permodesta, 24, 328 permodesta, 21, 324, 339 Atactodea glabrata, 250 Atherina pontica, 360 Atrosalarias fuscus, 291, 292, 296 Atys blainvilliana, 241 Audouinella purpurea, 187 Axinodon sp. nov., 341 Balaena glacialis, 334 Balaenoptera acutorostrata, 334, 373 borealis, 334 edeni, 334 musculus, 334 mysticetus, 334 physalis, 334 Balanus balanoides, 176 Bathydraconidae, 76 Bathygobius fuscus, 292 Bathykurila guaymasensis, 339 Bathymodiolinae, 341 Bathymodiolus, 343 Bathysiphon, 20 Batis maritima, 139 Beggiatoa, 1, 7, 28, 330 Benthescymus altus, 18, 19 Benthomangelia antonia, 14 Benthomodiolus, 343 Berthellina citrina, 241 Birgus latro, 138 Bittium proteum, 240 Bivalvia, 69, 78, 87, 241, 248, 265, 338, 341, 342 Blattodea, 129 Bledius, 136 Blenniidae, 291, 292, 293, 298, 299, 300 Blennius cristatus, 291 marmoreus, 291 Bolivina, 20 seminuda, 25 Bonellidae, 69 Bovichtyidae, 76

Cabereidae, 72 Caﬁus ragazii, 128 xantholoma, 135 Calcarea, 58, 59 Calidris alba, 123, 130 canutus, 123 pusilla, 123 Callicnemis latreillei, 120 Callinectes arcuatus, 28 Callipallenidae, 66 Calloporidae, 72 Callostracum gracile, 240 Caloria indica, 248 Calyptogena, 340 elongata, 339, 343 kilmeri, 339, 343 paciﬁca, 339 Cancellaria cancellata, 265, 266 Cancellaridae, 69 Canis latrans, 121 mesomeles, 122 Caprellidae, 62 Carabidae, 120, 126, 128, 129 Carapidae, 76 Cardiidae, 269 Carenioidae, 62 Caspialosa, 360 Cassidulina, 20 Catharacta skua, 373 Caulerpa, 182, 186 taxifolia, 260 Cellana rota, 244 Cellariidae, 72 Cellarinella, 72 Celleporidae, 72 Centropyge argi, 291

420

SYSTEMATIC INDEX

ferrugatus, 291 Centroscyllium coelolepis, 337 Cephalochordata, 52 Cephalopoda, 78 Cephalorhynchus heavisidii, 369 Ceramium, 187 Cerithiidae, 69, 269 Cerithioidea, 269 Cerithiopsis pulvis, 244 tenthrenois, 244 Cerithium caeruleum, 240 echinatum, 240 egenum, 244 erythraeoense, 240, 256 nesioticum, 244 scabridum, 237, 244, 255, 256, 257, 258, 262, 266, 270 Chaetognatha, 51 Chaetomorpha linum, 176 Chama paciﬁca, 248, 257 Chamelea gallina, 261 Channichthyidae, 76 Chaperidae, 72 Charadrius alexandrinus, 136 rubicollis, 135 semipalmatus, 136 Cheilopora praelonga, 210 Cheilostomatida, 72, 87 Chelicerata, 52, 65, 78 Chelidonura fulvipunctata, 246, 258 Chilomycterus schoepﬁ, 130 Chiton hululensis, 244, 256 Chlamys lischkei, 248, 272 Chlorophyta, 58 Chlorurus microrhinos, 292, 296, 297 sordidus, 292, 296, 297 Chondracanthus acicularis, 176 canaliculatus, 187 Chondrichthyes, 76, 78 Chondrus crispus, 186 Chordata, 52, 75, 78 Chromodoris clenchi, 241 quadricolor, 248, 255, 259, 260, 270, 272 Chrysallida ﬁscheri, 246 maiae, 246, 256, 257 pirintella, 246 Chrysopetalidae, 341 Chrysophyceae, 58 Chthamalus ﬁssus/dalli, 176 Cidaridae, 74 Cingulina isseli, 246 Circenita callipyga, 250 Cirripectes chelomatus, 292 Cirripedia, 63, 65, 78 Cirrophorus lyra, 15, 27 Cladophora, 187

columbiana, 176 rupestris, 198 Clanculus kraussi, 264 Clarenciidae, 62 Clathrofenella ferruginea, 244 Clavagella aperta, 265 Clelandella infucata, 240 Clementia papyracea, 250 Clitellata, 69, 78 Clupea harengus, 355, 359 pallasii, 360 Clypeomorus bifasciatus, 244 Cnidaria, 51, 60, 61, 78, 87 Cnidoglanis macrocephalus, 138 Cocculina, 337 craigsmithi, 323, 324, 325, 328, 339, 341, 342 Cocculiniformia, 337 Cocos nucifera, 121 Codium setchellii, 184, 187, 192, 196, 202 Coelopa, 128, 134 frigida, 134, 135 pilipes, 134, 135 vanduzeei, 134 Coleoptera, 50, 126, 127, 128, 129 Collembola, 126 Colossendeidae, 66 Colossendeis, 66 Congiopodidae, 76 Conus arenatus, 241, 256 fumigatus, 246, 255 Coragyps atratus, 129 Corallina, 184, 187, 188 pinnatifolia, 184 vancouveriensis, 184 Coralliobia madreporarum, 240 Corophiidae, 63 Corvus corax, 122 Coryphopterus glaucofraenum, 291 Cossura, 23, 27 chilensis, 23 Crania, 71 Crassatelloidea, 11 Crassostrea gigas, 248, 260, 261, 266, 268, 271 virginica, 241, 272 Crepidula aculeata, 244, 252, 261 fornicata, 244, 262, 268 Crinoidea, 74, 78, 87 Crustacea, 52, 62, 78, 87, 342 Cryptoarachnidium argilla, 187 Cryptosiphonia woodii, 199 Ctenochaetus striatus, 291, 292, 295, 296, 297 strigosus, 291 Ctenophora, 51, 61, 78 Ctenostomatida, 72 Cubozoa, 60

