The Effects of Fishing on Marine Ecosystems

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

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

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

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

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

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

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

1. GENERAL INTRODUCTION

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

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Tropical shelves Upwellings Non-tropical shelves

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

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

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

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

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

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7 1 Tunas, bonitos Miscellaneous

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Mackerels, snooks 8 cutlassfishes

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

2. BENTHIC FAUNA AND HABITAT 2.1. Introduction

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

t

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

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2.2.2. Static Fishing Gears

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

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

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

2.2.3. Drive Netting, Poisons and Explosives Techniques such as drive-netting, pull-seining, poison and explosive fishing are principally used by small-scale and artisanal fishers fishing on tropical reefs. Although the effects attributable to the activities of individual fishers are often small in comparison with those attributable to commercial fishing boats using towed gears, the combined effects of their activities are considerable given the large proportion of the coastal population involved in fishing (Pauly, 1988; Pauly er al., 1989; Dalzell et al., 1996). Many of the fishing techniques used to catch reef-associated fishes cause direct physical damage to the reef substratum. The most widely used destructive fishing techniques are drive netting (Carpenter and Alcala, 1977; Gomez et al., 1987), trapping (Munro et al., 1987) and explosive fishing (Munro et al., 1987). In addition, those poisons widely used to catch fishes for the aquarium trade and consumption have the potential to cause chemical damage to corals and nontarget fishes and invertebrates (Rubec, 1986; Eldredge, 1987; McAllister, 1988; Pyle, 1993). Corals perform several important functions in tropical environments. They provide substrata for primary production, habitats for invertebrates and fishes and often play a key role in protecting coasts from wave exposure and erosion. The rate at which reefs develop is determined by the balance between rates of accretion owing to the growth of corals, hydrocorals and coralline algae and erosion owing to mechanical processes and bioerosion. Fishing affects reefs directly when gears contact the reef substratum or indirectly by altering the relationships between those communities of plants, invertebrates and fishes which determine rates of reef accretion and bioerosion (see Section 2.3). Coral accretion relies upon the successful settlement of young corals, and the maintenance of suitable conditions for their growth (Pearson, 1981). These processes may be affected by fishing activities.

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Drive netting techniques are used to catch a range of reef-associated fishes which shelter within the reef matrix or shoal above the reef. These techniques are extensively used on coral reefs, and may range from smallscale village-based operations involving four or five fishers to large commercial operations which target offshore reefs in the Philippines and South China Sea and involve hundreds of divers (McManus, 1996). The process of drive netting requires that the fishers (who stand on the reef or dive) scare reef-associated fishes towards an encircling net or trap, using scaring devices such as weighted lines or poles. In shallow water, corals are often broken deliberately to scare closely reef-associated fishes such as groupers (Epinephelinae), snappers or emperors from their refuges. In deeper water, the Kayakas and muro-ami drive-netting techniques involve teams of swimmers which repeatedly drop weighted scarelines onto the reef in order to drive fishes towards a bag net. Carpenter and Alcala (1977) calculated the damage to 1ha of reef during a single muro-ami operation involving 50 fishers who each struck the bottom 50 times with a 4 kg weighted scareline. Six percent of the total area of coral present was damaged. Blast fishing is practised on many reefs in the Atlantic, Pacific and Indian Oceans (Gomez et al., 1981; Polunin, 1983; Alcala and Gomez, 1987; Galvez and Sadorra, 1988; Ruddle, 1996). A variety of explosives are used including those obtained from mines or removed from armaments. Pelagic fishes living above the reef are often targeted rather than fishes living in direct association with the reef (Saila et al., 1993). Owing to the considerable variation in the types and sizes of charges used, and the depths at which they explode, it is difficult to make useful generalizations about the damage which they will cause. Alcala and Gomez (1987) report that a bottle bomb exploding at or near the bottom will shatter all corals within a radius of 1.15m, and that a gallon-sized drum will have the same effect within a radius of 5 m. A “typical” charge will kill most marine organisms including invertebrates within a radius of 77m. Such techniques are highly unselective and Munro et al. (1987) report that post-larval and juvenile fishes are also killed. These young fishes would be about to recruit to the reef habitat, and the repeated effects of blast fishing on a large scale would reduce fish production from the reef. On many reefs from 15-30’ either side of the equator, which are susceptible to hurricane damage, the effects of blast fishing are often localized and negligible in comparison with those of hurricanes (S. Jennings, pers. obs.). However, intensively blast-fished areas will be subjected to chronic rather than the acute damage associated with natural events. In areas such as the Philippines, damage attributable to blast fishing is an increasing cause of concern. Stupefacients are widely used by reef fishers. Traditionally, poisons extracted from plants were extensively used for reef fishing, but in the last few decades, synthetic chemicals such as sodium cyanide and chlorine have

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been used more frequently (Rubec, 1986; Eldredge, 1987). McAllister (1988) estimated that 150 t of sodium cyanide is used annually on Philippine reefs to catch aquarium fishes. There is little knowledge of the effects of these chemicals on the various life-history stages of the reef biota (Rubec, 1986; Pyle, 1993) and while concentrations of stupefacients which have an acute effect are quickly dispersed, the chronic effects may be significant. The long-term direct effects of fishing on reefs are largely determined by the rate at which coral can accrete in relation to the rate at which it is damaged. The recovery and recolonization of coral communities following mechanical damage by fishing gears takes place when partially damaged colonies or coral fragments regrow and when the substratum becomes suitable for coral settlement (Pearson, 1981). Saila et al. (1993) developed a model to examine the effects of blast fishing on reefs in the Philippines. At present fishing intensities, the loss of diversity and coral cover would continue for approximately 25 years before recovery is expected. Coral growth rates are highly variable: 0.7-17.2 cm year-’ for branching species and 0.5-1.9 cm year-’ for massive species (Loya, 1976; Huston, 1985; Witman, 1988): Several studies of reef development following hurricanes and other natural events provide a useful guide to recovery rates. Published estimates of recovery time often vary widely because they reflect differences in the authors’ assumptions regarding the organization of coral communities and the meaning of “stability” (Moran, 1986, 1990; Done, 1987, 1988; Done et al., 1988; Endean et al., 1988; Turner, 1994; McClanahan et al. 1996). However, a coral community dominated by fast-growing branching species and which provides a suitable habitat for many reef fishes would develop within 5 years (Pearson, 1981). 2.3. Indirect Effects on Habitat

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

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

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

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sediment deposition has been shown to inhibit the settlement and growth of oysters and scallops (Moore, 1977; Jones, 1992). However, given the range of sediment types in the marine environment and the natural spatial and temporal variations in sediment load (Moore, 1977), it is unlikely that the population level consequences of sediment resuspension and deposition can be determined from small scale studies of siltation effects. The surface of marine sediments is an important site of benthic production. Brylinsky er al. (1994) demonstrated that the biomass of benthic diatoms (measured as chlorophyll a) was significantly less in trawl door furrows on a muddy substratum in shallow water. However, one month after the trawling had taken place there was a diatom bloom in the furrow, which Brylinsky et al. (1994) attributed to the release of nutrients from the sediment. Emerson (1989) considered the effects of sediment disturbance resulting from wind stress on production in the southern North Sea, and found a significant negative correlation between wind stress and total macro- and meiobenthic production. The intensive trawling of Posidoniu oceanica meadows in the Mediterranean Sea may lead to reductions in littoral primary productivity, since large areas of P. oceanica are reported to have been killed by the mechanical action of fishing gears and the deposition of resuspended sediment (GuillCn er al., 1994). These meadows are known to be important sources of primary production, although the consequences of losses in production are not known. It is unlikely that large-scale changes in primary production could be reliably correlated with changes in fishing intensity using existing data. However, given that a large proportion of the continental shelf area is now trawled, and that tools such as stable isotope analysis can be used to trace the origins and transformations of organic matter in the marine environment (Owens, 1987; see Section 5.3), it is increasingly likely that the impacts of fishing on the relative roles of benthicand planktonic-based food chains could be investigated. The most convincing evidence for the indirect effects of fishing on habitat structure comes from the study of fishing effects on coral reefs. The direct effects of fishing (see Section 2.2.3) have been widely reported because popular reef fishing techniques often cause rapid and highly visible damage. However, the intensity and selectivity of fishing practices may have been responsible for initiating the transition of reef communities between relatively stable algal and coral-dominated phases (Levitan, 1992). Understanding the ways in which fishing can lead to shifts in ecosystem state is dependent on an understanding of the roles of herbivores and corallivores in the reef ecosystem. Herbivorous and corallivorous species both erode the reef matrix and the rate at which coral reefs grow is determined by the relative rates of coral accretion and erosion. Some species of parrotfishes (Scaridae) and urchins erode the reef matrix while feeding on algae (Birkeland, 1989; Bak, 1990, 1994; McClanahan, 1992, 1995a; Bellwood,

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

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

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

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

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

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

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

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

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

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

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

3. FISH COMMUNITY STRUCTURE

3.1. Introduction

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

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

3.2. Diversity 3.2.1. Local Extinctions and Redundancy

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

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

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

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

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Jennings et al., 1995.)

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

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

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Year Figure 12 Estimated changes in the total biomass (open circles) and biomass composition (open triangles, herring and mackerel; filled squares, species other than herring and mackerel) of the North Sea fish community. Biomass estimates were calculated using a method similar to virtual population analysis. (Redrawn from Ursin, 1982.)

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

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

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

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

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

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

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3.4. Life History Traits

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

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

2 49

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

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

Ste. Anne Reserve (some fishing occurs) invertebrate feeders

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

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a

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

3.4.2. Intraspecijic Changes in L$e Histories

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

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

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

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

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

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

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

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

4. TROPHIC INTERACTIONS 4.1. Introduction

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

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

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

Decreases in the abundance of large fished predators, such as the bluefin tuna Thunnus thynnus (L.) or the blue whale Baluenopteru rnusculus (L.) have been amongst the most widespread and thoroughly documented direct effects of fishing. Given that many of these species were very abundant and that their prey consumption rates were high, there have been many attempts to correlate their decline with the proliferation of their prey (e.g. Laws, 1977; Tiews, 1978). On coral reefs, where the functional and species diversity of fishes is relatively high, the indirect effects of fishing on the abundance of unfished prey species appears to be minor. While Jones (1982), Grigg et al. (1984) and Munro and Williams (1985) have suggested that a fishing strategy that selectively targets piscivorous fishes may lead to increases in the production or biomass of their prey, empirical evidence suggests that any changes in the abundance of prey species following the capture of their predators may not be detectable. On the scale of kilometres to tens of kilometres Bohnsack (1982), Russ (1985), Jennings et al. (1995) and Jennings and Polunin (1997) documented significant decreases in the abundance of piscivorous target species following fishing and yet there was no evidence for a corresponding increase in the abundance of their prey (Figure 17). This result is apparently contradicted by those studies that demonstrate increases in the abundance of small fast-growing species from low trophic levels in catches from intensively fished areas. However, the changes in catch composition which have been reported may result from fishing activities altering the fished habitat, from fishers shifting their attention to the only remaining resources or from fishers reducing the amount of catch discarded. These factors have rarely been treated explicitly in previous assessments but would not have affected the results of Jennings et al. (1995) and Jennings and Polunin (1997). They used fishery independent visual census methods to estimate abundance, compared habitats at sites subject to different fishing intensities, and worked in areas where destructive fishing practices were not in use. It is clear that the power to detect minor changes in prey populations at the large spatial scales

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intensities (Jennings and Polunin, 1997).

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

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

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

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

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

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

4.3.1. Fishes

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

5. STUDY OF FISHING EFFECTS 5.1. Introduction

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

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

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

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

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

5.3. Investigating Marine Food Webs

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

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

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

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

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

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

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

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

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

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

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

5.7. Conclusions

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

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6. MANAGEMENT

6.1. Introduction

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

6.2. The Role of Fisheries in Marine Ecosystems

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

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

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6.3. From Population to Ecosystem Management

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

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

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

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

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

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

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

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

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

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

7. SUMMARY

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

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

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

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

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Abrams, P.A. (1996). Evolution and the consequences of species introductions and deletions. Ecology 77, 1321-1328. Adams, P.B. (1980). Life history patterns in marine fishes and their consequences for management. Fishery Bulletin 78, 1-12. Aebischer, N.J. and Wanless, S. (1992). Relationships between colony size, adult non breeding and environmental conditions for shags Phalacrocorax aristotelis on the Isle of May, Scotland. Bird Study 39, 43-52. Ainley, D.G., Sydeman, W.J. and Norton, J. (1995). Upper trophic level predators indicate interannual negative and positive anomalies in the California current food web. Marine Ecology Progress Series 118, 69-79. Akenhead, S.A., Carscadden, J.E., Lear, H., Lilly, G.R. and Wells, R. (1982). Codcapelin interactions off northeast Newfoundland and Labrador. Canadian Special Publications in Fisheries and Aquatic Sciences 59, 141-148. Alcala, A.C. (1988). Effects of marine reserves on coral fish abundances and yields of Philippine coral reefs. Ambio 17, 194-199. Alcala, A.C. and Gomez, E.D. (1987). Dynamiting coral reefs: a resource destructive fishing method. In “Human Impacts on Coral Reefs: Facts and Recommendations” (B. Salvat, ed.), pp. 51-60. Antenne Museum, French Polynesia. Alcala, A.C. and Russ, G.R. (1988). A direct test of the effects of protective management on the abundance and yield of tropical marine resources. Journal du Conseil International pour I’Exploration de la Mer 46, 40-47. Alm,G. (1959). Connection between maturity, size and age in fishes. Reports of the Institute of Freshwater Fisheries Research, Drottingholm 5, 5-145. Alverson, D.L., Freeberg, M.H., Pope, J.G. and Murawski, S.A. (1994). A global assessment of fisheries bycatch and discards. F A 0 Fisheries Technical Paper 339, 233. Andersen, K.P. and Ursin, E. (1977). A multispecies extension to the Beverton and Holt theory of fishing with acounts of phosphorus circulation and primary production. Meddelelser fra Danmarks Fiskeri- og Havundersogelser, New Series, 7, 319-435. Andersen, K.P. and Ursin, E. (1978). A multispecies analysis of the effects of variations of effort upon stock composition of eleven North Sea fish stocks. Rapports et Procis- Verbaux des Rkunions. Conseil Permanent International pour I’Exploration de la Mer 172, 286-291. Anderson, D.W., Gress, F. and Mais, K.F. (1982). Brown pelicans: influence of food on reproduction. Oikos 39, 23-3 1. Andrew, N.L. and Pepperell, J.D. (1992). The by-catch of shrimp trawl fisheries. Oceanography and Marine Biology, an Annual Review 30,527-565. Anker-Nilssen, T., Barrett, R.T. and Krasnov, J.V. (1997). Long- and short-term responses of seabirds in the Norwegian and Barents Seas to changes in stocks of prey fish. In “Role of Forage Fishes in Marine Ecosystems”, pp. 683-698. Alaska Sea Grant College Program Report, University of Alaska, Fairbanks. Anon. (1885). “Report of the Commissioners appointed to inquire and report upon the complaints that have been made by line and drift net fishermen of injuries sustained by them in their calling owing to the use of the trawl net and beam trawl in the territorial waters of the United Kingdom”. Eyre and Spottiswoode, London.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