421

SYSTEMATIC INDEX

Cucumariidae, 74 Cumella, 338 Cumella sp. A, 322 Cuna gambiensis, 265 Curculionidae, 128 Cuthona perca, 248 Cycliophora, 52 Cycloscala hyalina, 244, 258 Cyclostomatida, 72 Cyclostrematidae, 69 Cylichna cf. mongii, 241 Cylichnina girardi, 246 Cymbiola vespertilio, 243 Cymbium cucumis, 265, 266 olla, 265 rubiginosum, 241 tritonis, 265 Cymodocea, 118, 125, 142 Cymodoceaceae, 117 Cypraea pantherina, 240 tigris, 243 Cypraeidae, 143, 269 Cyprinodon variegatus, 281 Decapoda, 65, 78, 87 Delphinus phocoena, 355 Demospongiae, 58, 59 Demoulia obtusata, 264 Dendrodoris fumata, 248 Dendropoma petraeum, 264 Dendrostrea frons, 248, 258 Dermochelys coriacea, 129 Desmodora masira, 30 Desulfonema, 7 Diala varia, 244, 256 Diaphanidae, 69 Diaphorosoma, 10, 23 Dictyota, 183 dichotoma, 184 Dictyotales, 184 Dicyemida, 51 Digitaria digitaria, 265 Diodora ruppellii, 244 Diogenes edwardsii, 123 Diplodonta brocchii, 265 cf. subrotunda, 248 Diptera, 50, 126, 129 Dischistodus perspicillatus, 282, 292, 296 Discodoris lilacina, 246 Divalinga arabica, 248 Dolabrifera holboelli, 241 Donax serra, 137, 139 Dorvilleid sp. D, 339 Dorvilleidae, 10, 341 Dosinia erythraea, 250, 257 Durvillaea potatorum, 141

antarctica, 133 Eastonia rugosa, 265 Eatoniella, 68 Eatoniellidae, 69 Echinodermata, 52, 73, 78, 87 Echinogammarus obtusatus, 132 pirloti, 132 Echinoidea, 74, 78, 87 Echiura, 52, 69, 78 Echiuridae, 69 Ecklonia, 117 cava, 196, 201, 206, 220 maxima, 123, 132 radiata, 138 Ecsenius bicolor, 292 mandibularis, 292 stictus, 292 Ectoprocta, 52 Elpidiidae, 74 Encephaloides armstrongi, 18, 19, 26, 32 Engraulis encrasicholus, 360 Enopla, 67 Enteromorpha, 136, 142, 176, 178, 187 compressa, 194 intestinalis, 169, 176 Enteropneusta, 52 Entomacrodus nigricans, 291 Entoproct sp. B, 339 Entoprocta, 52, 78 Epidromus gladiolus, 254 Epimeriidae, 63 Epistominella, 20 bradyana, 25 Epitonium jolyi, 264 Epsilonematidae, 10, 24 Eptatretus deani, 318, 322, 338 Ergalatax obscura, 246, 258 Erosaria spurca, 264 turdus, 244 Erronea caurica, 240 Escarpia spicata, 339 Eschrichtius robustus, 334 Eubacteria, 49 Eucheuma uncinatum, 142 Eulimella lomana, 324, 339 Eulimidae, 69 Eusiridae, 63 Exochellidae, 72 Feldmannia, 215 Finella pupoides, 244, 257, 271 Fissurella nubecula, 264 Flabelligeridae, 71 Flabellina rubrolineata, 248 Flustridae, 72

422

SYSTEMATIC INDEX

Foraminifera, 2, 5, 8, 9, 11, 12, 14, 15, 20, 24, 28, 30 Formicidae, 130 Fratercula arctica, 373 Fucales, 117 Fucellia capensis, 128 costalis, 134 ruﬁtibia, 134 Fucus, 134, 190, 203 mytili, 185 serratus, 204, 206, 207 vesiculosus, 131, 186 vesiculosus (forma mytili), 186, 203 Fulvia australis, 250 fragilis, 250, 257, 258, 259, 270 Fursenkoina, 20 Fusinus verrucosus, 246, 257 Fusiturris similis, 265 Gadiculus argenteus thori, 363 Gadidae, 76 Gadiformes, 76 Gadus euxinus, 360 morhua, 359 morhua callarias, 360 Gafrarium pectinatum, 250, 257 Galeomma polita, 241 Gammarellidae, 62 Gammarus locusta, 126 Gari intermedia, 265 pseudoweinkaufﬁ, 265 Gastrochaena cymbium, 250 Gastroclonium coulteri, 187 Gastropoda, 69, 78, 87, 143, 240, 241, 244, 246, 248, 264, 265, 341, 342 Gastrotricha, 51, 78 Gavia immer, 130 Gelidium, 178, 187 corneum, 176, 190 sesquipedale, 176, 190 Gibberula caelata, 265 oryza, 264 secreta, 264 Gibborissoa virgata, 244 Gibbula, 272 albida, 239, 243, 244, 261 cineraria, 240 Giffordia, 187 Gigartina canaliculata, 187 papillata, 199 pectinata, 142 stellata, 186 Glochinema bathyperuvensis, 9, 10, 11, 24 Glycymeris arabica, 248 Glyptoparus delicatulus, 292, 293

Gnatholepis thompsoni, 291 Gnathophausia ingens, 11 Gnathostomula, 51 Gobiidae, 291, 292, 293 Gobius melanostomus, 360 rotan, 360 syrman, 360 Gobobulimina, 20 Golﬁngia, 69 Golﬁngiidae, 69 Gracilaria confervoides, 181 gracilis, 181 lemaneiformis, 183, 190, 199 Gracilariopsis lemaneiformis, 183, 190, 196, 199 Gryphaea angulata, 260 Gulo gulo, 122 Gymnammodytes semisquamatus, 371 Gymnogongrus linearis, 182, 186, 198, 202 platyphyllus, 186, 198 Gymnolaemata, 72 Hadromerida, 59 Halacritus algarum, 128 Haleciidae, 61 Haliaeetus leucocephalus, 122 Halimeda, 181, 186 tuna, 184 Haliotis cracherodii, 176 pustulata cruenta, 244 Haliotis roei, 138 rufescens, 202, 205 Halodule, 188 Halopitys incurvus, 176 Haminaea callidegenita, 246, 272 cyanomarginata, 246, 255 Haplosclerida, 59 Harmathoe, 70 craigsmithi, 339, 341 Harpagiferidae, 76 Hedophyllum sessile, 190 Hemichordata, 52, 78 Hemiglyphidodon plagiometopon, 282, 292 Heterobranchia, 246 Heterocarpus nesisi, 18, 19 Heterocyemida, 51 Heterokontophyta, 58 Heteroscelus incanus, 136 Heterozostera, 118 Hexactinellida, 58, 59 Hiatula rueppelliana, 250, 257 Himanthalia elongata, 196, 198 Hinemoa cylindrica, 246 Hippopus hippopus, 241, 243, 256 Histeridae, 128 Hochstetteria munieri, 241 Hoeglundina, 20