315

Anon. (1993a). “North Sea Quality Status Report 1993”. Olsen and Olsen, Fredensborg. Anon. (1993b). Biogeographical identification of channel fish and shellfish stocks. Report to Commission of the European Communities by MAFF Fisheries Laboratory, Lowestoft. Anon. (1 994). Report of the multispecies assessments working group. International Councilfor the Exploration of the Sea, Committee Meeting 19941Assess: 9, 117 pp. Anon. (1995a). Report of the herring assessment working group for the area south of 62“ N. International Council for the Exploration of the Sea. Committee Meeting Assess 13, 303 pp. Anon. (1995b). Report of the study group on ecosystem effects of fishing activities. ICES Co-operative Research Report 200, 120 pp. Anon. (1995~).Report of the working group for the assessment of demersal stocks in the North Sea and Skagerrak. International Councilfor the Exploration of the Sea, Committee Meeting Assess 8: (Part I), 460 pp. Anon. (1995d). Report of the working group on the assessment of Norway pout and sandeel. International Council for the Exploration of the Seas, Committee Meeting Assess 5, 145 pp. Anon. (1995e). Report on the working group on the assessment of mackerel, horse mackerel, sardine and anchovy. International Council for the Exploration of the Sea, Committee Meeting Assess 2, 329 pp. Anon. (19950. Understanding marine biodiversity: a research agenda for the nation. National Academy Press, Washington. Anon. (1996a). “Industrial Fisheries: from Fish to Fodder”. Greenpeace UK, London. Anon. (1996b). Report of the working group on the ecosystem effects of fishing activities. International Council for the Exploration of the Sea, Committee Meeting G:l. Anon. (1996~).Seabird/fish interactions, with particular reference to seabirds in the North Sea. ICES Cooperative Research Report 216, 87 pp. Anon. (1997). Briefing: fisheries science. Nature, London 386, 105-1 10. Appeldoorn, R.S. (1996). Model and method in reef fishery assessment. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 219-248. Chapman and Hall, London. Arntz, W.E. and Weber, W. (1970). Cyprina islandica (Mollusca: Bivalvia) als Nahrung von Dorsch und Kliesche in der Kieler Bucht. Berichte der Deutschen Wissenschaftlichen Kommission fur Meeresforschung 21, 193-209. Atkinson, R.J.A. and Nash, R.D.M. (1990). Some preliminary observations on the burrows of Callianassa subterranea (Montagu) Decapoda: Thalassinidea from the west coast of Scotland. Journal of Natural History 24, 403-413. Auster, P. J. and Malatesta, R. J. (1995). Assessing the role of non-extractive reserves for enhancing harvested populations in temperate and boreal marine systems. In “Marine Protected Areas and Sustainable Fisheries” (N. Shackell and J.H.N. Willison, eds), pp. 82-89. Science and Management of Protected Areas Association, Wolfville. Auster, P. J. and Shackell, N. L. (1997). Fishery reserves: a tool for managing groundfish resources. In “Northwest Atlantic Groundfish: Perspectives on a Fishery Collapse” (J.G. Boreman, H.W. Nakashima, J.A. Powles, J.A. Wilson and R.L. Kendall, eds), in press. American Fisheries Society, Bethesda. Auster, P.J., Malatesta, R.J., Langton, R.W., Watling, L., Valentine, P.C., Donaldson, C.L., Langton, E.W., Shepard, A.N. and Babb, I.G. (1996). The

316

S. JENNINGS AND M.J. KAISER

impacts of mobile fishing gear on seafloor habitats in the Gulf of Maine (Northwest Atlantic): implications for conservation of fish populations. Reviews in Fisheries Science 4, 185-202. Bailey, R.S. (1991). The interaction between sandeels and seabirds - a case study in Shetland. International Councilfor the Exploration of the Seas, Committee Meeting 1991/L41. Bailey, R.S., Furness, R.W., Gauld, J.A. and Kunzlik, P.A. (1991). Recent changes in the population of the sandeel (Ammodytes marinus Raitt) at Shetland in relation to estimates of seabird predation. ICES Marine Science Symposia 193, 209-216. Bak, R.P.M. (1990). Patterns of echinoid bioerosion in two Pacific coral reef lagoons. Marine Ecology Progress Series 66, 267-272. Bak, R.P.M. (1994). Sea urchin bioerosion on coral reefs: place in the carbonate budget and relevant variables. Coral Reefs 13, 99-103. Bak, R.P.M. and Engel, M.S. (1979). Distribution, abundance and survival of juvenile hermatypic corals (Scleractinia) and the importance of life history strategies in the parent coral community. Marine Biology 54,341-352. Bak, R.P.M., Carpay, M.J.E. and de Ruyter van Steveninck, E.D. (1984). Densities of the sea urchin Diadema antillarum (Philippi) before and after mass mortalities on the coral reefs of Curacao. Marine Ecology Progress Series 17, 105108. Bannerot, S.P.,Fox, W.W. and Powers, J.E. (1987). Reproductive strategies and the management of tropical snappers and groupers. In “Tropical Snappers and Groupers: Biology and Fisheries Management” (J.J. Polovina and S. Ralston, eds), pp. 561403. Westview Press, Boulder. Bannister, R.C.A. (1978). Changes in plaice stocks and plaice fisheries in the North Sea. Rapports et ProcPs-Verbaux des Rkunions, Conseil Permanent International pour 1’Exploration de la Mer 172,86-101. Barel, C.D.N., Ligtvoet, W., Goldschmidt, T., Witte, F. and Goudswaard, P.C. (1991). The haplochromine cichlids of Lake Victoria: an assessment of biological and fisheries interests. In “Cichlid Fishes: Behaviour, Ecology and Distribution” (M.H.A. Keenleyside, ed.), pp. 258-279. Chapman and Hall, London. Barrett, R.T., Anker-Nilssen, T., Rikardsen, F., Valde, K., Rov, N. and Vader, W. (1987). The food, growth and fledging success of Norwegian puffin chicks Fratercula arctica in 1980-1983. Ornis Scandanavia 18,73-83. Baumgartner, T.R., Soutar, A., Ferreira-Bartrina, V. (1992). Reconstruction of the Pacific sardine and northern anchovy population over the past two millenia from sediments of the Santa Barbara basin, California. California Co-operative Fishery Investigations Report 33, 2440. Bax, N.J. (1991). A comparison of the biomass flow to fish, fisheries and mammals in six marine ecosystems. ICES Marine Science Symposia 193,217-224. Bech, G. (1995). Retrieval of lost gillnets at Ilulissat Kangia. NAFO Science Council Research Document 95/67.

Beddington, J.R. and Cooke, J.G. (1983). The potential yield of fish stocks. F A 0 Fisheries Technical Paper 242. Beddington, J.R. and de la Mare, W.K. (1985). Marine mammal fishery Interactions: modelling and the Southern Ocean. In “Marine Mammals and Fisheries” (J.R. Beddington, R.J.H. Beverton and D.M. Lavigne, eds), pp. 94-105. George Allen and Unwin, London. Beddington, J.R., Beverton, R.J.H. and Lavigne, D.M. (eds) (1985). “Marine Mammals and Fisheries”. George Allen and Unwin, London.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

317

Beek, F.A.v., Leeuwen, P.1.v. and Rijnsdorp, A.D. (1990). On the survival of plaice and sole discards in the otter-trawl and beam-trawl fisheries in the North Sea. Netherlands Journal of Sea Research 26, 151-160. Bellwood, D.R. (1995). Direct estimate of bioerosion by two parrotfish species, Chlorurus gibbus and C . sordidus, on the Great Barrier Reef, Australia. Marine Biology 121, 419-429. Bengston, J.L. and Laws, R.M. (1985). Trends in crabeater seal age at maturity: an insight into antarctic marine interactions. In “Antarctic Nutrient Cycles and Food Webs” (W.R. Siegfried, P.R. Condy and R.M. Laws, eds), pp. 67M75. SpringerVerlag, New York. Berghahn, R. (1990). On the potential impact of shrimping on trophic relationships in the Wadden Sea. In “Trophic Relationships in the Marine System” (M. Barnes and R.N. Gibson, eds), pp. 13CL140. Aberdeen University Press, Aberdeen. Bergman, M.J.N. and Hup, M. (1992). Direct effects of beamtrawling on macrofauna in a sandy sediment in the southern North Sea. ICES Journal of Marine Science 49, 5-1 1. Bergman, M.J.N. and Santbrink, J.W.V. (1994). Direct effects of beam trawling on macrofauna in sandy areas off the Dutch coast. In “Environmental Impact of Bottom Gears on Benthic Fauna in Relation to Natural Resources Management and Protection of the North Sea” (S.J. de Groot and H.J. Lindeboom, eds), pp. 209-236. Netherlands Institute for Sea Research, Den Burg, Texel. Beukema, J.J. (1995). Long-term effects of mechanical harvesting of lugworms Arenicola marina on the zoobenthic community of a tidal flat in the Wadden Sea. Netherlands Journal of Sea Research 33, 219-221. Beverton, R.J.H. (1963). Maturation, growth and mortality of clupeid and engraulid stocks in relation to fishing. Rapports et Prods- Verbaux des RPunions, Conseil International pour I’Exploration de la Mer 154, 44-67. Beverton, R.J.H. (1985). Analysis of marine mammal-fisheries interactions. In “Marine Mammals and Fisheries” (J.R. Beddington, R.J.H. Beverton and D.M. Lavigne, eds), pp. 3-33. George Allen and Unwin, London. Beverton, R.J.H. (1987). Longevity in fish: some ecological and evolutionary perspectives. In “Ageing Processes in Animals” (A. D. Woodhead, M. Witten and K. Thompson, eds), pp. 161-186. Plenum Press, New York. Beverton, R.J.H. (1990). Small marine pelagic fish and the threat of fishing: are they endangered? Journal of Fish Biology 37 (Supplement l), 5-16. Beverton, R.J.H. (1992a). Fish resources, threats and protection. Netherlands Journal Of Zoology 42, 139-175. Beverton, R.J.H. (1992b). Patterns of reproductive strategy parameters in some marine teleost fishes. Journal of Fish Biology 41 (Supplement), 137-160. Beverton, R.J.H. and Holt, S.J. (1957). “On the Dynamics of Exploited Fish Populations”. Ministry of Agriculture, Fisheries and Food, London. Birkeland, C. (1989). The influence of echinoderms on coral-reef communities. In “Echinoderm Studies” (M. Jangoux and J.M. Lawrence, eds), pp. 1-79. Balkema, Rotterdam. Birkeland, C . (1997). Symbiosis, fisheries and economic development on coral reefs. Trendr in Ecology and Evolution 12, 364-367. Blaber, S.J.M., Milton, D.A. and Rawlinson, N.J.F. (1988). Diets of lagoon fishes of the Solomon Islands: predators of tuna baitfish and trophic effects of baitfishing on the subsistence fishery. Fisheries Research, London 5, 263-286.

318

S. JENNINGS AND M.J. KAISER

Blaber, S.J.M., Milton, D.A., Smith, G.C. and Farmer, M.J. (1995). Trawl discards in the diets of tropical seabirds of the northern Great Barrier Reef, Australia. Marine Ecology Progress Series 127, 1-13. Blaber, S.J.M. and Wassenberg, T.J. (1989). Feeding ecology of the piscivorous birds Phalocrocorax varius, P . melanoleucos and Sterna bergii in Moreton Bay, Australia: diets and dependence on trawler discards. Marine Biology 101, 1-10. Blindheim, J. and Skjoldal, H.R. (1993). Effects of climatic changes on the biomass yield of the Barents Sea, Norwegian Sea and West Greenland large marine ecosystems. In “Large Marine Ecosystems: Stress, Mitigation and Sustainability” (K. Sherman, L.M. Alexander and B.D. Gold, eds), pp. 185-189. American Association for the Advancement of Science, Washington. Blythe, S.P. and Stokes, T.K. (1993). Size selective harvesting and age at maturity 1. Some theoretical implications for management of evolving resources. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 222-23 1. Springer-Verlag, Berlin. Bohnsack, J.A. (1982). Effects of piscivorous predator removal on coral reef fish community structure. In “Gutshop ’81: Fish Food Habits and Studies” (G.M. Caillet and C.A. Simenstad, eds), pp. 258-267. University of Washington, Seattle. Bohnsack, J.A. (1990). The potential of marine fishery reserves for reef fish management in the US.southern Atlantic. NOAA Technical Memorandum NMFS-SEFC 261.

Bohnsack, J.A. (1996). Maintenance and recovery of reef fishery productivity. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 283-313. Chapman and Hall, London. Borisov, V.M. (1978). The selective effect of fishing on the population structure of species with a long life cycle. Journal of Ichthyology 18, 896-904. Bouchon-Navaro, Y.,Bouchon, C. and Harmelin-Vivien, M. (1985). Impact of coral degradation on a chaetodontid fish assemblage (Moorea, French Polynesia). Proceedings of the Fifth International Coral Reef Symposium 5, 427432. Boudreau, P.R. and Dickie, L.M. (1992). Biomass spectra of aquatic ecosystems in relation to fisheries yield. Canadian Journal of Fisheries and Aquatic Sciences 49, 1528-1 538. Boudreau, P.R., Dickie, L.M. and Kerr, S.R. (1991). Body-size spectra of production and biomass as system-level indicators of ecological dynamics. Journal of Theoretical Biology 152, 329-339. Bowman, R.E. and Michaels, W.L. (1984). Food of seventeen species of northwest Atlantic fish. NOAA Technical Memorandum NMFS-NEFC 28. Bradstock, M. and Gordon, D.P. (1983). Coral-like bryozoan growths in Tasman Bay, and their protection to conserve local fish stocks. New Zealand Journal of Marine and Freshwater Research 17, 159-163. Brand, A.R., Allison, E.H. and Murphy, E.J. (1991). North Irish Sea scallop fisheries: a review of changes. In “An International Compendium of Scallop Biology and Culture” (S.E. Shumway and P.A. Sandifer, eds), pp. 204-218. World Aquaculture Society, Baton Rouge. Brander, K. (1981). Disappearance of common skate, Raia batis, from the Irish Sea. Nature, London, 290, 48-49. Breen, P.A. (1987). Mortality of Dungeness crabs caught by lost traps in the Fraser River Estuary, British Columbia. North American Journal of Fisheries Management 7, 429435.

THE EFFECTS OF FISHING

ON MARINE ECOSYSTEMS

319

Brierley, AS. and Watkins, J.L. (1996). Acoustic targets at South Georgia and the South Orkney Islands during a season of krill scarcity. Marine Ecology Progress Series 138, 51-61. Brierley, AS., Watkins, J.L. and Murray, A.W.A. (1997). Interannual variability in krill abundance at South Georgia. Marine Ecology Progress Series 150, 87-98. Briggs, R.P. (1992). An assessment of nets with a square mesh panel as a whiting conservation tool in the Irish Sea Nephrops fishery. Fisheries Research 13, 133-152. Britton, J.C. and Morton, B. (1994). Marine carrion and scavengers. Oceanography and Marine Biology: an Annual Review 32, 369-434. Brock, R.E., Lewis, C. and Wass, R.C. (1979). Stability and structure of a fish community on a coral patch reef in Hawaii. Marine Biology 54, 281-292. Brothers, G. (1992). Lost or abandoned fishing gear in the Newfoundland-aquatic environment. Report of the Symposium on Marine Stewardship in the Northwest Atlantic, Department of Fisheries and Oceans, St. Johns, Newfoundland, Canada. Brown, J.S. and Parman, A.O. (1993). Consequences of size-selective harvesting as an evolutionary game. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 248-261. Springer-Verlag, Berlin. Brown, K.M. and Alexander, J.E.J. (1994). Group foraging in a marine gastropod predator: benefits and costs to individuals. Marine Ecology Progress Series 112, 97-105. Brown, M.W., Allen, J. M. and Kraus, S. D. (1995). The designation of seasonal right whale conservation areas in the waters of Atlantic Canada. In “Marine Protected Areas and Sustainable Fisheries” (N.L. Shackell and J.H. MartinWillison, eds.), pp. 9C98. Science and Management of Protected Areas Association, Wolfville. Brown, R.G.B. and Nettleship, D.N. (1984). Capelin and seabirds in the north-west Atlantic. In “Marine Birds: Their Feeding Ecology and Commercial Fisheries Relationships” (D.N. Nettleship, G.A. Sanger and P.F. Springer, eds), pp. 184194. Canadian Wildlife Service, Ottawa. Brylinsky, M., Gibson, J. and Gordon, D.C. (1994). Impacts of flounder trawls on the intertidal habitat and community of the Minas Basin, Bay of Fundy. Canadian Journal of Fisheries and Aquatic Sciences 51, 650-661. Burger, A.E. and Cooper, J. (1984). The effects of fisheries on seabirds in South Africa and Namibia. In “Marine Birds: Their Feeding Ecology and Commercial Fisheries Interactions” (D.N. Nettleship, G.A. Sanger and P.F. Springer, eds), pp. 155160. Canadian Wildlife Service, Ottawa. Burger, A.E. and Piatt, J.F. (1990). Flexible time budgets in breeding common murres: buffers against variable prey availability. Studies in Avian Biology 14, 71-83. Buxton, C.D. (1993). Life-history changes in exploited reef fishes on the east coast of South Africa. Environmental Biology of Fishes 36, 47-63. Caddy, J.F. (1 968). Underwater observations on scallop (Placopecten rnagellanicus) behaviour and drag efficiency. Journal of the Fisheries Research Board of Canada 25, 2123-2141. Caddy, J.F. (1973). Underwater observations on tracks of dredges and trawls and some effects of dredging on a scallop ground. Journal of the Fisheries Research Board of Canada 30, 173-180. Caley, M.J. (1993). Predation, recruitment and the dynamics of communities of coral-reef fishes. Marine Biology 117, 33-43. Camphuysen, C.J., Ensor, K., Furness, R.W., Garthe, S., Huppop, O., Leaper, G., Offringa, H. and Tasker, M.L. (1993). Seabirds feeding on discards in winter in the