423

SYSTEMATIC INDEX

Holothuroidea, 74, 78, 87 Homoptera, 50 Hyaena brunnea, 122 Hyalogyrina, 340 Hyalogyrina n. sp., 341 Hyalogyrinidae, 340 Hyalopomatus merenzelleri, 19 Hydrophilidae, 128 Hydrozoa, 59, 60, 61, 87 Hymenoptera, 50 Hyperoplus immaculatus, 371 lanceolatus, 371 Hypselodoris infucata, 248 Idas, 314, 337, 343 (Adipicola) simpsoni, 337 argenteus, 337 ghisottii, 337, 341 pelagica, 341 washingtonia, 313, 314, 323, 324, 325, 326, 327, 329, 330, 331, 333, 337, 339, 342 Idasola, 314 Ilyarachna profunda, 323, 324, 325, 327, 328, 339, 342 Indocrassatella indica, 11 Iolaea neofelixoides, 246 Iphimediidae, 63 Isochyroceridae, 63 Isopoda, 63, 64, 78, 87 Istigobius decoratus, 292 goldmanni, 292 Janiridae, 339 Kamptozoa, 52 Katharina tunicata, 192 Kinorhyncha, 51, 78 Kyphosidae, 299, 300 Laevicardium ﬂavum, 241 Lagenorhynchus obscurus, 368 Lamellaridae, 69 Laminaria, 117 digitata, 132, 139 hyperborea, 140, 141 longipes, 180 saccharina, 196, 197, 199, 201 sinclairii, 180, 181 Laminariales, 117 Larus argentatus, 130 atricilla, 130 glaucescens, 122 marinus, 130 novaehollandiae, 135 Laternula anatina, 250, 257

Latirus polygonus, 240 Laurencia, 187, 200 paciﬁca, 187 Lepidomeniidae, 23 Lepidoptera, 50 Leptogorgia virgulata, 208, 209, 217 Leptomedusae, 60 Leucotina cf. eva, 246 Ligia dilatata, 132 pallasii, 131 Limanda limanda, 360 Limnomedusae, 60 Limopsis multistriata, 248 Limulus polyphemus, 123 Linga aurantia, 241 Liothyrella, 71 Liparidae, 76 Lithothamnion, 214 Littorina littorea, 169, 177, 240, 243, 272 obtusata, 240 punctata, 264 saxatilis, 253 Littorinidae, 69 Littorophiloscia tropicalis, 128 Lobatocerebrum, 52 Loligo forbesi, 363 opalescens, 365 pealii, 360 Lophiotoma indica, 241, 256 Loricifera, 51 Lottia gigantea, 176 Lucicutia, 11 Lucinoidea, 11 Lucinoma, 11 aequizonata, 30 Lucioperca lucioperca, 360 Lumbrineridae, 10 Lutra canadensis, 122 Lyrocteis ﬂavopallidus, 61, 79 Lysianassoidae, 63 Lysippe, 23 Macarorchestia remyi, 120 Macrocystis pyrifera, 117, 119, 125, 190, 196, 198, 205, 206, 207, 220 Macrouridae, 23 Mactra lilacea, 250 olorina, 250, 255, 256, 257 Mactrinula tryphera, 241 Magelona phyllisae, 27 Majidae, 19 Malacostraca, 62, 63, 64, 65, 78 Maldanid sp. C, 339 Maldanidae, 71 Malleus regulus, 257 Mallotus villosus, 359

424

SYSTEMATIC INDEX

Malvufundus regulus, 248, 256, 258 Marginella glabella, 263, 264 monilis, 143 Mastocarpus papillatus, 199 stellatus, 186 Maurolicus weitzmani, 364 Mazatlania cosentini, 240, 243 Medusozoa, 60, 61, 78 Megalorchestia, 127 Megalorchestia californiana, 131 Megaptera novaeangliae, 334, 373 Melanochlamys seurati, 241 Meleagrina margaritifera, 260 savignyi, 256 Melibe ﬁmbriata, 248, 255, 258, 272 Membranipora membranacea, 209, 210 Mercenaria mercenaria, 250, 262, 266, 268, 271 Merlangius merlangus 355, 359 Merluccius bilinearis, 360 gayi peruanus, 28 Mesalia brevialis, 264 opalina, 240, 266 Mesites aquitanus, 120 Mesogobius batrachocephalus, 360 Mesophyllum lichenoides, 176 Mesozoa, 51 Metaxia bacillum, 244 Metazoa, 346 Microcladia borealis, 199 Micromesistius potassou, 360 Microporellidae, 72 Microspathodon chrysurus, 291 Microstomus kitt, 372 paciﬁcus, 23 Mimosaceae, 120 Minolia nedyma, 271 Minuspio, 26 Mitrella broderipi, 264 bruggeni, 264 permodesta, 24, 324 Mixine circifrons, 318, 338 Modiolus auriculatus, 248, 256 lulat, 265 stultorum, 265 Molgulidae, 75 Mollusca, 52, 67, 78, 87, 237, 239, 240, 253, 262, 263, 340 Monetaria annulus, 240, 243 moneta, 240 Monticellina, 11 Moridae, 76 Morus bassanus, 130 Mugil, 360 Munidopsis cf. hystrix, 18 scobina, 18, 26, 32 Munnopsidae, 63

Muraenolepididae, 76 Murchisonella columna, 246 Murex forskoehlii, 246 Muricidae, 69, 269 Musculista perfragilis, 248 senhousia, 237, 248, 261, 267, 268, 270, 272 Mya arenaria, 242, 250, 253, 262, 268 Myoforceps aristatus, 265 Myrina (Adipicola) paciﬁca, 341 Mytilidae, 269, 337 Mytilus californianus, 176, 189 edulis, 176 galloprovincialis, 254, 268 Myxilla incrustans, 210 Myxinidae, 76 Myxiniformes, 76 Narcomedusae, 60 Nassarius arcularius plicatus, 246 denticulatus, 264 elatus, 264 festivus, 122, 123 goreensis, 264 heynemanni, 264 pfeifferi, 263, 264 vaucheri, 264 Natica fanel, 264 gualteriana, 244 marochiensis, 240 vittata, 264 Naticidae, 69 Nematoda, 10, 51, 78 Nematomorpha, 51 Nemertea, 52, 67, 78 Neogoniolithon strictum, 179, 190 Neoleptopsis, 339 Neorhodomela larix, 183, 191, 196, 199, 202, 205, 206 Nephasoma, 69 Nephrops, 52 Nephtys 27 ferruginea, 12 Nereid sp. 1, 324 Nereis, 27, 360 anoculis, 328, 337 Nerita sanguinolenta, 244 Neverita josephina, 261 Nezumia liolepus, 19, 23 stelgidolepis, 322 Nicrophorus, 129 Ninoe, 10 Nonionella, 20 stella, 25 Notarchus indicus, 241 Nototheniidae, 76