320

S. JENNINGS AND M.J. KAISER

North Sea. Reports of the Netherlands Institute for Sea Research, Den Burg, Texel. Camphuysen, C.J., Calvo, B., Durinck, J., Ensor, K., Follestad, A., Furness, R.W., Garthe, S., Leaper, G., Skov, H., Tasker, M.L. and Winter, C.J.N. (1995). Consumption of discards by seabirds in the North Sea. Reports of the Netherlands Institute for Sea Research, Den Burg, Texel. Carpenter, K.E. and Alcala, A.C. (1977). Philippine coral reef fisheries resources, Part 2, Muro-ami and kayakas reef fisheries, benefit or bane? Philippine Journal of Fisheries 15, 217-235. Carpenter, K.E., Miclat, R.I., Albaladego, V.D. and Corpuz, V.T. (1981). The influence of substrate structure on the local abundance and diversity of Philippine reef fishes. Proceedings of the Fourth International Coral Reef Symposium 2, 497-502. Carpenter, R.C. (1985). Sea urchin mass-mortality: effects on reef algal abundance, species composition and metabolism and other coral reef herbivores. Proceedings of the Fifth International Coral Reef Symposium 4, 53-60. Carpenter, R.C. (1986). Partitioning herbivory and its effects on coral reef algal communities. Ecological Monographs 56, 345-363. Carpenter, R.C. (1988a). Mass mortality of a Caribbean sea urchin: immediate effects on community metabolism and other herbivores. Proceedings of the U S National Academy of Sciences 85, 51 1-514. Carpenter, R.C. (1990). Mass mortality of Diadema antillarum: effects on population density of parrotfish and surgeonfish. Marine Biology 104, 79-86. Carpenter, S.R. (1988b). Transmission of variance through lake food webs. In “Complex Interactions in Lake Communities” (S.R. Carpenter, ed.), pp. 119135. Springer-Verlag, New York. Carpenter, S.R. and Kitchell, J.F. (1984). Planktonic community structure and limnetic primary production. The American Naturalist 124, 159-1 72. Carpenter, S.R. and Leavitt, P.R. (1991). Temporal variation in a paleolimnological record arising from a trophic cascade. Ecology 12, 277-285. Carpenter, S.R., Kitchell, J.F. and Hodgson, J.R. (1985). Cascading trophic interactions and lake productivity. Bioscience 35, 634-639. Carr, H.A., Blott, A.J. and Caruso, P.G. (1992). A study of ghost gillnets in the inshore waters of southern New England. In “MTS ’92: Global Ocean Partnership”, pp. 361-367. Marine Technology Society, Washington D.C. Carr, M.H. and Hixon, M.A. (1995). Predation effects on early post-settlement survivorship of coral-reef fishes. Marine Ecology Progress Series 124, 3 1-42. Carvalho, G.R. and Hauser, L. (1994). Molecular genetics and the stock concept in fisheries. Reviews in Fish Biology and Fisheries 4, 326-350. Chapdelaine, G. and Rail, J.-F. (1997). Relationship between cod fishery activities and the population of herring gulls on the North Shore of the Gulf of St Lawrence, QuCbec, Canada. ICES Journal of Marine Science 54, 708-713. Charnov, E.L. (1982). The theory of sex allocation. Monographs in Population Biology 18, 1-335. Charnov, E.L. (1993). “Life History Invariants: Some Explorations of Symmetry in Evolutionary Ecology”. Oxford University Press, Oxford. Choat, J.H. (1982). Fish feeding and the structure of benthic communities in temperate waters. Annual Review of Ecology and Systematics 13, 429-449. Christensen, I. (1990). A review of the distribution, migrations, food, reproduction, exploitation and present abundance of humpback whales (Megaptera novaengliae)

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

321

in the northeast Atlantic. International Council for the Exploration of the Sea, Committee Meeting 1990/N:10. Christensen, V. (1996). Managing fisheries involving predator and prey species. Reviews in Fish Biology and Fisheries 6, 417442. Christensen, V. and Pauly, D. (1992). ECOPATH 11: a software for balancing steady-state ecosystem models and calculating network characteristics. Ecological Modelling 61, 169-185. Christensen, V. and Pauly, D. (1993). “Trophic Models of Aquatic Ecosystems”. ICLARM, Manila. Churchill, J.H. (1989). The effect of commercial trawling on sediment resuspension and transport over the Middle Atlantic Bight continental shelf. Continental Sherf Research 9, 841-864. Clark, J.R., Causey, B. and Bohnsack, J.A. (1989). Benefits from coral reef protection: Looe Key Reef, Florida. In “Coastal Zone ’89” (O.T. Magoon, H. Converse, D. Miner, L.T. Tobin and D. Clark, eds), pp. 307C3086. American Society of Civil Engineers, New York. Clark, M.R. (1996). Biomass estimation of orange roughy: a summary and evaluation of techniques for measuring stock size of a deep-water fish species in New Zealand. Journal of Fish Biology 49 (Supplement A), 114-131. Clarke, K.R. (1993). Non-parametric multivariate analysis of changes in community structure. Australian Journal of Ecology 18, 117-143. Clarke, K.R. and Ainsworth, M. (1993). A method of linking multivariate community structure to environmental variables. Marine Ecology Progress Series 92, 205-2 19. Clarke, K.R. and Green, R.H. (1988). Statistical design and analysis for a ’biological effects’ study. Marine Ecology Progress Series 46, 213-226. Cohen, J.E. and Briand, F. (1984). Trophic links of community food webs. Proceedings of the National Academy of Science 81, 41054109. Cohen, J.E., Pimm, S.L., Yodzis, P. and Saldafia, J. (1993). Body sizes of animal predators and animal prey in food webs. Journal of Animal Ecology 62, 67-78. Cole, H.A. (1971). The heavy tickler chain - right or wrong? The view of Dr H.A. Cole. World Fishing 1011971, 8-10. Collie, J.S., Escanero, G.A. and Valentine, P.C. (1997). Effects of bottom fishing on the benthic megafauna of Georges Bank. Marine Ecology Progress Series 155, 159-172. Collins, S.L. and Benning, T.L. (1996). Spatial and temporal patterns in functional diversity. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed.), pp. 253-280. Blackwell Science, Oxford. Connell, J.H. (1978). Diversity in tropical rain forests and coral reefs. Science, N.Y., 199, 1302-1310. Connell, S.D. and Jones, D.P. (1991). The influence of habitat composition on post recruitment processes in a temperate reef fish population. Journal of Experimental Marine Biology and Ecology 151, 271-294. Cook, R.M., Sinclair, A. and Stefannson, G. (1997). Potential collapse of North Sea cod stocks. Nature 385, 521-522. Cooke, J.G. (1985). Has the age at sexual maturity of southern hemisphere minke whales declined? Report of the International Whaling Commission 32, 633-642. Corten, A. (1986). On the causes of the recruitment failure of herring in the central and northern North Sea in the years 1972-1978. Journal du Conseil, Conseil International pour I’Exploration de la Mer 42, 28 1-294.

322

S. JENNINGS AND M.J. KAISER

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. Couperus, A.S.A.F. (1994). Killer whales (Orcinus orca) scavenging on discards of freezer trawlers North East of the Shetland Islands. Aquatic Mammology 20, 4751. Crawford, R.J.M. and Shelton, P.A. (1978). Pelagic fish and seabird interrelationships off the coasts of southwest and South Africa. Biological Conservation 14, 85-1 09. Crawford, R.J.M., Shannon, L.V. and Pollock, D.E. (1987). The Benguela ecosystem: 4. The major fish and invertebrate resources. Oceanography and Marine Biology: an Annual Review 25, 353-505. Crawford, R.J.M., Underhill, L.G., Raubenheimer, C.M., Dyer, B.M. and Martin, J. (1992). Top predators in the Benguela ecosystem - implications of their trophic position. South African Journal of Marine Science 12, 675-687. Croxall J.P., McCann T.S., Prince P.A., Rothery P. (1988). Reproductive performance of seabirds and seals at South Georgia and Signy Island, South Orkney Islands, 1976-1987: implications for Southern Ocean monitoring studies. In “Antarctic Ocean and Resources Variability” (D. Sahrhage, ed.), pp 261-285. Springer-Verlag, Berlin. Cruetzberg, F., Duineveld, G.C.A. and van Noort, G.J. (1987). The effect of different numbers of tickler chains on beam trawl catches. Journal du Conseil International pour I’Exploration de la Mer 43, 159-168. Culley, M. (1971). “The Pilchard: Biology and Exploitation”. Pergamon Press, Oxford. Currie, D.R. and Parry, G.D. (1996). Effects of scallop dredging on a soft sediment community: a large-scale experimental study. Marine Ecology Progress Series 134, 131-1 50. Cushing, D.H. (1961). On the failure of the Plymouth herring fishery. Journal of the Marine Biological Association of the United Kingdom 41, 799-8 16. Cushing, D.H. (1968). “Fisheries Ecology: a Study in Population Dynamics”. University of Wisconsin Press, Madison. Cushing, D.H. (1975). “Marine Ecology and Fisheries”. Cambridge University Press, Cambridge. Cushing, D.H. (1982). “Climate and Fisheries”. Academic Press, London. Cushing, D.H. (1988a). The study of stock and recruitment. In “Fish Population Dynamics” (J.A. Gulland, ed.), pp. 105-128. Wiley, Chichester. Cushing, D.H. (1988b). “The Provident Sea”. Cambridge University Press, Cambridge. Daan, N. (1980). A review of replacement of depleted stocks by other species and the mechanisms underlying such replacement. Rapports et Procis- Verbaux des RPunions, Conseil International pour I’Exploration de la Mer 177, 405-421. Daan, N. (1987). Multispecies versus single-species assessment of North Sea fish stocks. Canadian Journal of Fisheries and Aquatic Sciences 44 (Supplement 2), 360-370. Daan, N. (1993). Simulation study of effects of closed areas to all fishing, with particular reference to the North Sea ecosystem. In “Large Marine Ecosystems: Stress, Mitigation and Sustainability” (K. Sherman, L.M. Alexander and B.D. Gold, eds), pp. 252-258. American Association for the Advancement of Science, Washington.

M E EFFECTS OF FISHING ON MARINE ECOSYSTEMS

323

Daly, K.L. (1990). Overwintering development, growth and feeding of larval Euphausia superba in the Antarctic marginal ice zone. Limnology and Oceanography 35, 1564-1 576. Dalzell, P. (1996). Catch rates, selectivity and yields of reef fishing. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 161-192. Chapman and Hall, London. Dalzell, P., Adams, T.J.H. and Polunin, N.V.C. (1996). Coastal fisheries in the Pacific Islands. Oceanography and Marine Biology: an Annual Review 34, 395-531. Dare, P.J., Key, D. and Connor, P.M. (1993). The efficiency of spring-loaded dredges used in the western English Channel fishery for scallops, Pecten maximus (L.). International Council for the Exploration of the Sea, Committee Meeting 1993/B:15,8 pp. Dauer, D.M. (1984). High resilience to disturbance of an estuarine polychaete community. Bulletin of Marine Science 34, 17C174. Davidson, N.C., Laffoley, D. d’A., Doody, J.P., Way, L.S., Gordon, J., Key, R., Drake, C.M., Pienkowski, M.W., Mitchell, R. and Duff, K.L. (1991). “Nature Conservation and Estuaries in Great Britain”. Nature Conservancy Council, Peterborough. Dayton, P.K. and Hessler, R.R. (1972). Role of biological disturbance in maintaining diversity in the deep sea. Deep-sea Research 19, 199-208. Dayton, P.K., Thrush, S.F., Agardy, M.T. and Hofman, R.J. (1995). Environmental effects of marine fishing. Aquatic Conservation: Marine and Freshwater Ecosystems 5, 205-232. de Boer, B.A. (1978). Factors influencing the distribution of the damselfish Chromis cyanea (Poey), Pomacentridae, on a reef at Curacao, Netherlands Antilles. Bulletin of Marine Science 28, 550-565. de Groot, S.J. and Lindeboom, H.J. (1994). Environmental impact of bottom gears on benthic fauna in relation to natural resources management and protection of the North Sea. Reports of the Netherlands Institute for Sea Research, Texel. De Martini, E.E., Ellis, D.M. and Honda, V.A. (1992). Comparisons of spiny lobster Panilirus marginatus fecundity, egg size and spawning frequency before and after exploitation. Fishery Bulletin 91, 1-7. de Veen, J.F. (1976). On changes in some biological parameters in the North Sea sole (Solea solea L.). Journal du Conseil, Conseil International pour I’Exploration de la Mer 37, 6&90. De Vries, J.G. and Pearcy, W.G. (1983). Fish debris in sediments of the upwelling area off central Peru: a late quaternary record. Deep Sea Research 28, 87-109. de Wolf, P. and Mulder, M. (1985). Spatial variation of macrobenthos in the southern North Sea. Estuaries 8,64A (abstract only). Dickie, L.M., Kerr, S.R. and Boudreau, P.R. (1987). Size-dependent processes underlying regularities in ecosystem structure. Ecological Monographs 57, 233250. Dickson, R.R., Meincke, J., Malmberg, S.A. and Lee, A.J. (1988). The great ’salinity anomaly’ in the northern North Atlantic 1968-1982. Progress in Oceanography 20, 103-1 5 1. Doherty, P.J. (1991). Spatial and temporal patterns in recruitment. In “The ecology of fishes on coral reefs” (P.F. Sale, ed.), pp. 261-293. Academic Press, San Diego. Doherty, P.J. and Fowler, T. (1994). An empirical test of recruitment limitation in a coral reef fish. Science, N.Y. 263,935-939.

324

S. JENNINGS AND M.J. KAISER

Done, T.J. (1987). Simulation of the effects of Acanthaster planci on the population structure of massive corals of the genus Porites: evidence of population resilience. Coral Reefs 6, 75-90. Done, T.J. (1988). Simulation of recovery of predisturbace size structure in populations of Porites corals damaged by crown of thorns starfish Acanthaster planci L. Marine Biology 100, 51-61. Done, T.J. (1992). Phase-shifts in coral-reef communities and their ecological significance. Hydrobiologia 247, 121-132. Done, T.J., Osborne, K. and Navin, K.F. (1988). Recovery of corals postAcanthaster: progress and prospects. Proceedings of the Sixth International Coral Reef Symposium 2, 137-142. Dragesund, 0. and Gjosaeter, J. (1988). The Barents Sea. In “Continental Shelves (Ecosystems of the World 27)” (H. Postma and J.J. Zijlstra, eds), pp. 339-361. Elsevier, Amsterdam. Duffy, D.C. (1983). Environmental uncertainty and commercial fishing - effects on Peruvian guano birds. Biological Conservation 26, 227-238. Dugan, J.E. and Davis, G.E. (1993). Applications of marine refugia to coastal fisheries management. Canadian Journal of Fisheries and Aquatic Sciences 50, 2029-2042. Duineveld, G.C.A., Kunitzer, A. and Heyman, R.P. (1987). Amphiura jiliformis (Ophiuroidea: Echinodermata) in the North Sea. Distribution, present and former abundance and size composition. Netherlands Journal of Sea Research 21, 317-329. Duplisea, D.E. and Kerr, S.R. (1995). Application of a biomass size spectrum model to demersal fish data from the Scotian shelf. Journal of Theoretical Biology 177, 263-269. Earle, M. (1996). Ecological interactions between cetaceans and fisheries. In “The Conservation of Whales and Dolphins” (M.P. Simmonds and J.D. Hutchinson, eds), pp. 167-204. John Wiley, Chichester. Edley, M.T. and Law, R. (1988). Evolution of life histories and yields in experimental populations of Daphnia magna. Biological Journal of the Linnean Society 34, 309-326. Eggers, D.M., Bartoo, N.W., Rickard, N.A., Nelson, R.E., Wissmar, R.C., Burgner, R.L. and Devol, A.H. (1978). The Lake Washington ecosystem: the perspective from the fish community and forage base. Journal of the Fisheries Research Board of Canada 35, 1553-1571. Ehrlich, P.R. and Daily, G.C. (1993). Population extinction and saving biodiversity. Ambio 22, 64-68. Ehrlich, P.R. and Wilson, E.O. (1991). Biodiversity studies: science and policy. Science 253, 758-762. Eisenbud, R. (1985). The pelagic driftnet. Salt Water Sportsman 5/1985, 65-72 . Eldredge, L.G. (1987). Poisons for fishing on coral reefs. In “Human Impacts on Coral Reefs: Facts and Recommendations” (B. Salvat, ed.), pp. 6146. Antenne Museum, French Polynesia. Eleftheriou, A. and Robertson, M.R. (1992). The effects of experimental scallop dredging on the fauna and physical environment of a shallow sandy community. Netherlands Journal of Sea Research 30, 289-299. Elner, R.W. and Vadas, R.L. (1990). Inference in ecology: the sea urchin phenomenon in the north western Atlantic. The American Naturalist 136, 108-125. Elton, C.S. (1927). “Animal Ecology”. Sidgwick and Jackson, London.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