425

SYSTEMATIC INDEX

Nucella, 192 Nuculana bicuspidata, 265 Nuculanidae, 69 Numenius phaeopus, 136 Nymphon, 66 Nymphonidae, 66 Octocoral sp. A, 339 Ocypode ryderi, 143 Odobenus rosmarus, 122 Odontasteridae, 74 Odostomia lorioli, 246 Olavius algarvensis, 24, 31 crassitunicatus, 23, 30, 31 Onoba, 68 Onuphidae, 71 Onychophora, 51 Opalia crenata, 264 Opheliidae, 71 Ophiacanthidae, 74 Ophidiiformes, 76 Ophioblennius atlanticus, 291 Ophiolepidae, 74 Ophiolymna antarctica, 18 Ophiopthalmus normani, 18, 337 Ophiura, 73 Ophiuroidea, 74, 78, 87 Ophryotrocha, 322, 338 sp. A, 322 Opisthobranchia, 241, 246 Orania fusulus, 264 Orchestia, 131 gammarellus, 131 platensis, 126 scutigerula, 132 Orchomene, 62 nanus, 121, 122 obtusus, 11 Orthonecta, 51 Oscilla jocosa, 246 Osilinus sauciatus, 263, 264 Osmeridae, 360 Osmerus mordax, 360 Osteichthyes, 76, 78 Osteopelta, 337 ceticola, 341 cf. mirabilis, 343 mirabilis, 341 praeceps, 341 Osteopeltidae, 312 Ostracoda, 78 Ostreidae, 269 Oswaldella, 60 Pallenopsis, 66 Palmadusta lentiginosa, 244

Palmae, 120 Palmaria palmata, 193 Palpiphitime sp. nov., 341 Panarthropoda, 62 Pandanaceae, 120 Panopea glycimeris, 265 Panthera leo, 122 Panulirus argus, 200 Paphia textile, 250, 257 undulata, 256 Paracocculina, 337 cervae, 327, 341 Paractora dreuxi, 133 trichosterna, 128, 133 Paracucumidae, 74 Paracyathus stearnsii, 202 Paralomis manningi, 341, 342 multispina, 318, 322 Paraonis gracilis, 23 Paraprionospio pinnata, 12, 23, 27, 30 Parazoa, 58 Parborlasia corrugatus, 67 Pareledone, 68 Parvicardium hauniense, 241 Patella granularis, 186, 200 nigra, 263, 264 vulgata, 243 Pectinidae, 69 Pectinoidea, 11 Peinaleopolynoe, 338 santacatalina, 341 Pelagodiscus, 71 Pelosina, 21 Pelvetia canaliculata, 198 Penaeus, 28 Penicillus, 182, 186 vaginiferus, 241 Pentastoma, 52 Peracarida, 63 Perciformes, 76 Perna perna, 265 picta, 239, 244, 248 Perviata, 52 Petalifera gravieri, 241 Petricola hemprichii, 241 pholadiformis, 250, 262, 268 Petromyzontidae, 76 Petromyzontiformes, 76 Petroscirtes breviceps, 291 mitratus, 291 Phaeophyceae, 58 Phaeophyta, 58 Phaeostrophion irregulare, 181 Phalacrocorax aristotelis, 373 Phaleria, 130, 133 bimaculata, 133

426

SYSTEMATIC INDEX

provincialis, 133 Phallodrilinae, 24 Phascolidae, 69 Phascolion, 69 Phascolosoma, 338 saprophagicum, 312, 313, 338, 341 turnerae, 338 Phidoloporidae, 72 Philinidae, 69 Phoca vitulina, 130 Phocoena phocoena, 355–390 phocoena phocoena, 356 phocoena relicta, 356 phocoena vomerina, 356 Phocoenidae, 355 Phorona, 52 Phoxocephalidae, 63 Phragmatopoma californica, 176 Phyllodocidae, 71 Phyllospadix scouleri, 176 torreyi, 182 Phylobryidae, 69 Physeter macrocephalus, 142, 334 Phytosus nigriventris, 120 Pinctada margaritifera, 248, 255, 266, 272 radiata, 237, 248, 255, 256, 257, 258, 266, 267 Pinna rudis, 265 Pitar sewelli, 11 Placopecten magellanicus, 241 Placozoa, 51 Planaxis griseus, 244, 256, 257 Platyhelminthes, 51, 78 Pleurobranchus forskalii, 246 Pleuroncodes monodon 8, 28 planipes, 28 Pleuronectes ﬂesus ﬂesus, 360 platessa, 374 Pleuronectiformes, 76 Plocamopherus ocellatus, 246 Plumaria elegans, 198 plumosa, 198 Plumulariidae, 61 Pluvialis squatarola, 131, 136 Poecilosclerida, 59 Pogonophora, 52, 71, 78 Pollachius virens, 360 Pollicipes elegans, 28 Polycera hedgpethi, 246, 272 Polycerella emertoni, 246, 272 Polychaeta, 10, 70, 78, 87, 341 Polyclinidae, 75 Polyides rotundus, 181 Polynices lacteus, 240 Polynoidae, 71, 341 Polyplacophora, 68, 78, 244

Polysiphonia, 208 lanosa, 198 setacea, 184, 188 Polyzoa, 52 Pomacanthidae, 291 Pomacentridae, 291, 292, 293, 298, 299, 300 Pomacentrus fuscus, 291 variabilis, 291 Porifera, 51, 58, 59, 78, 87 Porphyra umbilicalis, 176, 198 Posidonia, 118, 140, 141, 142, 269 angustifolia, 118 oceanica, 118 sinuosa, 118 Posidoniaceae, 117 Potamides conicus, 238, 240, 242 Potamogetonales, 117 Priapula, 51, 62, 78 Primnoidae, 61 Prionospio, 23, 24 (Minuspio), 10, 26 Propeamussium cf. alcocki, 11 Prosipho, 68 Prosobranchia, 240, 241, 244, 254 Protoctista, 58 Protodorvillea, 15, 27 Protolira thorvaldsoni, 341 Provanna lomana, 324, 328, 339 Psammotreta praerupta, 250, 258 Pseudemys scripta elegans, 130 Pseudochama corbieri, 248 Pseudochitinopoma occidentalis, 209, 210 Pseudominolia nedyma, 244 Psolidae, 74 Pterasteridae, 74 Pterobranchia, 52 Pterocladia capillacea, 183 Pulmonata, 248, 265 Purpuradusta gracilis notata, 244 Pusionella nifat, 241 Pycnogonida, 65, 66, 78, 87 Pyramidellidae, 269 Pyropelta, 337, 340 corymba, 324, 325, 339, 340, 342 musaica, 323, 324, 325, 328, 329, 340, 342 wakeﬁeldi, 340, 341 Pyruidae, 75 Pyrunculus fourierii, 246 Rajidae, 76 Rajiformes, 76 Ranella, 254 Rapana rapiformis, 240 venosa, 237, 246, 253, 267, 268, 270, 272 Rathbunaster californicus, 18 Reinhardtius hippoglossoides, 360