325

Emerson, C.W. (1989). Wind stress limitation of benthic secondary production in shallow, soft-sediment communities. Marine Ecology Progress Series 53, 65-77. Endean, R. and Stablum, W. (1973). A study of some aspects of the crown-of-thorns starfish (Acanthaster planci) infestations of reefs of Australia’s Great Barrier Reef. Atoll Research Bulletin 167, 1-60. Endean, R., Cameron, A.M. and de Vantier, L.M. (1988). Acanthaster planci predation on massive corals: the myth of rapid recovery of devastated reefs. Proceedings of the Sixth International Coral Reef Symposium 2, 143-1 55. Eno, N.C., MacDonald, D. and Amos, S.C. (1996). A study on the effects of fish (Crustacea/Mollusc) traps on benthic habitats and species. Report to European Commission Directorate General XIV, Studies Contract 94/076. Faure, G. (1989). Degradation of coral reefs at Moorea Island (French Polynesia) by Acanthaster planci. Journal of Coastal Research 5, 295-305. Fenchel, T. (1996). Worm burrows and oxic microniches in marine sediments. 1. Spatial and temporal scales. Marine Biology 127, 289-295. Fenchel, T. and Finlay, B.J. (1995). “Ecology and Evolution in Anoxic Worlds”. Oxford University Press, Oxford. Fogarty, M.J., Cohen, E.B., Michaels, W.L. and Morse, W.W. (1991). Predation and the regulation of sand lance populations: an exploratory analysis. ICES Marine Science Symposia 193, 120-124. Fonseca, M.S., Thayer, G.W. and Chester, A.J. (1984). Impact of scallop harvesting on eelgrass (Zostera marina) meadows: implications for management. North American Journal of Fisheries Management 4, 286-293. Francis, R.C., Awbry, F.T., Goudey, C.L., Hall, M.A., King, D.M., Medina, H., Norris, K.S., Orbach, M.K., Payne, R.and Pikitch, E. (1992). Dolphins and the Tuna Industry. National Research Council, Washington. Frank, K.T. and Leggett, W.C. (1994). Fisheries ecology in the context of ecological and evolutionary theory. Annual Review of Ecology and Systematics 25, 401-422. Fry, B. (1988). Food web structure on Georges Bank from stable C, N and S isotopic compositions. Limnology and Oceanography 33, 1182-1 190. Fry, B. and Sherr, E.B. (1984). Delta l 3 C measurements as indicators of carbon flow in marine and freshwater ecosystems. Contributions in Marine Science 27, 1347. Furness, R.W. (1982). Competition between fisheries and seabird communities. Advances in Marine Biology 20, 225-327. Furness, R.W. (1992). Implications of changes in net mesh size, fishing effort and minimum landing size regulations in the North Sea for seabird populations. Joint Nature Conservaton Committee Report 133, Peterborough. Furness, R.W. (1996). A review of seabird responses to natural or fisheries-induced changes in food supply. In “Aquatic Predators and their Prey” (S.P.R. Greenstreet and M.L. Tasker, eds), pp. 168-173. Blackwell Scientific Publications, Oxford. Furness, R.W. and Barrett, R.T. (1991). Seabirds and fish declines. Research and Exploration 7, 82-95. Furness, R.W. and Camphuysen, C.J. (1997). Seabirds as monitors of the marine environment. ICES Journal of Marine Science 54, 726-737. Furness, R.W., Ensor, K. and Hudson, A.V. (1992). The use of fishery waste by gull populations around the British Isles. Ardea 80, 105-1 13. Furness, R.W., Hudson, A.V. and Ensor, K. (1988). Interactions between scavenging seabirds and commercial fisheries around the British Isles. In “Seabirds and Other Marine Vertebrates: Competition, Predation and Other Interactions” (J. Burger, ed.), pp. 24Ck268. Columbia University Press, New York.

326

S. JENNINGS AND M.J. KAISER

Gabriel, W. L. (1992). Persistence of demersal fish assemblages between Cape Hatteras and Nova Scotia, Northwest Atlantic. Journal of Northwest Atlantic Fishery Science, 14, 29-46. Gadgil, M. (1996). Managing biodiversity. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed.), pp. 345-366. Blackwell Science, Oxford. Galvez, R. and Sadorra, M.S.M. (1988). Blast fishing: a Philippine case study. Tropical Coastal Area Management 3, 9-10. Garstang, W. (1900). The impoverishment of the sea. Journal of the Marine Biological Association of the United Kingdom 6, 1-69. Garthe, S., Camphuysen, K.C.J. and Furness, R.W. (1996). Amounts of discards by commercial fisheries and their significance as food for seabirds in the North Sea. Marine Ecology Progress Series 136, 1-1 1. Gaspar, M.B., Richardson, C.A. and Monteiro, C.C. (1994). The effects of dredging on shell formation in the razor clam Ensis siliqua from Barrinha, southern Portugal. Journal of the Marine Biological Association of the United Kingdom 74, 927-938. Gaston, K.J. (1996). What is biodiversity. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed.), pp. 1-9. Blackwell Science, Oxford. Gilkinson, K., Paulin, M., Hurley, S. and Schwinghamer, P. (1998) Impacts of trawl door scouring on infaunal bivalves: results of a physical trawl doorldense sand interaction. Journal of Experimental Marine Biology and Ecology 224, 29 1-3 12. Gislason, H. (1994). Ecosystem effects of fishing activities in the North Sea. Marine Pollution Bulletin 29, 520-527. Glynn, P.W. (1973). Acanthaster: effect on coral reef growth in Panama. Science 180, 504506. Glynn, P.W., Wellington, G.M. and Birkeland, C. (1979). Coral reef growth in GalLpagos: limitation by sea urchins. Science 203, 4749. Gomez, E.D., Alcala, A.C. and San Diego, A.C. (1981). Status of Philippine coral reefs. Proceedings of the Fourth International Coral Reef Symposium 1, 275-282. Gomez, E.D., Alcala, A.C. and Yap, H.T. (1987). Other fishing methods destructive to coral. In “Human Impacts on Coral Reefs: Facts and Recommendations” (B. Salvat, ed.), pp. 67-75. Antenne Museum, French Polynesia. Goreau, T.F., Lang, J.C., Graham, E.H. and Goreau, P.D. (1972). Structure and ecology of the Saipan reefs in relation to predation by Acanthaster planci (L). Bulletin of Marine Science 22, 113-152. Goss-Custard, J.D., Caldow, R.W.G., Clarke, R.T., Durell, S.E.A. le V. dit and Sutherland, W.J. (1995a). Deriving population parameters from individual variations in foraging behaviour: I. Empirical game theory distribution model of oystercatchers Haematopus ostralegus feeding on mussels Mytilus edulis. Journal of Animal Ecology 64, 265-276. Goss-Custard, J.D., Caldow, R.W.G., Clarke, R.T. and West, A.D. (1995b). Deriving population parameters from individual variations in foraging behaviour: 11. Model tests and population parameters. Journal of Animal Ecology 64, 277-289. Graham, M. (1935). Modern theory of exploiting a fishery and application to North Sea trawling. Journal du Conseil International pour I’Exploration a2 la Mer 10, 264274. Graham, M. (1955). Effect of trawling on animals of the sea bed. Deep Sea Research 3 (Supplement), 1-16. Grassle, J.F. and Saunders, H.L. (1973). Life histories and the role of disturbance. Deep Sea Research 20, 643-659.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

327

Green, E.P., Mumby, P.J., Edwards, A.J. and Clark, C.D. (1996). A review of remote sensing for the assessment and management of tropical coastal resources. Coastal Management 24, 140. Greenstreet, S.P.R. and Hall, S.J. (1996). Fishing and ground-fish assemblage structure in the north-western North Sea: an analysis of long-term and spatial trends. Journal of Animal Ecology 65, 577-598. Greer-Walker, M. and Emerson, L. (1990). The seasonal migration of soles (Solea solea) through the Dover Strait. Netherlands Journal of Sea Research 25,417-422. Grigg, R.W. (1994). Effects of sewage discharge, fishing pressure and habitat complexity on coral ecosystems and reef fishes in Hawaii. Marine Ecology Progress Series 103, 25-34. Grigg, R.W., Polovina, J.J. and Atkinson, M.J. (1984). Model of a coral reef ecosystem 111. Resource limitation, community regulation, fisheries yield and resource management. Coral Reefs 3, 23-27. Grimes, C.B., Idelberger, C.F., Able, K.W. and Turner, S.C. (1988). The reproductive biology of tilefish Lopholatilus chamaeleonticeps Goode & Bean, from the United States mid-Atlantic Bight, and the effects of fishing on the breeding system. Fishery Bulletin 86, 745-162. Grosslein, M.D., Langton, R.W. and Sissenwine, M.P. (1980). Recent fluctuations in pelagic fish stocks of the northwest Atlantic, Georges Bank region, in relation to species interactions. Rapports et Procis- Verbaux des Rdunions, Conseil International pour I’Exploration de la Mer 177, 374-404. Guillen, J.E., Ramos, A.A., Martinez, L. and Sanchez Lizaso, J.L. (1994). Antitrawling reefs and the protection of Posidonia oceanica (L.) meadows in the western Mediterranean Sea: demands and aims. Bulletin of Marine Science 55, 645-6 50. Guillory, V. (1993). Ghost fishing by blue crab traps. North American Journal of Fisheries Management 13, 459466. Gulland, J.A. (1970). Food chain studies and some problems in world fisheries. In “Marine Food Chains” (J.H. Steele, ed.), pp. 296318. Oliver and Boyd, Edinburgh. Gulland, J.A. (1977). Fish population dynamics. John Wiley & Sons, London. Gulland, J.A. (1991). Under what conditions will multispecies models lead to better fisheries management? ICES Marine Science Symposia 193, 348-352. Hall, M.A. (1996). On bycatches. Reviews in Fish Biology and Fisheries 6, 319-352. Hall, S.J. (1994). Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology, an Annual Review 32. 179-239. Hall,’ S.J. and Harding, M.J.C. (1997). Physical disturbance and marine benthic communities: the effects of mechanical harvesting of cockles on non-target benthic infauna. Journal of Applied Ecology 34, 497-517. Hall, S. J., Basford, D. J., and Robertson, M. R. (1990). The impact of hydraulic dredging for razor clams Ensis sp. on an infaunal community. Netherlands Journal of Sea Research 27, 119-125. Hall, S.J., Raffaelli, D. and Thrush, S.F. (1994). Patchiness and disturbance in shallow water benthic assemblages. In “Aquatic Ecology” (P.S. Giller, A.G. Hildrew and D. Raffaelli, eds), pp. 333-376. Blackwell Science, Oxford. Hall, S.J., Robertson, M.R., Basford, D.J. and Heaney, S.D. (1993). The possible effects of fishing disturbance in the northern North Sea - an analysis of spatial patterns in community structure around a wreck. Netherlands Journal of Sea Research 31, 201-208.

328

S. JENNINGS AND M.J. KAISER

Hall-Spencer, J.M. (1995). Evaluation of the direct impact of fishing gears on the substratum and on the benthos. Report to European Commission, Brussels. PEM\93\08. Hamer, K.C., Furness, R.W. and Caldow, R.W.G. (1991). The effects of changes in food availability on the breeding ecology of Great Skuas Catharacta skua in Shetland. Journal of Zoology 223, 175-188. Hamer, K.C., Monaghan, P., Uttley, J.D., Walton, P. and Bums, M.D. (1993). The influence of food supply on the breeding ecology of kittiwakes Rissa tridactyla in Shetland. Ibis 135, 255-263. Hammond, P.S., Hall, A.J. and Prime, J.H. (1994a). The diet of grey seals around Orkney and other island and mainland sites in north-eastern Scotland. Journal of Applied Ecology 31, 34C350. Hammond, P.S., Hall, A.J. and Prime, J.H. (1994b). The diet of grey seals in the Inner and Outer Hebrides. Journal of Applied Ecology 31, 737-746. Hamre, J. (1988). Some aspects of the interrelation between the herring in the Norwegian Sea and stocks of capelin and cod in the Barents Sea. International Council for the Exploration of the Sea, Committee Meeting 19881 H42. Hamre, J. (1991). Interrelation between environmental changes and fluctuating fish populations in the Barents Sea. In “Long Term Variability of Pelagic Fish Populations and Their Environment” (T. Kawasaki, S. Tanaka, Y. Toba and A. Taniguchi, eds), pp. 259-270. Pergamon Press, Oxford. Hamre, J. (1994). Biodiversity and exploitation of the main fish stocks in the Norwegian-Barents Sea ecosystem. Biodiversity and Conservation 3, 473494. Hansen, D. (1997). Decline of harbor seals (Phoca vitulina) and northern sea lions (Eumetopiasjubatus) in the western Gulf of Alaska. In “Role of Forage Fishes in Marine Ecosystems”, pp. 197-208. Alaska Sea Grant College Program Report, University of Alaska, Fairbanks. Harden-Jones, F.R. and Scholes, P. (1974). The effect of door-to-door tickler chain on the catch-rate of plaice (Pleuronectes platessa L.) taken by an otter trawl. Journal du Conseil International pour l’exploration de la Mer 35, 21@-212. Hardy A.C., Gunther E.R. (1935). The plankton of the South Georgia whaling grounds and adjacent waters, 1926-1932. Discovery Reports 11, 1-456. Harmelin, J.G., Bachet, F. and Garcia, F. (1995). Mediterranean marine reserves fish indexes as tests of protection efficiency. Marine Ecology 16, 233-250. Harris, A.N. and Poiner, I.R. (1991). Changes in species composition of demersal fish fauna of southeast Gulf of Carpentaria, Australia, after 20 years of fishing. Marine Biology 111, 503-519. Harris, M.P. and Wanless, S. (1991). The importance of the lesser sandeel Ammodytes marinus in the diet of shag Phalacrocorax aristotelis. Ornis Scandanavia 22, 375-382. Hart, M.W. and Scheibling, R.E. (1988). Heat waves, baby booms, and the destruction of kelp beds by sea urchins. Marine Biology 99, 167-176. Harvey, P.H. and Pagel, M.D. (1991). “The Comparative Method in Evolutionary Biology”. Oxford University Press, Oxford. Harwood, J. and Croxall, J.P. (1988). The assessment of competition between seals and commercial fisheries in the North Sea and the Antarctic. Marine Mammal Science 4, 13-33. Haug, T., Kroyer, A.B., Nilssen, K.T., Ugland, K.I. and Aspholm, P.E. (1991). Harp seal (Phoca groenlandica) invasions of Norwegian coastal waters: age composition and feeding habits. ICES Journal of Marine Science 48, 363-371.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

329

Hay, M.E. (1984). Patterns of fish and urchin grazing on Caribbean coral reefs: are previous results typical? Ecology 65, 446-454. Hay, M.E. (1985). Spatial patterns of herbivore impact and their importance in maintaining algal species richness. Proceedings of the Fifth International Coral Reef Symposium 4, 29-34. Hay, M.E. and Taylor, P.R. (1985). Competition between herbivorous fishes and urchins on Caribbean reefs. Oecologia 65, 591-598. Hayes, D.L. and Kuletz, K.J. (1997). Decline of pigeon guillemot populations in Prince William Sound, Alaksa, and apparent changes in the distribution and abundance of their prey. In “Role of Forage Fishes in Marine Ecosystems”, pp. 699-702 Alaska Sea Grant College Program Report, University of Alaska, Fairbanks. Heessen, H.J.L. (1996). Time-series data for a selection of forty fish species caught during the International Bottom Trawl Survey. ICES Journal of Marine Science 53, 1079-1084. Heessen, H.J.L. and Daan, N. (1996). Long-term changes in ten non-target North Sea fish species. ICES Journal of Marine Science 53, 1063-1078. Hiatt, R.W. and Strasburg, D.W. (1960). Ecological relationships of the fish fauna on coral reefs of the Marshall Islands. Ecological Monographs 30, 65-127. High, W.L. (1976). Escape of Dungeness crabs from pots. Marine Fisheries Reviews 38, 19-23. Hilborn, R. and Walters, C.J. (1992). “Quantitative Fisheries Stock Assessment: Choice, Dynamics and Uncertainty”. Chapman and Hall, New York. Hill, B.J. and Wassenberg, T.J. (1990). Fate of discards from prawn trawlers in Torres Strait. Australian Journal of Marine and Freshwater Research 41, 53-64. Hislop, J.R.G. (1996). Changes in North Sea gadoid stocks. ICES Journal of’Marine Science 53, 11461 156. Hixon, M.A. (1991). Predation as a process structuring coral reef fish communities. In “The Ecology of Fishes on Coral Reefs” (P.F. Sale, ed.), pp. 475-508. Academic Press, San Diego. Hixon, M.A. and Beets, J.P. (1993). Predation, prey refuges and the structure of coral reef fish assembleges. Ecological Monographs 63, 77-101. Hjort, J. (1914). Fluctuations in the great fisheries of northern Europe viewed in the light of biological research. Rapports et Procis- Verbaux des RPunions, Conseil International pour I’Exploration de la Mer 20, 1-13. Hjort, J., Jahn, G . and Ottestad, P. (1933). The optimum catch. Hvalradets Skrifter 7, 92-107. Hobson, K.A. and Welch, H.E. (1992). Determination of trophic relationships within a high arctic marine food web using Sk3Cand d 5 N analysis. Marine Ecology Progress Series 84, 9-18. Hobson, K.A. and Welch, H.E. (1994). Cannabilism and trophic structure in a high Arctic lake: insights from stable-isotope analysis. Canadian Journal of Fisheries and Aquatic Sciences 52, 1195-1201. Hodgson, W.C. (1957). “The Herring and Its Fishery”. Routledge & Kegan Paul, London. Hoffman, S.G. and Robertson, D.R. (1983). Foraging and reproduction of two Caribbean reef toadfishes (Batrachoididae). Bulletin of Marine Science 33, 919926. Holden, M.J. (1978). Long-term changes in landings of fish from the North Sea. Rapports et Procis- Verbaux des RPunions, Conseil International pour I’Exploration de la Mer 172, 11-26.