427

SYSTEMATIC INDEX

Rhagidiidae, 65 Rhinoclavis kochi, 237, 244, 257, 266, 270, 271 Rhodochorton ﬂoridulum, 181, 187 purpureum, 187 Rhodomela larix, 183, 191, 199, 206 Rhodophyta, 58 Rhodothamniella ﬂoridula, 181, 187 Rhombozoa, 51 Rissa tridactyla, 373 Rissoidae, 69, 254, 269 Rissoina bertholleti, 244 chesneli, 240 decussatam, 240 spirata, 244, 255 Rotifera, 51 Ruditapes decussatus, 261, 272 philippinarum, 250, 261, 266, 268, 272 Sabellaria bella, 27 Sabellid sp. C, 339 Sabellidae, 71 Sabia conica, 244 Saccorhiza polyschides, 199 Saccostrea commercialis, 248, 261, 266, 272 cucullata, 248, 258 Salarias fasciatus, 291, 292, 297 guttatus, 292 luctuosus, 291 patzneri, 282, 285, 286, 292, 294, 297 Salariini, 295 Salinella salva, 52 Salmo salar, 359 Sardinops coerulea, 360 Sargassum, 119, 207, 338 horneri, 191 johnstonii, 142 microphyllum, 185, 213, 215 sinicola, 142, 183 Saxidomus purpuratus, 241 Scacchia zorni, 265 Scaliola elata, 240 Scalpellidae, 63, 65 Scaphopoda, 68, 78 Scarabaeidae, 120, 126, 128, 129 Scaridae, 291, 292, 293, 297, 299, 300 Scarus altipinnis, 291 ghobban, 291 schlegeli, 291, 292, 297 sordidus, 291 Schizasteridae, 74 Sclerodomidae, 72 Sclerodoris cf. tuberculata, 241 Sclerospongiae, 58, 59 Scomber scombrus, 359, 360 Scorpaeniformes, 76

Scyphozoa, 60 Sebastes marinus, 360 Sebastolobus alascanus, 11, 12, 23 altivelis, 23 Semibalanus balanoides, 176 Senoturbella bocki, 52 Sepiola atlantica, 370 Sepiolidae, 370 Septifer forskali, 248, 258 Serolidae, 63, 64 Serpulidae, 71 Sertulariidae, 61 Sicyona, 28 Siganidae, 299, 300 Sillago bassensis, 138 Silphidae, 129, 130 Sinum bifasciatum, 264, 266 Sinupharus combieri, 265 Siphonaria capensis, 184, 186, 196, 200 crenata, 248 pectinata, 239, 243, 248, 262, 265, 266 Sipuncula, 52, 68, 78, 341 Smaragdia souverbiana, 244, 258 Smittinidae, 72 Solatia piscatoria, 263, 265 Solea nasuta, 360 solea, 359, 363 Solenogastres, 68 Somniosus paciﬁcus, 318, 322, 337 Sphacelaria furcigera, 183 radicans, 176, 182 rigidula, 183 Sphenia rueppelli, 250 Spionidae, 10 Spirobrachia, 71 Spondylus, 255 cf. multisetosus, 248 groschi, 248 limbatus, 241 spectrum, 241 spinosus, 248 Sprattus sprattus, 360 Staphylaea nucleus, 240 Staphylinidae, 120, 128 Stegastes nigricans, 282, 292, 293, 296, 297 Stenella, 130 Stenolaemata, 72 Stenothoidae, 63 Stercorarius parasiticus, 373 Sterculiaceae, 120 Sterna paradisaea, 373 Sticteulima cf. lentiginosa, 244, 258 Stomatella impertusa, 244, 258 Stramonita haemastoma, 264 Strigatella virgata, 240 Strombus decorus, 270

428

SYSTEMATIC INDEX

lentiginosus, 240, 256 mutabilis, 244 persicus, 237, 244, 258, 266 Styela gagetyleri, 11, 24 Stylasterina, 60 Stylerasteridae, 61 Stylidae, 75 Styloptygma beatrix, 246 Syllid sp. A, 339 Syllidae, 71 Symbion pandora, 52 Symplasma, 51, 58, 59, 78, 87 Synallactidae, 74 Synopiidae, 62 Syringodium, 118 isoetifolium, 119 Syrnola cinctella, 246 fasciata, 246, 257

Trisopterus, 360, 361, 363, 370 esmarkii, 363 Trivia bitou, 264 Trochammina, 20 Trochidae, 69, 269 Trochus erythraeus, 244 Trophon, 68 Tubiﬁcidae, 23 Tubilopora, 209 Tunicata, 52 Turbellaria, 51, 78 Turbonilla edgari, 246 Turridae, 69 Tursiops, 357 truncatus, 357 Tylos capensis, 127, 133 granulatus, 133 punctatus, 133

Talitrus saltator, 131, 144 Talorchestia, 131 brito, 131 martensii, 127 Tamu, 343 Tanaidacea, 63, 64, 78 Tardigrada, 51, 78 Tectonatica ﬁlosa, 264 Tegula funebralis, 182 Tellina compressa, 265 valtonis, 250 Tenebrionidae, 126, 128 Terebellidae, 71 Teredo navalis, 254, 268 Tetraclita rubescens, 176 Thais lacera, 246, 255, 256, 257 sacellum, 246, 255 Thalassia hemprichii, 118 Thalassodendron, 118 ciliatum, 118, 119, 125, 142 Tharyx, 15, 27 luticastellus, 11 Thiomargarita, 1 nammibiensis, 7 Thioploca, 1, 7, 8, 16, 28, 30 araucae, 7 chileae, 7 Thoracica, 63 Thoracochaeta zostera, 135 Thyasiridae, 341 Tilapia mossambicus, 284 Timoclea roemeriana, 250 Trapezium oblongum, 248 Traskorchestia traskiana, 131 Trichoplax adhaerans, 51 Tricolia miniata, 263, 264 Tripterygiidae, 76

Udotea, 181, 186 Ulva, 178, 187, 190, 196, 199, 204 curvata, 204 lactuca, 142, 169, 193, 194, 197, 203, 204, 206, 208 lobata, 176 rigida, 204 scandinavica, 204 Umbonium vestiarium, 240 Undaria pinnatiﬁda, 196, 206, 220 Ungulina cuneata, 265 Uniramia, 52 Urechis caupo, 12 Uria aalge, 122, 373 Urochordata, 52, 75, 78, 87 Urodasys anorektoxys, 24 Urophycis tenuis, 360 Ursus americanus, 122 maritimus, 122 Uvigerina, 20 Vaejovis littoralis, 137 Valenciennea muralis, 292 Valettidae, 62 Vasum turbinellus, 240 Vaucheria thuretii, 181 velutina, 181 Velutinidae, 69 Veneridae, 269 Veneroidea, 11 Vertebrata, 52, 76, 78 Vesicomya, 340 gigas, 329, 339, 343 Vestimentifera, 52 Vexillum depexum, 241 Vigtorniella n. sp., 319, 322, 333, 338, 341 Voorwindia tiberiana, 244

429

SYSTEMATIC INDEX

Vulpes vulpes, 122 Womersleyella setacea, 184, 188 Xenonerilla bactericola, 12, 24, 30 Xenostrobus securis, 237, 248, 261, 267, 272 Xiphopenaeus riveti, 28 Xylodiscula, 337, 340 osteophila, 341 Xyloplax, 52

selasphora, 246 Zoantharia, 60 Zoarces viviparus, 363 Zoarcidae, 76 Zonaria farlowii, 181, 196, 197, 205 pyrum, 264 Zostera, 118, 125, 142 angustifolia, 136 marina, 126, 131, 133 noltii, 136 Zosteraceae, 117, 118

Zafra savignyi, 246, 257

430

SUBJECT INDEX

References to whole articles are given in bold type; references to sections of articles are given in italics; references to pages are given in normal type. Algae, in beach cast, 117 Ambergris, 142, 143 Antarctic marine benthic diversity, 47–114 aims of study, 55 biogeographic patterns, 88–90 biological diversity, 48–49 endemism, 86 estimates of benthic diversity, 80–83 important taxa, 83–85 inventory of marine benthic diversity, 57–77 Annelida, 69–71 Polychaeta 70–71 Brachiopoda, 71 Bryozoa, 72–73 Chelicerata, 65–66 Pycnogonida 65–66 Chordata, 75–77 Urochordata, 75 Vertebrata, 76–77 Cnidaria, 60–61 Crustacea, 62–65 Amphipoda, 62–63 Cirripedia, 65 Decapoda, 65 Isopoda, 64 Tanaidacea, 64 Ctenophora, 61 Echinodermata, 73–74 Echiura, 69 macroalgae, 58 Mollusca, 67–68 Nemertea, 67 Pogonophora, 71 Porifera and Symplasma, 58–59 Priapula, 62 Sipuncula, 68–69 latitudinal diversity variation, 86–88 macroecological analyses, 90–95 distribution of species among genera, 92 latitudinal diversity cline, 90–92 size and abundance, 92–95 marine diversity, 50–52 methods, 56–57 data compilation, 56 distributional analyses, 56–57 historical and macroecological analyses, 57 Southern Ocean, 53–55 spatial variation in richness, 86 species numbers in Antarctica, 77–80

Arabian Sea OMZ, 1, 3, 4, 5, 6, 7, 11, 15, 16, 20, 25, 27 Bacteria, in OMZs, 7–8, 12, 13 in reef detritus, 284, 285, 286 Bacterial breakdown of beach cast, 123, 124 Baltic Sea OMZ, 3, 12 Beach cleaning, effects on invertebrates, 143–144 Biodiversity patterns, whale falls, 329–330 Biogeographic and evolutionary relationships, of whale falls, 332–344 Biotechnological spinoffs, from whale falls, 344–345 Bioturbation and lifestyles, in OMZ, 28–30 Black Sea OMZ, 3, 12 By-catch, of porpoises, 355, 357, 363, 366, 367, 368, 369, 373, 374, 375, 376 C:N ratios, in detritus, 281, 283, 284 Carrion, in beach cast, 121–123 Chemoautotrophic assemblages, 311, 313, 314, 317, 324, 327, 332, 343 bacteria, 323, 326, 329 endosymbionts, 314, 323, 326, 328, 329, 330, 331, 342, 343 symbiosis, in OMZs, 31 Chesapeake Bay, hypoxia, 13 Chile margin OMZ, 7, 9, 10, 22, 26, 27 Competition with other predators, porpoises, 372–373 Convention on Biological Diversity, 48, 77 Cryptogenic species, 251, 253–255 DDTs, in porpoises, 376 Debris, inﬂuence of on beach ecosystems, 145–146 Detritus in epilithic algal matrix, use by coral reef ﬁshes, 279–309 detrital consumption, 290–295 detrital production processes, 287–290 from ﬁsh faeces, 290 from organic matter, 287–288 from water column, contribution to coral reef ﬁsh assemblages, 288–290 detritivorous ﬁshes, 297–302 deﬁnition, 290–295 feeding traits, 295–297 future research, 302–303 nature of, 281

431

SUBJECT INDEX

introduced range, 252 mode of arrival, 252–253 native range, 252 recorded Mediterranean range, 243–250 relation to current Mediterranean range, 250–252 type of larval development, 253 year and place of ﬁrst collection, 243 criteria for data selection, 240–243 geographic scope, 239 reasons for success, 266–271 characters related to establishment, 267–270 larval development, 269–270 factors related to recipient sites, 270–271 sources of introduced species, 253–271 Atlantic species, 262–266 cryptogenic species, 253–255 Indo-Paciﬁc immigrants, 255–260 introduction via shipping, 261–262 introductions for mariculture, 260–261 invasion pathways, 266

Detritus in epilithic algal matrix continued nutritional value, 283–286 quantity, 281–283 sources, 286–287 Diatoms, in reef detritus, 285, 286 Diet of NE Atlantic harbour porpoise, 355–390 competition with other predators, 372–373 consequences of diet for health and populations, 376–377 diet composition, 359–361 food consumption, 371–372 future research, 377–379 interactions with ﬁsheries, 373–374 long-term trends in diet, 374–376 opportunism and selection, 362–369 cause of death and diet variability, 367–369 geographical variation, 363 interannual variation, 365 male and female diets, 366–367 ontogenetic variation, 365–366 seasonal variation, 364 study methods, 358–359 where and how porpoises feed, 369–371 echolocation, 370 inferences from diet studies, 370–371 surface observations, 369 Dispersal stepping stones, whale falls, 343–344, 347 Dissolved organic matter, as source of detritus, 281, 288 Disturbance, by sediments, 163, 168, 171, 178, 187, 189, 191, 192, 195, 196, 217, 218, 219, 220, 221, 223 Diversity, Antarctic marine benthos, biogeographic patterns, 88–90 endemism, 86 estimates of, 80–83 latitudinal diversity cline, 90–92 Driftwood, 120 Dune formation and erosion 140, 141