330

S. JENNINGS AND M.J. KAISER

Holden, M.J. (1994). “The Common Fisheries Policy”. Blackwell Scientific Publications, Oxford. Holtmann, S.E., Belgers, J.J.M., Kracht, B. and Daan, R. (1996). The macrobenthic fauna in the Dutch sector of the North Sea in 1995 and a comparison with previous data. Reports of the Netherlands Institute for Sea Research, Texel. Horwood, J.W. (1987). “The Sei Whale: Population Biology, Ecology and Management”. Croom Helm, Beckenham. Horwood, J.W., Bannister, R.C.A. and Howlett, G.J. (1986). Comparative fecundity of North Sea plaice (Pleuronectes platessa L.). Proceedings of the Royal Society of London B 228, 401-43 1. Houghton, R.G. (1979). Density-dependent growth in demersal fish. International Councilfor the Exploration of the Sea, Committee Meeting 1979/6:22. Houghton, R.G. and Harding, D. (1976). The plaice of the English Channel: spawning and migration. Journal du Conseil International pour I’Exploration de la Mer 36,229-239. Hudson, A.V. and Furness, R.W. (1988). Utilization of discarded fish by scavenging seabirds behind whitefish trawlers in Shetland. Journal of Zoology 215, 151-166. Hughes, R.N. (1993). Introduction. In “Diet Selection: an Interdisciplinary Approach to Foraging Behaviour” (R.N. Hughes, ed). pp. 1-9. Blackwell Science, Oxford. Hughes, R.N. and Croy, M.I. (1993). An experimental analysis of frequency-dependent predation (switching) in the 15-spined stickleback, Spinachia spinachia. Journal of Animal Ecology 62, 341-352. Hughes, T.P. (1994). Catastrophes, phase shifts and large-scale degradation of a Caribbean coral reef. Science 265, 1547-1551. Hughes, T.P., Keller, B.D., Jackson, J.B.C. and Boyle, M.-J. (1987a). Massmortality of the echinoid Diadema antillarum Philippi in Jamaica. Bulletin of Marine Science 36, 377-384. Hughes, T.P., Reed, D.C. and Boyle, M.-J. (1987b). Herbivory on coral reefs: community structure following mass mortalities of sea urchins. Journal of Experimental Marine Biology and Ecology 113, 39-59. Huppert, D.D. (1991). Managing the groundfish fisheries of Alaska, history and prospects. Reviews in Aquatic Sciences 4, 339-373. Huston, M. (1985). Variation in coral growth rates with depth at Discovery Bay, Jamaica. Coral Reefs 4, 19-25. Huston, M.A. (1994). “Biological Diversity: the Coexistence of Species on Changing Landscapes”. Cambridge University Press, Cambridge. Hutchings, J.A. (1995). Seasonal marine protected areas within the context of spatiotemporal variation in the northern cod fishery. In “Marine Protected Areas and Sustainable Fisheries” (N.L. Shackell and J.H. Martin-Willison, eds.), pp. 39-47. Science and Management of Protected Areas Association, Wolfville. Hutchings, P. (1990). Review of the effects of trawling on macrobenthic epifaunal communities. Australian Journal of Marine and Freshwater Research 41, 111-120. Hutchings, P.A. (1986). Biological extinction of coral reefs. Coral Reefs 4, 239-252. Hutchinson, J. (1996). Fisheries interactions: the harbour porpoise - a review. In “The Conservation of Whales and Dolphins” (M.P. Simmonds and J.D. Hutchinson, eds), pp. 128-165. John Wiley, Chichester. Idyll, C.P. (1973). The anchovy crisis. Scientijic American 228, 22-29. Jarre-Teichmann, A. (1998). The potential role of mass balance models for the management of upwelling ecosystems. Ecological Applications 8, 93-103.

THE EFFECTS OF FISHING.ON MARINE ECOSYSTEMS

33 1

Jefferson, T.A. and Curry, B.E. (1994). A global review of porpoise (Cetacea: Phocoenidae) mortality in gillnets. Biological Conservation 67, 167-183. Jenkins, J.T. (1927). “The Herring and the Herring Fisheries”. King & Son, London. Jennings, S. (1992). Potential effects of estuarine development on the success of management strategies for the British bass fishery. Ambio 21, 468-470. Jennings, S. and Beverton, R.J.H. (1991). Intraspecific variation in the life history tactics of Atlantic herring (Clupea harengus L.) stocks. ICES Journal of Marine Science 48, 117-125. Jennings, S. and Lock, J.M. (1996). Population and ecosystem effects of fishing. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 193-218. Chapman and Hall, London. Jennings, S. and Polunin, N.V.C. (1995). Comparative size and composition of yield from six Fijian reef fisheries. Journal of Fish Biology 46, 2846. Jennings, S. and Polunin, N.V.C. (1996a). Effects of fishing effort and catch rate upon the structure and biomass of Fijian reef fish communities. Journal of Applied Ecology 33, 400-412. Jennings, S. and Polunin, N.V.C. (1996b). Impacts of fishing on tropical reef ecosystems. Ambio 25, 44-49. Jennings, S. and Polunin, N.V.C. (1997). Impacts of predator depletion by fishing on the biomass and diversity of non-target reef fish communities. Coral Reefs 16, 71-82. Jennings, S., Howlett, G . J. and Flatman, S. (1993). The distribution, migrations and stock integrity of lemon sole Microstomus kitt in the western English Channel. Fisheries Research, London 18, 377-388. Jennings, S., Grandcourt, E.M. and Polunin, N.V.C. (1995). The effects of fishing on the diversity, biomass and trophic structure of Seychelles’ reef fish communities. Coral Reefs 14, 225-235. Jennings, S., Boullk, D.B. and Polunin, N.V.C. (1996a). Habitat correlates of the distribution and biomass of Seychelles’ reef fishes. Environmental Biology of Fishes 46, 15-25. Jennings, S., Marshall, S.S. and Polunin, N.V.C. (1996b). Seychelles’ marine protected areas: comparative structure and status of reef fish communities. Biological Conservation 75, 201-209. Jennings, S., Reiiones, O., Morales-Nin, B., Polunin, N.V.C., Moranta, J. and Coll, J. (1997). Spatial variation in the I5N and I3C stable isotope composition of plants, invertebrates and fishes on Mediterranean reefs: implications for the study of trophic pathways. Marine Ecology Progress Series 146, 109-1 16. Jennings, S., Reynolds, J.D. and Mills, S.C. (1998). Life history correlates of responses to fisheries exploitation. Proceedings of the Royal Society B 265, 333-339. Johnson, K.H., Vogt, K.A., Clark, H.J., Schmitz, O.J. and Vogt, D.J. (1996). Biodiversity and the productivity and stability of ecosystems. Trends in Ecology and Evolution 11, 372-377. Jones, G.P. (1988). Experimental evaluation of the effects of habitat structure and competitive interactions on the juveniles of two coral reef fishes. Journal of Experimental Marine Biology and Ecology 123, 115-126. Jones, J.B. (1992). Environmental imact of trawling on the seabed: a review. New Zealand Journal of Marine and Freshwater Research 26, 59-67. Jones, R. (1973). Stock and recruitment with special reference to cod and haddock. Rapports et Proce‘s-verbaux des Re‘unions, Conseil Permanent International pour I’ Exploration de la mer 164, 156173.

332

S. JENNINGS AND M.J. KAISER

Jones, R. (1982). Ecosystems, food chains and fish yields. In “Theory and Management of Tropical Fisheries” (D. Pauly and G.I. Murphy, eds), pp. 195239. ICLARM, Manila. Kaiser, M.J. (1996). Starfish damage as an indicator of trawling intensity. Marine Ecology Progress Series 134, 303-307. Kaiser, M.J. and Ramsay, K. (1997). Opportunistic feeding by dabs within areas of trawl disturbance: possible implications for increased survival. Marine Ecology Progress Series 152, 307-3 10. Kaiser, M.J. and Spencer, B.E. (1994). Fish scavenging behaviour in recently trawled areas. Marine Ecology Progress Series 112, 41-49. Kaiser, M.J. and Spencer, B.E. (1996a). The behavioural response of scavengers to beam-trawl disturbance. In “Aquatic Predators and their Prey” (S.P.R. Greenstreet and M.L. Tasker, eds), pp. 117-123. Blackwell Scientific Publications, Oxford. Kaiser, M.J. and Spencer, B.E. (1996b). The effects of beam-trawl disturbance on infaunal communities in different habitats. Journal of Animal Ecology 65,348-358. Kaiser, M.J., Rogers, S.I. and McCandless, D.T. (1994). Improving quantitative surveys of epibenthic communities using a modified 2m beam trawl. Marine Ecology Progress Series 106, 131-138. Kaiser, M.J., Bullimore, B., Newman, P., Lock, K. and Gilbert, S. (1996a). Catches in ’ghost-fishing’ set nets. Marine Ecology Progress Series 145, 11-16. Kaiser, M.J., Hill, A.S., Ramsay, K., Spencer, B.E., Brand, A.R., Veale, L.O., Prudden, K., Rees, E.I.S., Munday, B.W., Ball, B. and Hawkins, S.J. (1996b). Benthic disturbance by fishing gear in the Irish Sea: a comparison of beam trawling and scallop dredging. Aquatic Conservation: Marine and Freshwater Ecosystems 6, 269-285. Kaiser, M.J., Edwards, D.B., Armstrong, P.J., Radford, K., Lough, N.E.L., Flatt, R.P. and Jones H.D. (1998). Changes in megafaunal benthic communities in different habitats after trawling disturbance. ICES Journal of Marine Science, in press. Kaufman, L.S. and Ebersole, J.P. (1984). Microtopography and the organisation of two assemblages of coral reef fishes in the West Indies. Journal of Experimental Marine Biology and Ecology 78, 253-268. Kawasaki, T., Tanaka, S., Toba, Y. and Taniguchi, A. (eds) (1991). “Long-Term Variability of Pelagic Fish Populations and Their Environment”. Pergamon Press, Oxford. Keesing, J.K. and Halford, A.R. (1992). Field measurement of survival rates of juvenile Acanthaster planci - techniques and preliminary results. Marine Ecology Progress Series 85, 107-1 14. Kerfoot, W.C. and Sih, A. (1987). “Predation: Direct and Indirect Effects on Aquatic Communities”. University Press of New England, Hanover. Kerr, S.R. (1974). Theory of size distribution in ecological communities. Journal of the Fisheries Research Board of Canada 31, 1859-1862. Kirkwood, G. P., Beddington, J.R. and Rossouw, J.A. (1994). Harvesting species of different lifespans. In “Large-Scale Ecology and Conservation Biology” (P.J. Edwards, R.M. May and N.R. Webb eds), pp. 199-227. Blackwell Scientific, Oxford. Klomp, N.I. and Furness, R.W. (1992). Nonbreeders as a buffer against environmental stress - declines in numbers of Great Skuas on Foula, Shetland, and prediction of future recruitment. Journal of Applied Ecology 29, 341-348.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

333

Knijn, R.J., Boon, T.W., Heessen, H.J.L. and Hislop, J.R.G. (1993). Atlas of North Sea fishes. ICES Co-operative Research Report 194, 268 pp. Knowlton, N. (1992). Thresholds and multiple stable states in coral-reef community dynamics. American Zoologist 32, 674-682. Kock, K.H. (1994). Fishing and conservation in southern waters. Polar Record 30, 3-22. Kock, K.H. and Shimadzu, Y. (1994). Trophic relationships and trends in population size and reproductive parameters in antarctic high-level predators. In “Southern Ocean Ecology: the BIOMASS Perspective” (S.Z. El-Sayed, ed.), pp. 287-3 12. Cambridge University Press, Cambridge. Koslow, J.A., Hanley, F. and Wicklaud, R. (1988). Effects of fishing on reef fish communities at Pedro Bank and Port Royal Cays, Jamaica. Marine Ecology Progress Series 43, 201-212. Koslow, J.A., Bell, J., Virtue, P. and Smith, D.C. (1995). Fecundity and its variability in orange roughy - effects of population density, condition, egg size and senescence. Journal of Fish Biology 47, 1063-1080. Krost, P., Bernhard, M., Werner, F. and Hukriede, W. (1990). Otter trawl tracks in Kiel Bay (Western Baltic) mapped by side-scan sonar. Meeresforschung 32, 344-3 53. Kruse, G.H. and Kimber, A. (1993). “Degradable escape mechanisms for pot gear: a summary report to the Alaska Board of Fisheries”. Alaska Department of Fisheries and Game, Juneau. Kuletz, K.J., Irons, D., Piatt, J.F. and Duffy, D.C. (1997). Long-term changes in populations and diets of piscivorous birds and mammals in Prince William Sound, Alaska, reflect a shift in prey species abundance. In “Role of Forage Fishes in Marine Ecosystems”, pp. 703-706. Alaska Sea Grant College Program Report, University of Alaska, Fairbanks. Kunin, W.E. and Lawton, J.H. (1996). Does biodiversity matter? Evaluating the case for conserving species. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed.), pp. 283-308. Blackwell Science, Oxford. Langton, R.W. and Robinson, W.E. (1990). Faunal associations on scallop grounds in the western Gulf of Maine. Journal of Experimental Marine Biology and Ecology 144, 157-171. Langton, R.W., Stenech, R.S., Gotceitas, V., Juanes, F. and Lawton, P. (1996). The interface between fisheries research and habitat management. North American Journal of Fisheries Management 16, 1-7. Larkin, P.A. and Gazey, W. (1982). Applications of ecological simulation models to management of tropical multispecies fisheries. ICLARM Conference Proceedings 9, 123-140. Law, R. (1991). On the quantitative genetics of correlated characters under directional selection in age structured populations. Philosophical Transactions of the Royal Society B 331, 213-223. Law, R. and Grey, D.R. (1989). Evolution of yields from populations with agespecific cropping. Evolutionary Ecology 3, 343-359. Law, R. and Rowell, C.A. (1993). Cohort structured populations, selection responses, and exploitation of the North Sea cod. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 155173. Springer-Verlag, Berlin. Lawrence, J.M. (1975). On the relationships between marine plants and sea urchins. Oceanography and Marine Biology: an Annual Review 13, 213-286.