Facilitation mechanisms in beach cast succession, 129 Facilitation models, whale falls, 331, 332 Fatty acids, in porpoises, 359, 379 Feeding preferences, beach wrack fauna, 131, 132 Fisheries interactions, porpoises, 373–374 Genetics, of seaweed ﬂies, 134 Grazing, inhibition by sediment, 191–193 Great Barrier Reef, 279, 281, 282, 284, 287, 289, 292, 297, 303 Gulf of California OMZ, 6, 31 Harbour porpoise, diet in NE Atlantic, 355–390 Harvesting, seaweeds and seagrasses, 141, 142 Hydrothermal vents, 311, 314, 328, 329, 333, 338, 342, 343, 344, 345, 346, 347 Hypoxia, adaptations to in OMZs, 8–14

Echolocation, by porpoises, 370 Edge effects, in OMZ, 20 El Niño, 7, 16, 18, 26, 27, 28, 29, 30 Endemics, in OMZs, 24 Endemism, Antarctic benthos, 86 Endobiotic bacteria, in OMZs, 30, 31 ENSO, 5, 16 Epilithic algal matrix, use of detritus in by coral reef ﬁshes 279–309 Erythrean invasion, 266 Exotic molluscs in the Mediterranean, 237–277 impact on native fauna and prospects, 271–272 methods, 239–253 attributes of species, 243–253 establishment status, 252

Introduced species, molluscs, 238, 252, 260, 261, 267, 270, 271, 272 Invasion pathways, exotic molluscs, 266 Invasive species, 238, 245, 266, 267 Island biogeography, 48 Isotope signatures, in OMZs, 30, 31 Isotope values, carbon, 328 nitrogen, 328, 329 Landslides, as source of sediments, 165, 166, 176, 177, 189 Larval development, molluscs, 253, 260, 269–270 Lessepsian immigrants, 239, 252, 255, 267, 271

432

SUBJECT INDEX

Litter decomposition, 125, 126 Lizard Island, 282, 284, 286, 292, 298, 299, 300, 301 Macroalgae, Antarctica, 58 Macrofauna in OMZs, 16–18, 25 Mariculture, 239, 253, 260–261, 266, 267, 271, 272 Marine allochthonous input, effect on sandy beaches, 115–159 breakdown of beach-cast material, 123–126 leaching and microbial processes, meiofauna, 123–126 effects of removal of wracks on invertebrates, 143–144 effects of stranded wrack, 138–140 beneﬁcial efects, 138–139 detrimental effects, 139–140 geomorphological importance, 140–141 human use of material, 141–143 harvesting seaweeds and seagrasses, 141–142 other uses, 142–143 inorganic beach-cast material, 144–146 inﬂuence of debris on beach ecosystems, 145–146 man-made marine debris, 144–145 macrofaunal beach-wrack communities, species succession, 126–135 herbivores, detritivores, predators, 126–127 inter- and interspeciﬁc interactions, 131–135 species succession in carrion, 129–131 in macrophyte wracks, 127–129 offshore consumers of wracks, 137–138 ﬁlter feeders, grazers, 137–138 organic components, 117–123 abundance of stranded macrophytes, 117–118, 118–119 algae, 117 seagrasses, 117–118 other organic material, 120–123 carrion, 121–123 driftwood, 120 fruits and seeds, 120–121 shorebirds and terrestrial animals linked to wracks, 135–137 Mediterranean, exotic molluscs, 237–277 history of fauna, 238 Megafauna in OMZs, 18–19 Meiofauna in OMZs, 16 Meiofauna, effect of beach cast on, 124, 125 Molluscs, exotic Mediterranean, 237–277 Okinawa Island, 291, 292

Oman margin OMZ, 6, 7, 9, 10, 11, 15, 17, 21, 22, 24, 26, 29, 31, 32 Organic pollutants, in porpoises, 376, 377 Oxygen minimum zone (OMZ) benthos, 1–45 biotic response to OMZs, 7–32 adaptations to permanent hypoxia, 8–14 community-level responses, 14–28 abundance and biomass, Foraminifera, 15–19 macrofauna, 15 megafauna, 16–18 metazoan meiofauna, 16, 18–19 body size and morphology, 14–15 community composition, 20–24 edge effects, 20 interannual and seasonal variability, 27–28 OMZ endemics, 24 reproductive modes, 26–27 spatial heterogeneity, 26 Foraminifera, 24–25 macrofauna, 24–25 ecosystem level responses, 28–32 bioturbation and lifestyles, 28–30 trophic pathways, 30–32 large sulphur bacteria, 7–8 nature of OMZs, 3–7 distribution, formation, temporal stability, 3–5 historical aspects, 7 sediments and oceanographic environment, 6 OMZ frontiers, 33–34 palaeoceanographic and evolutionary implications, 32–33 Paciﬁc Ocean OMZ, 3, 7, 11, 14 PCBs, in porpoises, 376, 377 Peru margin OMZ, 4, 5, 6, 9, 15, 17, 19, 20, 21, 22, 24, 27, 28, 29, 31 Polar front, 53, 54, 56, 62 Predation, inhibition by sediment, 191–193 Preston plots, 94, 95 Production, of detritus, 287–290 Recruitment, inhibition by sediment, 190–191 Rocky coast assemblages, sedimentation effects, 161–236 San Clemente Basin, whale falls, 316, 318, 324, 327 328, 340 San Diego Trough, whale falls, 316, 318, 319, 321, 322, 325, 327, 328, 329 Sandy beaches, effects of marine allochthonous input, 115–159 San Nicolas Slope, whale falls, 316, 318, 324, 327, 329

433

SUBJECT INDEX

Santa Catalina Basin OMZ, 18, 21 Santa Catalina Basin, whale falls, 313, 316, 317, 318, 324, 327, 328, 329 Santa Cruz basin, whale falls, 316, 317, 318, 321, 322, 328 Scale, importance in sedimentation studies, 223–224 Scavenger assemblages, on whale falls, 337 Scour, by sediments, 163, 170, 171, 173, 179, 180, 189, 190, 191, 195, 196, 206, 217, 218, 221, 222, 224 Seagrasses, harvesting, 142 in beach cast, 117–118 Sediment traps, 6, 165, 173, 190, 287, 288, 289 Sedimentation, effects on rocky coast assemblages, 161–236 context of study, 161–162 ecological role of sedimentation 221–222 effects of sedimentation, 174–221 ﬁeld experiments, 207–221 measured effects, 216–220 methodological problems, 207–216 laboratory experiments, 196–207 modelling, 220–221 observations, 175–196 attributes of species, 179–187 changes in species composition and abundance, 176–178 effects of sediments from human discharges, 193–194 inhibition of recruitment, growth and survival, 189–193 responses of individual species, 178–179 sediment- species diversity relationship, 194–196 species trapping and binding sediments, 187–189 future research, 224–225 importance of scale, 223–224 predicting impacts of sediments, 222–223 sediment load to rocky coasts, 164–173 control by biological factors, 168–169 measurement of deposition and accumulation, 171–173 intertidal habitats, 172 subtidal habitats, 172–173 measurement of scour, 173 mechanisms, 170–171 problems in measurement and comparison, 171 sources, 164–167 human-related processes, 166–167 natural processes, 164–166 variability of sediment deposition, 167–168 Selective feeding, coral reef fishes, 295, 296, 297