334

S. JENNINGS AND M.J. KAISER

Laws, R.M. (1977). Seals and whales of the Southern Ocean. Philosophical Transactions of the Royal Society B 279, 81-96. Lawton, J.H. (1989). Food Webs. In “Ecological Concepts: the Contribution of Ecology to an Understanding of the Natural World” (J.M. Cherrett, ed.), pp. 43-78. Blackwell Scientific Publications, London. Lawton, J.H. (1994). What do species do in ecosystems? Oikos 71, 1-8. Lawton, J.H. and Brown, V.K. (1993). Redundancy in ecosystems. In “Biodiversity and Ecosystem Function” (E.D. Schulze and H.A. Mooney, eds), pp. 255-270. Springer-Verlag, Berlin. Leaman, B.M. (1991). Reproductive styles and life history variables relative to exploitation and management of Sebastes stocks. Environmental Biology of Fishes 30, 253-27 1. LeCren, E.D., Kipling, C. and McCormack, J. (1977). A study of the numbers, biomass and yearclass strengths of perch (Perca Jluviatilis L.) in Windermere from 1941 to 1966. Journal of Animal Ecology 46, 281-307. Leggett, W. C. and Carscadden, J. F. (1978). Latitudinal variation in the reproductive characteristics of American shad Alosa sapidissima: evidence for population specific life history strategies in fish. Journal of the Fisheries Research Board of Canada 35, 1469-1478. Leggett, W.C., Frank, K.T. and Carscadden, J.E. (1984). Meteorological and hydrographic regulation of year-class strength in the capelin (Mallotus villosus) . Canadian Journal of Fisheries and Aquatic Sciences 41, 1193-1201. Leibold, M.A. (1996). A graphical model of keystone predators in food webs: trophic regulation of abundance, incidence and diversity patterns in communities. The American Naturalist 147, 784-812. Lessios, H.A. (1988). Mass mortality of Diadema antillarum in the Caribbean: what have we learned? Annual Reviews of Ecology and Systematics, 19, 371-393. Lessios, H.A., Cubit, J.D., Robertson, D.R., Shulman, M.J., Parker, M.R., Garrity, S.D. and Levings, S.C. (1984). Mass mortality of Diadema antillarum on the Caribbean coast of Panama. Coral Reefs 3, 173-182. Lessios, H.A., Robertson, D.R. and Cubit, J.D. (1985). Spread of Diadema mass mortality through the Caribbean. Science, N. Y., 226, 335-337. Lett, P.F. and Kohler, A.C. (1976). Recruitment: a problem of multispecies interaction and environmental perturbations with special reference to Gulf of St. Lawrence Atlantic herring. Canadian Journal of Fisheries and Aquatic Sciences 33, 1353-1371. Levin, L.A. (1984). Life history and dispersal patterns in a dense infaunal polychaete assemblage: community structure and response to disturbance. Ecology 65, 1185-1200. Levitan, D.R. (1992). Community structure in times past: influence of human fishing pressure on algal-urchin interactions. Ecology 73, 1597-1 605. Lewis, S.M. (1986). The role of herbivorous fishes in the organisation of a Caribbean reef community. Ecological Monographs 56, 183-200. Lewis, S.M. and Wainwright, P.C. (1985). Herbivore abundance and grazing intensity on a Caribbean coral reef. Journal of Experimental Marine Biology and ECOlOgy 87, 215-228. Lilly, G.R. (1991). Interannual variability in predation by cod (Gudus morhua) on capelin (Mallotus villosus) and other prey off southern Labrador and northeastern Newfoundland. ICES Marine Science Symposia 193, 133-1 46.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

335

Lindeboom, H.J. and de Groot, S.J. (1998). The effect of different types of fisheries on the North Sea and Irish Sea benthic ecosystems. Netherlands Institute of Sea Research, Report 1998-1. Den Burg, Texel. Lindeboom, H.J., Raaphorst, W.V., Beukema, J.J., Cadee, G. and Swennen, C. (1995). Sudden changes in the North Sea and Wadden Sea: Oceanic influences underestimated? In “Actual Problems of the Marine Environment. Lectures of the 4th International Scientific Symposium” (Anon. ed.), pp. 87-100 Hamburg, Germany. Lindley, J.A., Gamble, J.C. and Hunt, H.G. (1995). A change in the zooplankton of the central North Sea (55” to 58”N): a possible consequence of changes in the benthos. Marine Ecology Progress Series 119,299-303. Lloyd, C.S., Tasker, M.L. and Partridge, K. (1991). “The Status of Seabirds in Britain and Ireland”. T. & A.D. Poyser, London. Lo, N.C.H. and Smith, T.D. (1986). Incidental mortality of dolphins in the eastern tropical Pacific. Fishery Bulletin 84, 27-34. Lockwood, S.J. (1988). “The Mackerel”. Fishing News Books, Farnham. Lockwood, S.J., Pawson, M.G. and Eaton, D.R. (1983). The effects of crowding on mackerel (Scomber scombrus L.) - physical condition and mortality. Fisheries Research, London 2, 129-147. Lotka, A.J. (1925). “Elements of Physical Biology”. Williams & Wilkins, Baltimore. Loya, Y. (1976). Recolonization of Red Sea corals affected by natural catastrophes and man-made perturbations. Ecology 57, 278-289. Luckhurst, B.E. and Luckhurst, K. (1978). Analysis of the influence of substrate variables on coral reef fish communities. Marine Biology 49,317-323. Lyster, S. (1985). “International Wildlife Law”. Grotius Publications, Cambridge. MacArthur, R.H. (1955). Fluctuations of animal populations and a measure of community stability. Ecology 36, 533-536. MacCall, A.D. (1986). Changes in the biomass of the California current system. In “Variability and Management of Large Marine Ecosystems” (K. Sherman and L.M. Alexander, eds), pp. 33-54. Westview Press, Boulder. MacDonald, D.S., Little, M., Eno, N.C. and Hiscock, K. (1996). Towards assessing the sensitivity of benthic species and biotopes in relation to fishing activities. Aquatic Conservation: Marine and Freshwater Ecosystems 6,257-268. Macintyre, R.G., Glynn, P.W. and Cortts, J. (1992). Holocene reef history in the eastern Pacific: mainland Costa Rica, Can0 Island, COCOS Island and Galapagos Islands. Proceedings of the Seventh International Coral Reef Symposium 2, 1 1 7 4 1184 Mackintosh, N.A. (1972). Life cycle of Antarctic krill in relation to ice and water conditions. Discovery Reports, 36, 1-94. Magnusson, K.G. and Palsson, O.K. (1991). Predator-prey interactions of cod and capelin in Icelendic waters. ICES Marine Science Symposia 193, 153-170. Magurran, A.E. (1988). “Ecological Diversity and Its Measurement”. Croom Helm, London. Main, J. and Sangster, G.I. (1981). A study of sand clouds produced by trawl boards and their possible effect on fish capture. Department of Agriculture and Fisheries for Scotland, Aberdeen, Scottish Fisheries Research Report 20. Majluf, P. (1989). Reproductive ecology of South American fur seals in Peru. ICLARM Conference Proceedings 18, 332-343. Malthus, T.R. (1798). “An Essay on the Principle of Population as it Affects the Future Improvement of Society, with Remarks on the Speculations of Mr Godwin, M. Condoret and Other Writers”. Johnson, London.

336

S. JENNINGS AND M.J. KAISER

Mann, K.H. (1982). “Ecology of Coastal Waters: a Systems Approach”. Blackwell Scientific Publications, Oxford. Mann, K.H. and Breen, P.A. (1972). The relation between lobster abundance, sea urchins and kelp beds. Journal of the Fisheries Research Board of Canada 29,603605. Marten, G.G. (1979). Predator removal: effects on fisheries yields in Lake Victoria (East Africa). Science 203, 646-648. Martinez, N.D. (1993a). Effect of scale on food web structures. Science 260,242-243. Martinez, N.D. (1993b). Effects of resolution on food-web structure. Oikos 66, 403-412. Martinez, N.D. (1994). Scale-dependent constraints on food web structure. The American Naturalist 144, 935-953. Martinez, N.D. (1996). Defining and measuring functional aspects of biodiversity. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed.), pp. 114148. Blackwell Science, Oxford. Martins, E.P. (ed.) (1996) “Phylogenies and the Comparative Method in Animal Behaviour”. Oxford University Press, Oxford. Matishov, G.G. and Pavlova, L.G. (1994). Degradation of ecosystems of the north European seas under the effects of fishing and pathways of their recovery. Izvestiya Akademii Nauk Seriya Biologicheskaya 199411, 1 19-126. May, R.M. (1972). Will a large complex system be stable? Nature, London, 238,413414. May, R.M. (1973). “Stability and Complexity in Model Ecosystems”. Princeton University Press, New Jersey. May, R.M. (ed) (1984). “Exploitation of Marine Communities”. Springer-Verlag, Berlin. Mayo, R.K., Fogarty, M.J. and Serchuk, F.M. (1992). Aggregate fish biomass and yield on Georges Bank, 1960-87. Journal of Northwest Atlantic Fishery Science, 14, 59-78. McAllister, D.E. (1988). Environmental, economic and social costs of coral reef destruction in the Philippines. Galaxea 7 , 161-178. McClanahan, T.R. (1989). Kenyan coral reef-associated gastropod fauna: a comparison between protected and unprotected reefs. Marine Ecology Progress Series 53, 1-20. McClanahan, T.R. (1 990). Hierarchical control of coral reef ecosystems. PhD Thesis, University of Florida, 218 pp. McClanahan, T.R. (1992). Resource utilization, competition and predation: a model and example from coral reef grazers. Ecological Modelling 61, 195-215. McClanahan, T.R. (1994a). Kenyan coral reef lagoon fish: effects of fishing, substrate complexity, and sea urchins. Coral Reefs 13, 231-241. McClanahan, T.R. (1994b). Coral-eating snail Drupella cornus population increases in Kenyan coral reef lagoons. Marine Ecology Progress Series 115, 131-137. McClanahan, T.R. (1995a). A coral-reef ecosystem-fisheries model - impacts of fishing intensity and catch selection on reef structure and processes. Ecological Modelling 80, 1-19. McClanahan, T.R. (1995b). Fish predators and scavengers of the sea urchin Echinometra mathaei in Kenyan coral-reef marine parks. Environmental Biology of Fishes 43, 187-193. McClanahan, T.R. and Kaundaarara, B. (1996). Fishery recovery in a coral reef marine park and its effect on the adjacent fishery. Conservation Biology 10, 1187-1 199.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

337

McClanahan, T.R. and Muthiga, N.A. (1988). Changes in Kenyan coral reef community structure and function due to exploitation. Hydrobiologia 166, 269-276. McClanahan, T.R. and Muthiga, N.A. (1989). Patterns of predation on a sea urchin Echinometra mathaei (de Blainville) on Kenyan coral reefs. Journal of Experimental Marine Biology and Ecology 126, 77-94. McClanahan, T.R. and Obura, D. (1995). Status of Kenyan coral reefs. Coastal Management 23, 57-76. McClanahan, T.R. and Shafir, S.H. (1990). Causes and consequences of sea urchin abundance and diversity in Kenyan coral reef lagoons. Oecologia 83, 362-370. McClanahan, T.R., Kakamura, A.T., Muthiga, N.A., Gilagabher Yebio, M. and Obura, D. (1996). Effects of sea-urchin reductions on algae, coral and fish populations. Conservation Biology 10, 136154. McKenna, J.E. and Saila, S.B. (1991). Shifts in the antarctic demersal fish community of South Georgia Island. Fisheries Research 12, 109-124. McKenzie, W.D., Crews, D., Kallman, K.D., Policansky, D. and Sohn, J.J. (1983). Age, weight and the genetics of sexual maturation in the platyfish, Xiphophorus maculatus. Copeia 1983, 77&774. McManus, J.W. (1996). Social and economic aspects of reef fisheries and their management. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 24928 1 . Chapman and Hall, London. Mehl, S. (1986). Stomach contents of north-east Arctic cod and possible changes in their diet. International Councilfor the Exploration of the Sea, Committee Meeting 1986/G29.

Mehl, S. (1987). The northeast Arctic cod stocks consumption of commercially exploited prey species in 1984-1986. International Council for the Exploration of the Sea, Committee Meeting 1981/S:9. Mehl, S. and Sunnani, K. (1991). Changes in the growth of northeast Arctic cod in relation to food consumption in 1984-1988. ICES Marine Science Symposia 193, 109-1 12. Meyer, T., Cooper, R.A. and Pecci, K.J. (1981). The performance and environmental effects of a hydraulic clam dredge. Marine Fisheries Review 43, 14-22. Millar, R.B., Fahrig, L. and Shelton, P.A. (1990). Effect of capelin biomass on cod growth. International Council for the Exploration of the Sea, Committee Meeting 1990/G:25.

Miller, R.J. (1977). Resource underutilization in a spider crab industry. Fisheries 2, 9-13. Miller, R.J. (1985). Seaweeds, sea urchins and lobsters: a reappraisal. Canadian Journal of Fisheries and Aquatic Sciences 42, 2061-2072. Miller, R.R. (1957). Have the genetic patterns of fishes been altered by introductions or selective fishing. Journal of the Fisheries Research Board of Canada 14, 797-806. Millner, R.S. (1985). The use of anchored gill and tangle nets in the sea fisheries of England and Wales. Laboratory Leaflets, MAFF, Directorate of Fisheries Research, Lowestoft. Millner, R.S. and Whiting, C.L. (1996). Long-term changes in growth and population abundance of sole in the North Sea from 1940 to the present. ICES Journal of Marine Science 53, 1185-1 195. Monaghan, P. (1992). Seabirds and sandeels - the conflict between exploitation and conservation in the northern North Sea. Biodiversity and Conservation 1, 98-1 11. Monaghan, P., Uttley, J.D. and Okill, J.D. (1989). Terns and sandeels - seabirds as indicators of changes in marine fish populations. Journal of Fish Biology 35, 339-340.

338

S. JENNINGS AND M.J. KAISER

Monaghan, P., Uttley, J.D. and Burns, M.D. (1992). The effects of changes in food availability on reproductive effort in arctic terns Sternaparadisaea. Ardea 80,7 1-8 1. Montevecchi, W.A. and Myers, R.A. (1995). Prey harvests of seabirds reflect pelagic fish and squid abundance on multiple spatial and temporal scales. Marine Ecology Progress Series 117, 1-9. Montevecchi, W.A. and Myers, R.A. (1996). Dietary changes of seabirds indicate shifts in pelagic food webs. Sarsia 80, 313-322. 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. Moore, P.G. and Howarth, J. (1997). Foraging by marine scavengers: effects of relatedness, bait damage and hunger. Journal of Sea Research 36, 267-273. Moran, P.J. (1986). The Acanthaster phenomenon. Oceanography and Marine Biology: an Annual Review 24, 379480. Moran, P.J. (1990). Acanthaster planci (L.): biographical data. Coral Reefs 9,95-96. Muck, P. (1989). Major trends in the pelagic ecosystem off Peru and their implications for management. ICLARM Conference Proceedings 18, 386403. Munro, J.L. and Williams, D.M. (1985). Assessment and management of coral reef fisheries: biological, environmental and socioeconomic aspects. Proceedings of the Fourth International Coral Reef Symposium 4, 545-581. Munro, J.L., Parrish, J.D. and Talbot, F.H. (1987). The biological effects of intensive fishing upon reef fish communities. In “Human Impacts on Coral Reefs: Facts and Recommendations” (B. Salvat, ed.), pp. 41-49. Antenne Museum E.P.H.E., French Polynesia. Murphy, E.C., Springer, A.M. and Roseneau, D.G. (1991). High annual variability in reproductive success of kittiwakes (Rissa tridactyla L.) at a colony in western Australia. Journal of Animal Ecology 60,515-534. Myers, R.A., Barrowman, N.J., Hutchings, J.A. and Rosenberg, A.A. (1995). Population dynamics of exploited fish stocks at low population levels. Science, N.Y., 269, 1106-1108. Myers, R.A., Hutchings, J.A. and Barrowman, N.J. (1996). Hypothesis for the decline of cod in the North Atlantic. Marine Ecology Progress Series 138,293-308. Naeem, S., Thompson, L.J., Lawler, S.P., Lawton, J.H. and Woodfin, R.M. (1994). Declining biodiversity can alter the performance of ecosystems. Nature, London, 368, 734-737. Naeem, S., Thompson, L.J., Lawler, S.P., Lawton, J.H. and Woodfin, R.M. (1995). Empirical evidence that declining biodiversity may alter the performance of terrestrial ecosystems. Philosophical Transactions of the Royal Society B 347, 249-262. Neill, W.M. (1994). Spatial and temporal scaling and the organisation of limnetic communities.In “Aquatic Ecology: Scale, Pattern and Process” (P.S. Giller, A.G. Hildrew and D.G. Rafaelli, eds), pp. 189-231. Blackwell Scientific Publications, Oxford. Nelson, K. and SoulB, M. (1987). Genetical conservation of exploited fishes. In “Population Genetics and Fishery Management” (N. Ryman and F. Utter, eds), pp. 345-368. University of Washington Press, Seattle. Newman, G.G. and Crawford, R.J.M. (1980). Population biology and management of mixed species pelagic stocks off South Africa. Rapports et Proc& Verbaux des Rbnions, Conseil International pour I’Exploration de la Mer 177, 219-29 1. Nickell, L.A. and Atkinson, R.J.A. (1995). Functional morphology of burrows and trophic modes of three thalassinidean shrimp species, and a new approach to the classification of thalassinidean burrow morphology. Marine Ecology Progress Series 128, 181-197.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

339

Nickell, T.D. and Moore, P.G. (1992). The behavioural ecology of epibenthic scavenging invertebrates in the Clyde Sea area: laboratory experiments on attractions to bait in moving water, underwater TV observations in situ and general conclusions. Journal of Experimental Marine Biology and Ecology 159, 15-35. Nikolskii, G.V. (1969). “Theory of Fish Population Dynamics as the Background for Rational Exploitation and Management of Fishery Resources”. Oliver and Boyd, Edinburgh. Nilsson, C. and Grelsson, G. (1995). The fragility of ecosystems: a review. Journal of Applied Ecology 32, 671-692. Nishihira, M. and Yanmazato, K. (1974). Human interference with the coral reef community and Acanthaster infestation in Okinawa. Proceedings of the Second International Coral Reef Symposium 1, 577-590. Norris, J.E. and Parrish, J.D. (1988). Predator prey relationships among fishes in pristine coral reef communities. Proceedings of the Sixth International Coral Reef Symposium 2, 107-1 13. Olaso, I., Velasco, F., Pereda, P. and PCrez, N. (1996). Importance of blue whiting (Micromesistius poutassou) discarded in the diet of lesser-spotted dogfish (Scyliorhinus canicula) in the Cantabrian Sea. International Council for the Exploration of the Sea CM 19961Mini: 2. Oliver, R.S. and Slattery, P.N. (1985). Destruction and opportunity on the sea floor: effects of gray whale feeding. Ecology 66, 1965-1975. Ormond, R.F.G., Bradbury, R., Bainbridge, S., Fabricus, K., Keesing, J., De Vantier, L., Medley, P. and Steven, A. (1991). Test of a model of regulation of Crown-of-Thorns starfish by fish predators. In “Acanthaster and the Coral Reef: a Theoretical Perspective” (R. Bradbury, ed.). Springer-Verlag, Berlin. Oro, D. and Ruiz, X. (1997). Exploitation of trawler discards by breeding seabirds in the north-western Mediterranean: differences between the Ebro Delta and the Balearic Islands areas. ICES Journal of Marine Science 54, 695-707. Oro, D., Genovart, X., Ruiz, X., Jimenez, J. and Garcia-Gans, J. (1996). Differences in diet, population size and reproductive performance between two colonies of Audouin’s Gull Larus audouinii affected by a trawling moratorium. Journal of Avian Biology 21, 245-251. Overholtz, W.J. and Tyler, A.V. (1985). Long term responses of the demersal fish assemblages of Georges Bank. Fishery Bulletin, US.,83, 507-520. Overholtz, W.J., Marawski, S.A. and Foster, K.L. (1991). Impact of predatory fish, marine mammals and seabirds on the pelagic fish ecosystem of the northeastern USA. ICES Marine Science Symposia 193, 198-208. Owens, N.J.P. (1987). Natural variations in ‘’N in the marine environment. Advances in Marine Biology 24, 389-451. Paine, R.T. (1992). Food web analysis through measurement of per capita interaction strength. Nature 355, 73-75. Parrish, B.B. and Saville, A. (1965). The biology of north-east Atlantic herring populations. Oceanography and Marine Biology, an Annual Review 3, 323-373. Parrish, F.A. and Kazama, T.K. (1992). Evaluation of ghost fishing in the Hawaiian lobster fishery. Fisheries Bulletin 90, 72@725. Parrish, J.D., Callahan, M.W. and Norris, J.E. (1985). Fish trophic relationships that structure reef communities. Proceedings of the Ffth International Coral Reef Symposium 4, 73-18. Parrish, J.D., Norris, J.E., Callahan, M.W., Magarifugi, E.J. and Schroeder, R.E. (1986). Piscivory in a coral reef community. In “Gutshop ’81: Fish Food Habits

340

S. JENNINGS AND M.J. KAISER

and Studies” (G.M. Caillet and C.A. Simenstad, eds), pp. 73-78. University of Washington, Seattle. Patton, M.L., Grove, R.S. and Harman, R.F. (1985). What do natural reefs tell us about designing artificial reefs in southern California? Bulletin of Marine Science 37,279-298. Pauly, D. (1979). Theory and management of tropical multispecies stocks: a review with emphasis on the southeast Asian demersal fisheries. ICLARM Studies and Reviews 1, 1-35. Pauly, D. (1988). Some definitions of overfishing relevant to coastal zone management in Southeast Asia. Tropical Coastal Area Management 3, 14-15. Pauly, D. (1997). Putting fishery management back in places. Reviews in Fish Biology and Fisheries 7, 125-127. Pauly, D. and Christensen, V. (1995). Primary production required to sustain global fisheries. Nature 374, 255-257. Pauly, D., Silvestre, G. and Smith, I.R. (1989). On development, fisheries and dynamite: a brief review of tropical fisheries management. Natural Resource Modeling 3, 307-329. Pawson, M.G. and Jennings, S. (1996). A critique of methods for stock identification in marine capture fisheries. Fisheries Research 25, 203-217. Pawson, M.G., Kelley, D.F., and Pickett, G.D. (1987). The distribution and migrations of bass Dicentrarchus labrax (L.) in waters around England and Wales as shown by tagging. Journal of the Marine Biological Association of the United Kingdom, 67, 183-217. Payne, M.R. (1977). Growth of a fur seal population. Philosophical Transactions of the Royal Society B 279, 67-79. Payne, P.M., Wiley, D.N., Young, S.B., Pittman, S., Clapham, P.J. and Jossi, J.W. (1990). Recent fluctuations in the abundance of baleen whales in the southern Gulf of Maine in relation to their selected prey. Fishery Bulletin, U.S., 88, 687496. Pearson, R.G. (1981). Recovery and recolonisation of coral reefs. Marine Ecology Progress Series 4, 105-122. Pearson, T.H., Josefson, A.B. and Rosenberg, R. (1985). Petersen’s stations revisited. 1. Is the Kattegatt becoming eutrophic? Journal of Experimental Marine Biology and Ecology 92, 157-206. Perrin, W.F., Donovan, G. and Barlow, J. (1994). “Gillnets and Cetaceans”. International Whaling Commission, Cambridge. Peterman, R.M. (1990). Statistical power analysis can improve fisheries research management. Canadian Journal of Fisheries and Aquatic Sciences 47, 2-15. Peterman, R.M. and M’Gonigle, M. (1992). Statistical power analysis and the precautionary principle. Marine Pollution Bulletin 24, 231-234. Peters, R.H. (1983). “The Ecological Implications of Body Size”. Cambridge University Press, Cambridge. Petersen, C.G.J. (1894). On the biology of our flatfishes and on the decrease of our flatfish fisheries. Reports of the Danish Biological Station N (1893), 146 pp. Peterson, B.J. and Fry, B. (1987). Stable isotopes in ecosystem studies. Annual Review of Ecology and Systematics 18, 293-320. Peterson, B.J., Howarth, R.W. and Garritt, R.H. (1985). Multiple stable isotopes used to trace the flow of organic matter in estuarine food webs. Science, N . Y., 227, 1361-1363. Philippart, C.J.M. (1998). Long-term impacts of bottom fisheries on several by-catch species of demersal fish and benthic invertebrates. ICES Journal of Marine Science. In press.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

341

Phillips, R.A., Caldow, R.W.G. and Furness, R.W. (1996). The influence of food availability on the breeding effort and reproductive success of Arctic skuas. Ibis 138,41@419. Piatt, J.F. (1987). Behavioural ecology of common murre and Atlantic puffin predation on capelin: implications for population biology. PhD Thesis, Department of Biology, Memorial University of Newfoundland, St. Johns, 31 1 pp. Piatt, J.F. (1990). Aggregative response of common murres and Atlantic puffins to their prey. Studies in Avian Biology 14, 3 6 5 1 . Pimm, S.L. (1982). “Food Webs”. Chapman and Hall, London. Pimm, S.L. (1991). “The Balance of Nature? Ecological Issues in the Conservation of Species and Communities”. University of Chicago Press, Chicago. Pimm, S.L., Lawton, J.H. and Cohen, J.E. (1991). Food web patterns and their consequences. Nature 350, 669-674. Pitcher, C.R., Burridge, C.Y., Wassenberg, T.J. and Poiner, I.R. (1997). The effects of prawn trawl fisheries on GBR seabed habitats. In “The Great Barrier Reef, Science, Use and Management, a National Conference Proceedings”, pp. 107-123. Great Barrier Reef Marine Park Authority, Townsville. Platt, T. and Denman, K. (1977). Organisation in the pelagic ecosystem. Helgolunder Wissenchaftliches Meeresuntersuchungen 30,575-58 1. Platt, T. and Denman, K. (1978). The structure of the pelagic marine ecosystem. Rapports et Procds- Verbaux des Reunions, Conseil International pour I’Exploration de la Mer 173, 60-65. Poiner, I.R., Buckworth, R.C. and Harris, A.N.M. (1990). Incidental capture and mortality of sea turtles in Australia’s northern prawn fishery. Australian Journal of Marine and Freshwater Research 41, 97-1 10. Polet, H., Blom, W. and Thiele, W. (1994). An inventory of vessels and gear types engaged in the Belgian, Dutch and German bottom trawling. In “Environmental impact of bottom gears on benthic fauna in relation to natural resources management and protection of the North Sea” (S.J. de Groot and H.J. Lindeboom, eds), pp. 7-20. Netherlands Institute for Sea Research, Den Burg, Texel. Policansky, D. (1993). Fishing as a cause of evolution in fishes. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 1-18. Springer-Verlag, Berlin. Polis, G.A. and Strong, G.R. (1996). Food web complexity and community dynamics. The American Naturalist 147, 8 13-846. Polovina, J.J. (1984). Model of a coral reef ecosystem: the ECOPATH model and its application to French Frigate Shoals. Coral Reefs 3, 1-1 1. Polovina, J.J. (1994). The lobster fishery in the north-western Hawaiian islands. In “Spiny Lobster Management” (B.F. Phillips, J.S. Cobb and J. Kittaka, eds), pp. 83-90. Blackwell Scientific Publications, London. Polunin, N.V.C. (1983). The marine resources of Indonesia. Oceanography and Marine Biology, an Annual Review 21, 455-531. Polunin, N.V.C. and Klumpp, D.W. (1992). A trophodynamic model of fish production on a windward reef flat. In “Plant-Animal Interactions in the Marine Benthos” (D.M. John, S.J. Hawkins and J.H. Price, eds), pp. 213-233. Clarendon Press, Oxford. Polunin, N.V.C. and Roberts, C.M. (1993). Greater biomass and value of target coral-reef fishes in two small Caribbean marine reserves. Marine Ecology Progress Series 100, 167-176.

342

S.JENNINGS AND M.J. KAISER

Pope, J.G. (1979). A modified cohort analysis in which constant natural mortality is replaced by estimates of predation levels. International Councilfor the Exploration of the Sea, Committee Meeting 1979, H:16. Pope, J.G. and Macer, C.T. (1996). An evaluation of the stock structure of North Sea cod, haddock, and whiting since 1920, together with a consideration of the impacts of fisheries and predation effects on their biomass and recruitment. ICES Journal of Marine Science 53, 1157-1 169. Pope, J.G., Stokes, T.K., Murawski, S.A. and Iodoine, S.I. (1988). A comparison of fish size composition in the North Sea and on Georges Bank. In “Ecodynamics: Contributions to Theoretical Ecology” (W. Wolff, C.J. Soeder and F.R. Drepper, eds). Springer-Verlag, Berlin. Porter, J.W., Porter, K.G. and Batac-Catalan, Z. (1977). Quantitative sampling of Indo-Pacific demersal reef plankton. Proceedings of the Third International Coral Reef Symposium 1, 105-1 12. Posey, M., Lindberg, W., Alphin, T. and Vose, F. (1996). Influence of storm disturbance on an offshore benthic community. Bulletin of Marine Science 59, 523-529. Potin, P., Floc’h, J.Y., Augris, C. and Cabioch, J. (1990). Annual growth rate of the calcareous red alga Lithothamnion corallioides (Corallinales, Rhodophyta) in the Bay of Brest, France. Hydrobiologia 204, 263-267. Potter, E.C.E. and Pawson, M.G. (1991). Gill netting. Laboratory Leajets, MAFF, Directorate of Fisheries Research, Lowestoft, 69, 34 pp. Potts, D.C. (1977). Suppression of coral populations by filamentous algae within damselfish territories. Journal of Experimental Marine Biology and Ecology 28, 207-216. Powers, K.D. and Brown, R.G.B. (1987). Seabirds. In “Georges Bank” (R.H. Backus, ed.), pp. 359-371, MIT Press, Cambridge, MA. Preston, T. (1992). The measurement of stable isotope natural abundance variations. Plant Cell and Environment 15, 1091-1097. Priddle J., Croxall J.P., Everson I., Heywood R.B., Murphy E.J., Prince P.A., Sear C.B. (1988). Large-scale fluctuations in distribution and abundance of krill - a discussion of possible causes. In: “Antarctic Ocean and Resources Variability” (D. Sahrhage, ed), pp 169-1 82. Springer-Verlag, Berlin. Pyle, R.L. (1993). Marine aquarium fish. In “Nearshore Marine Resources of the South Pacific” (A. Wright and L. Hill, eds), pp. 135-176. Institute of Pacific Studies, Suva. Raffaelli, D. and Milne, H. (1987). An experimental investigation of the effects of shorebird and flatfish predation on estuarine invertebrates. Estuarine, Coastal and Shelf Science, 24, 1-13. Raffaelli, D., Conacher, A., McLachlan, H. and Emes, C. (1989). The role of epibenthic crustacean predators in an estuarine food web. Estuarine, Coastal and Shelf Science, 28, 149-160. Raloff, J. (1996). Trawling: the bottom line. Science News 1996, pp. 268-271. Ramsay, K., Kaiser, M.J. and Hughes, R.N. (1996). Changes in hermit crab feeding patterns in response to trawling disturbance. Marine Ecology Progress Series 144, 63-12. Ramsay, K., Kaiser, M.J. and Hughes, R.N. (1997a). A field study of intraspecific competition for food in hermit crabs (Pagurus bernhardus). Estuarine, Coastal and Shelf Science 44,2 13-220.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

343

Ramsay, K., Kaiser, M.J., Moore, P.G. and Hughes, R.N. (1997b). Consumption of fisheries discards by benthic scavengers: utilisation of energy subsidies in different marine habitats. Journal of Animal Ecology 66, 884-896. Randall, J.E. (1967). Food habits of reef fishes of the West Indies. Studies in Tropical Oceanography 5, 665-847. Randall, J.E. and Heemstra, P.C. (1991). Revision of Indo-Pacific groupers (Perciformes: Serranidae: Epinephelinae) with descriptions of five new species. Indo-Pacific Fishes 20, 1-332. Reinthal, P.N., Kensley, B. and Lewis, S.M. (1984). Dietary shifts in the Queen Triggerfish Balistes vetula in the absence of its primary food item Diadema antillarum. Marine Ecology 5, 191-195. Reise, K. (1981). High abundance of small zoobenthos around biogenic structures in tidal sediments of the Waddensea. Helgolander Wissenschaftliches Meeresuntersuchungen 34, 413-425. Reznick, D.A. and Bryga, H. (1987). Life history evolution in guppies (Poecilia reticulata): 1: Phenotypic change in an introduction experiment. Evolution 41, 137Ck1385. Reznick, D.A., Bryga, H. and Endler, J.A. (1990). Experimentally induced life-history evolution in a natural population. Nature 346, 357-359. Reznick, D.N. (1993). Norms of reaction in fishes. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 72-90. SpringerVerlag, Berlin. Rhoads, D.C. (1974). Organism-sediment relations on the muddy sea floor. Oceanography and Marine Biology: an Annual Review 12, 263-300. Rice, J. and Gislason, H. (1996). Patterns of change in the size spectra of numbers and diversity of the North Sea fish assemblage, as reflected in surveys and models. ICES Journal of Marine Science 53, 1214-1225. Ricker, W.E. (1958). Handbook of computations for biological statistics of fish populations. Bulletin of the Fisheries Research Board of Canada 119, 300. Riesen, W. and Reise, K. (1982). Macrobenthos of the subtidal Wadden Sea: revisited after 55 years. Helgolander Wissenschaftliches Meeresuntersuchungen 35,409423. Rijnsdorp, A.D. (1989). Maturation of male and female North Sea plaice (Pleuronectes platessa L.). Journal du Conseil, Conseil International pour I’Exploration de la Mer 46, 35-51. Rijnsdorp, A.D. (1990). The mechanism of energy allocation over reproduction and somatic growth in North Sea plaice, Pleuronectes platessa L. Netherlands Journal of Sea Research 25, 279-290. Rijnsdorp, A.D. (1993a). Fisheries as a large scale experiment on life history evolution: disentangling phenotypic and genetic effects in changes in maturation and reproduction of North Sea plaice, Pleuronectes platessa L. Oecologia 96, 391401. Rijnsdorp, A.D. (1993b). Selection differentials in male and female North Sea plaice and changes in maturation and fecundity. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 19-36. SpringerVerlag, Berlin. Rijnsdorp, A.D. and van Leeuwen, P.I. (1996). Changes in growth of North Sea plaice since 1950 in relation to density, eutrophication, beam-trawl effort and temperature. ICES Journal of Marine Science 53, 1 199-121 3. Rijnsdorp, A.D., Daan, N., van Beek, F.A. and Heessen, H.J.L. (1991a). Reproductive variability in North Sea plaice, sole and cod. Journal du Conseil, Conseil International pour I’Exploration de la Mer 47, 352-375.