Settlement, inhibition by sediment, 190–191 Shipping, and introduced species 239, 253, 261–262, 266, 267, 272 Shorebirds linked to wracks, 135–137 Short chain fatty acids, in herbivorous ﬁshes, 294 Species diversity, in OMZs, 24–25 Species richness, Antarctica, 48, 57, 77, 86 Mediterranean molluscs, 270, 271, 272 Species succession in macrophyte wracks, 127–131 Stable isotopes, in porpoises, 359 Succession, on whale falls, 315–326, 346–347 Suez canal, molluscs, 238, 239, 252, 253, 257, 258, 260 Sulphophilic communities, 338–339 Sulphur bacteria, in OMZs, 7–8 Trophic pathways, in OMZs, 30–32 Trophic relationships, on whale falls, 326–329 Tropicalisation, Mediterranean molluscs 239, 262, 263, 266 Turfs, 180, 187, 188, 189, 280, 286 Veil line, 95 Vent and seep afﬁnities, of whale falls, 338–340 Walvis Bay OMZ, 14, 16, 17, 25, 28 Whale falls, ecology at deep-sea ﬂoor, 311–354 anthropogenic inﬂuences, 345 biogeographic and evolutionary relationships, 332–344 ancient/evolutionary relationships, 342–345 dispersal stepping stones, 343–344 whales and reptiles, 342–343 modern relationships, 332–342 abundance, 332–336 relation of California whale falls to other modern communities, 337–342 enrichment opportunists, 338 hard substratum biota, 340 plant/other organic substrata, 337–338 scavenger assemblages, 337 vent and seep afﬁnities, 338–340 whale-fall specialists, 340–342 biotechnological spinoffs, 344–345 future directions, 346–347 biogeography and evolution, 347 macrofaunal reproduction, dispersal, gene ﬂow, 346 microbial community structure, 346 succession, 346–347 whale-, kelp- and wood-fall relations, 347 manipulative studies off S. California, 314–332

434

SUBJECT INDEX

biodiversity patterns, 329–330 patterns of succession, 315–326 enrichment-opportunist stage, 319–322 mobile-scavenger stage, 318–319 reef stage, 326

sulphophilic stage, 322–325 succession structure and mechanisms, 330–332 trophic relationships, 326–329 scientiﬁc history, 312–314

435

Oceanography and Marine Biology, An Annual Review, Volume 41: An Annual Review: Volume 41 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review

Oceanography and Marine Biology: An Annual Review

Oceanography and Marine Biology, An Annual Review, Volume 40: An Annual Review: Volume 40 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review: Volume 39

Oceanography and Marine Biology: An Annual Review, Volume 43 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, Volume 45 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, Volume 48

Oceanography and Marine Biology: An Annual Review, volume 47 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review Volume 42 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, volume 47 (Oceanography and Marine Biology)

Oceanography And Marine Biology (Oceanography and Marine Biology - An Annual Review)

Oceanography and Marine Biology: An Annual Review, Vol. 36

Oceanography And Marine Biology

Oceanography And Marine Biology

Advances in Marine Biology, Volume 41 (Advances in Marine Biology)

Marine Optics (Elsevier Oceanography Series, Volume 14)

Biotechnology Annual Review, Volume 12 (Biotechnology Annual Review)

Biotechnology Annual Review, Volume 14

Biotechnology Annual Review, Volume 9 (Biotechnology Annual Review)

Biotechnology Annual Review, Volume 10 (Biotechnology Annual Review)

Marine Optics (Elsevier Oceanography Series, Volume 14)

Biotechnology Annual Review, Volume 2

Biotechnology Annual Review, Volume 3

Biotechnology Annual Review, Volume 1

Marine Mineral Exploration (Elsevier Oceanography Series V.41)

Oceanography And Marine Biology: A Book of Techniques

Advances In Marine Biology, Volume 46 (Advances in Marine Biology)

Advances in Marine Biology, Volume 44 (Advances in Marine Biology)

Advances In Marine Biology, Volume 51 (Advances in Marine Biology)

Annual Reports in Medicinal Chemistry, Volume 41

Oceanography and Marine Biology, An Annual Review, Volume 41: An Annual Review: Volume 41 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review

Oceanography and Marine Biology: An Annual Review

Oceanography and Marine Biology, An Annual Review, Volume 40: An Annual Review: Volume 40 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review: Volume 39

Oceanography and Marine Biology: An Annual Review, Volume 43 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, Volume 45 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, Volume 48

Oceanography and Marine Biology: An Annual Review, volume 47 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review Volume 42 (Oceanography and Marine Biology)

Oceanography and Marine Biology: An Annual Review, volume 47 (Oceanography and Marine Biology)

Oceanography And Marine Biology (Oceanography and Marine Biology - An Annual Review)

Oceanography and Marine Biology: An Annual Review, Vol. 36

Oceanography And Marine Biology

Oceanography And Marine Biology

Advances in Marine Biology, Volume 41 (Advances in Marine Biology)

Marine Optics (Elsevier Oceanography Series, Volume 14)

Biotechnology Annual Review, Volume 12 (Biotechnology Annual Review)

Biotechnology Annual Review, Volume 14

Biotechnology Annual Review, Volume 9 (Biotechnology Annual Review)

Biotechnology Annual Review, Volume 10 (Biotechnology Annual Review)

Marine Optics (Elsevier Oceanography Series, Volume 14)

Biotechnology Annual Review, Volume 2

Biotechnology Annual Review, Volume 3

Biotechnology Annual Review, Volume 1

Marine Mineral Exploration (Elsevier Oceanography Series V.41)

Oceanography And Marine Biology: A Book of Techniques

Advances In Marine Biology, Volume 46 (Advances in Marine Biology)

Advances in Marine Biology, Volume 44 (Advances in Marine Biology)

Advances In Marine Biology, Volume 51 (Advances in Marine Biology)

Annual Reports in Medicinal Chemistry, Volume 41

Recommend Documents