344

S. JENNINGS AND M.J. KAISER

Rijnsdorp, A.D., Groot, P. and van Beek, F.A. (1991b). The micro distribution of beam trawl effort in the southern North Sea. International Council for the Exploration of the Sea, Committee Meeting 1991, G:49. Rijnsdorp, A.D., Buijs, A.M., Storbeck, F. and Visser, E. (1996a). Micro-scale distribution of beam trawl effort in the southern North Sea between 1993 and 1996 in relation to the trawling frequency of the sea bed and the impact on benthic organisms. International Council for the Exploration of the Sea, Committee Meeting ICES 1996/mini 11. Rijnsdorp, A.D., van Leeuwen, P.I., Daan, N. and Heessen, H.J.L. (1996b). Changes in abundance of demersal fish species in the North Sea between 1906-1909 and 1990-1995. ICES Journal of Marine Science 53, 1054-1062. Rijnsdorp, A. D., Buys, A. M., Storbeck, F. and Visser, E. (1997). The micro-distribution of the Dutch beam trawl fishery between April 1993 and March 1996. Reports of the Netherlands Institute for Fisheries Research, Ijmuiden. Risk, M.J. (1972). Fish diversity on a coral reef in the Virgin Islands. Atoll Research Bulletin 153, 1-6. Roberts, C.M. (1995). The effects of fishing on the ecosystem structure of coral reefs. Conservation Biology 9, 988-995. Roberts, C.M. and Ormond, R.F.G. (1987). Habitat complexity and coral reef fish diversity and abundance on Red Sea fringing reefs. Marine Ecology Progress Series 41, 1-8. Roberts, C.M. and Polunin, N.V.C. (1991). Are marine reserves effective in management of reef fisheries? Reviews in Fish Biology and Fisheries 1, 65-91. Roberts, C.M. and Polunin, N.V.C. (1992). Effects of marine reserve protection on northern Red Sea fish populations. Proceedings of the Seventh International Coral Reef Symposium 2, 969-977. Roberts, C.M. and Polunin, N.V.C. (1993). Marine reserves: simple solutions to managing complex fisheries. Ambio 22, 363-368. Robin, J.-P. (1992). The brown shrimp fishery of the Loire Estuary: production and by-catch of juvenile fish. Fisheries Research 13, 153-172. Roff, D.A. (1984). The evolution of life history parameters in teleosts. Canadian Journal of Fisheries and Aquatic Sciences 41, 989-1000. Ross, R.M. (1990). The evolution of sex change mechanisms in fishes. Environmental Biology of Fishes 29, 81-93. Rothschild, B.J. (1991). Multispecies interactions on Georges Bank. ICES Marine Science Symposia 193, 86-92. Rowden, A.A. and Jones, M.B. (1993). Critical evaluation of sediment turnover estimates for Callianassidae (Decapoda: Thalassinidea). Journal of Experimental Marine Biology and Ecology 173, 265-272. Rowell, C.A. (1993). The effects of fishing on the timing of maturity in North Sea cod (Gadus morhua L.). In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 4461. Springer-Verlag, Berlin. Rowley, R.J. (1994). Marine reserves in fisheries management. Aquatic Conservation: Marine and Freshwater Ecosystems 4, 233-254. Rubec, P.J. (1986). The effects of sodium cyanide on coral reefs and marine fish in the Philippines. In “The First Asian Fisheries Forum” (J.L. McClean, L.B. Dizon and L.V. Hosillos, eds), pp. 297-302. Asian Fisheries Society, Manila. Ruddle, K. (1996). Geography and human ecology of reef fisheries. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 137-160. Chapman and Hall, London.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

345

Rumohr, H. and Krost, P. (1991). Experimental evidence of damage to the benthos by bottom trawling with special reference to Arctica islandica. Meeresforschung 33, 34c345. Russ, G.R. (1985). Effects of protective management on coral reef fishes in the central Philippines. Proceedings of the Fifth International Coral Reef Symposium 4, 219-224. Russ, G.R. (1991). Coral reef Fisheries: effects and yields. In “The Ecology of Fishes on Coral Reefs” (P.F. Sale, ed), pp. 601-635. Academic Press, San Diego. Russ, G.R. and Alcala, A.C. (1989). Effects of intense fishing pressure on an assemblage of coral reef fishes. Marine Ecology Progress Series 56, 13-27. Russ, G.R. and Alcala, A.C. (1996a). Do marine reserves export adult fish biomass? - evidence from Apo Island, Central Philippines. Marine Ecology Progress Series 132, 1-9. Russ, G.R. and Alcala, A.C. (1996b). Marine reserves - rates and patterns of recovery and decline of large predatory fish. Ecological Applications 6, 947-961. Russell, E.S. (1931). Some theoretical considerations of the overfishing problem. Journal du Conseil International pour I’Exploration de la Mer 6, 3-20. Russell, E.S. (1939). An elementary treatment of the overfishing problem. Journal du Conseil International pour I’Exploration de la Mer 11, 5-14. Russell, F.S., Southward, A.J., Boalch, G.T. and Butler, E.I. (1971). Changes in biological conditions in the English Channel off Plymouth during the last halfcentury. Nature 234, 468470. Ryman, N., Utter, F. and Laikre, L. (1995). Protection of intraspecific diversity of exploited fishes. Reviews in Fish Biology and Fisheries 5, 4 1 7 4 6 . Ryther, J.H. (1969). Relationships of photosynthesis to fish production in the sea. Science, N . Y., 166, 72-76. Sadovy, Y.J. (1996). Reproduction of reef fishery species. In “Reef Fisheries” (N.V.C. Polunin and C.M. Roberts, eds), pp. 15-59. Chapman and Hall, London. Saila, S.B., Kocic, V.L. and McManus, J.W. (1993). Modeling the effects of destructive fishing practices on tropical coral reefs. Marine Ecology Progress Series 94, 5140. Sainsbury, K.J. (1987). Assessment and management of the demersal fishery on the continental shelf of northwestern Australia. In “Tropical Snappers and Groupers - Biology and Fisheries Management” (J.J. Polovina and S. Ralston, eds), pp. 465503. Westview Press, Boulder, Colorado. Sainsbury, K.J. (1988). The ecological basis of multispecies fisheries and management of a demersal fishery in tropical Australia. In “Fish Population Dynamics” (J.A. Gulland, ed), pp. 349-382. John Wiley, London. Sainte-Marie, B. (1986). Effect of bait size and sampling time on the attraction of the lysianassid amphipods Anonyx sarsi Steele and Brunel and Orchomenella pinguis (Boeck). Journal of Experimental Marine Biology and Ecology 99, 63-77. Sainte-Marie, B. and Hargrave, B.T. (1987). Estimation of scavenger abundance and distance of attraction to bait. Marine Biology 94, 4 3 1 4 3 . Sala, E. and Zabala, M. (1996). Fish predation and the structure of the sea-urchin Paracentrotus Iividus populations in the NW Mediterranean. Marine Ecology Progress Series 140, 7 1-8 1 . Sammarco, P.W. (1980). Diadema and its relationship with coral spat mortality: grazing, competition and biological disturbance. Journal of Experimental Marine Biology and Ecology 45, 245-272.

346

S. JENNINGS AND M.J. KAISER

Samoilys, M. (1988). Abundance and species richness of coral reef fish on the Kenyan coast: the effects of protective management and fishing. Proceedings of the Sixth International Coral Reef Symposium 2, 261-266. Sano, M., Shimizu, M. and Nose, Y. (1984). Changes in structure of coral reef fish communities by destruction of hermatypic corals: observational and experimental views. Pacifc Science 38, 51-79. Santander, A., Alheit, J., MacCall, A. and Alamo, A. (1983). Egg mortality of the Peruvian anchovy (Engraulis ringens) caused by cannibalism and predation by sardines (Sardinops sagax). FA0 Fisheries Report 291(3), 1011-1025. Schaefer, M.B. (1954). Fisheries dynamics and the concept of maximum equilibrium catch. Proceedings of the Gulf and Caribbean Fisheries Institute 6, 1-1 1. Schubel, J.R., Carter, H.H., Wilson, R.E., Wise, W.M., Heaton, M.G. and Gross, M.G. (1978). “Field Investigations of the Nature, Degree and Extent of Turbidity Generated by Open Water Pipeline Disposal Operations”. Marine Sciences Research Centre, State University of New York, New York. Schulze, E.D. and Mooney, H.A. (1993). Ecosystem function of biodiversity: a summary. In “Biodiversity and Ecosystem Function” (E.D. Schulze and H.A. Mooney, eds), pp. 497-5 10. Springer-Verlag, Berlin. Schwinghamer, P., Guignk, J.Y. and Siu, W.C. (1996). Quantifying the impact of trawling on benthic habitat structure using high resolution acoustics and chaos theory. Canadian Journal of Fisheries and Aquatic Sciences 53, 288-296. Shackell, N.L. and Martin Willison, J.H. (eds) (1995). “Marine Protected Areas and Sustainable Fisheries”. Science and Management of Protected Areas Association, Halifax, Nova Scotia. Shackelton, L.Y. (1987). A comparative study of fossil fish scales from three upwelling regions. South African Journal of Marine Science 5, 79-84. Shapiro, D.Y., Sadovy, Y. and McGehee, M.A. (1993). Periodicity of sex change and reproduction in the red hind, Epinephelus guttatus, protogynous grouper. Bulletin of Marine Science 53, 1151-1162. Shapiro, D.Y., Garciamoliner, G. and Sadovy, Y. (1994a). Social system of an inshore stock of the red hind grouper Epinephelus guttatus (Pisces, Serranidae). Environmental Biology of Fishes 41, 415422. Shapiro, D.Y., Marconato, A. and Yoshikawa, T. (1994b). Sperm economy in a coral reef fish Thalassoma bifasciatum. Ecology 75, 1334-1344. Shapiro, J. and Wright, D.I. (1984). Lake restoration by biomanipulation. Freshwater Biology 14, 371-383. Sharp, G.D. (1988). Fish populations and fisheries: their perturbations, natural and man-induced. In “Ecosystems of the World 27: Continental Shelves” (H. Postma and J.J. Zijlstra, eds), pp. 155-202. Elsevier, Amsterdam. Sheldon, R.W., Prakash, A. and Sutcliffe, W.H.(1972). The size distribution of particles in the Ocean. Limnology and Oceanography 17, 327-340. Sheldon, R.W., Sutcliffe, W.H. and Prakash, A. (1973). The production of particles in the surface waters of the ocean with particular reference to the Sargasso Sea. Limnology and Oceanography 18, 719-733. Shelton, P.A. (1992). Detecting and incorporating multispecies effects into fisheries management in the north-west and south-east Atlantic. South African Journal of Marine Science 12, 723-737. Shepherd, J.G. and Cushing, D.H. (1990). Regulation in fish populations: myth or mirage? Philosophical Transactions of the Royal Society B 330, 151-164. Shepherd, S.A. (1983). The epifauna of megaripples: species’ adaptations and population responses to disturbance. Australian Journal of Ecology 8, 3-8.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

347

Sherman, K. (1991). The large marine ecosystem concept: research and management strategy for living marine resources. Ecological Applications 1, 349-360. Sherman, K. and Alexander, L.M. (eds) (1986). “Variability and Management of Large Marine Ecosystems”. Westview Press, Boulder. Sherman, K., Alexander, L.M. and Gold, B.D. (eds) (1991). “Food Chains, Yields, Models and Management of Large Marine Ecosystems”. Westview Press, Boulder. Sherman, K., Alexander, L.M. and Gold, B.D. (eds) (1993). “Large marine ecosystems: stress, mitigation and sustainability”. American Association for the Advancement of Science Press, Washington. Sherman, K., Jones, C., Sullivan, L., Smith, W., Berrien, P. and Ejsymont, L. (1981). Congruent shifts in sandeel abundance in western and eastern North Atlantic ecosystems. Nature, London, 291, 4 8 H 8 9 . Short, F.T. and Wyllie-Echeverria, S. (1996). Natural and human-induced disturbance of seagrasses. Environmental Conservation 23, 17-28. Siegel, V. and Loeb, V. (1995). Recruitment of Antarctic krill Euphausia superba and possible causes for its variability. Marine Ecology Progress Series 123, 45-56. Siegfried, W.R. and Crawford, R.J.M. (1978). Jackass penguins, eggs and guano: diminishing resources at Dassen Islands. South African Journal of Science 74, 389-390. Silliman, R. ( 1975). Selective and unselective exploitation of experimental populations of Tilapia mossambica. Fishery Bulletin US.73, 495-501. Simmonds, M.P. and Hutchinson, J.D. (eds) (1996). “The Conservation of Whales and Dolphins”. John Wiley and Sons, Chichester. Sissenwine, M.P. (1986). Perturbation of a predator controlled continental shelf ecosystem. In “Variability and Management of Large Marine Ecosystems” (K. Sheman and L.M. Alexander, eds), pp. 55-85. Westview Press, Boulder. Sissenwine, M.P., Brown, B.E., Palmer, J.E. and Essig, R.J. (1982). Empirical examination of population interactions for the fishery resources off the northeastern USA. Canadian Special Publications in Fisheries and Aquatic Sciences 59, 82-94. Sissenwine, M.P., Cohen, E.B. and Grosslein, M.D. (1984). Structure of the Georges Bank ecosystem. Rapports et Procis- Verbaux des RPunions, Conseil International pour I’Exploration de la Mer 183, 243-254. Sladen, W.J.L. (1964). The distribution of the Adelie and chinstrap penguins. In “Biologie Antarctique” (R. Carrick, M.W. Holdgate and J. Prevost, eds), pp. 359-365. Hermann, Paris. Smith, B.D. and Jamieson, G.S. (1991). Possible consequences of intensive fishing for males on the mating opportunities of Dungeness crabs. Transactions of the American Fisheries Society 120, 650-653. 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 32, 41 7 4 2 . Smith, P.J., Francis, R.I.C.C. and McVeagh, M. (1991). Loss of genetic diversity due to fishing pressure. Fisheries Research, London 10, 309-3 16. Smith, T.D. (1983). Changes in size of three dolphin (Stenella sp.) populations in the eastern tropical Pacific. Fishery Bulletin U S .81, 1-13. Smith, T.D. (1994). “Scaling fisheries: the science of measuring the effects of fishing, 1855-1955”. Cambridge University Press, Cambridge. Smolowitz, R.J. (1978). Trap design and ghost fishing: An overview. Marine Fisheries Review 40,2-8. Soutar, A. and Isaacs, J.D. (1974). Abundance of pelagic fish during the 19th and 20th centuries as recorded in anaerobic sediment off the Californias. Fishery Bulletin, U.S., 72, 257-273.

348

S. JENNINGS AND M.J. KAISER

Southward, A.J. (1963). The distribution of some plankton animals in the English Channel and approaches: I11 Theories about long-term biological changes including fish. Journal of the Marine Biological Association of the United Kingdom 43, 1-29. Southward, A.J. (1980). The western English Channel - an inconstant ecosystem. Nature, London 285, 361-366. Southward, A.J., Boalch, G.T. and Maddock, L. (1988). Fluctuations in the herring and pilchard fisheries of Devon and Cornwall linked to changes in climate since the 16th century. Journal of the Marine Biological Association of the United Kingdom 68,423-445. Sparholt, H. (1990). An estimate of the total biomass of fish in the North Sea. Journal du Conseil, Conseil International pour I’Exploration de la Mer 46,200-210. Sparre, P. (1991). Introduction to multispecies virtual population analysis. ICES Marine Science Symposia 193, 12-21. Spencer B.E., Kaiser M.J. and Edwards D.B. (1997). Ecological effects of Manila clam cultivation: observations at the end of the cultivation phase. Journal of Applied Ecology 34, 444-452. Sprules, W.G. and Goyke, A.P. (1994). Size-based structure and production in the pelagia of Lakes Ontario and Michigan. Canadian Journal of Fisheries and Aquatic Sciences 51, 2603-261 1. Sprules, W.G. and Stockwell, J.D. (1995). Size based biomass and production models in the St Lawrence Great Lakes. ICES Journal of Marine Science 52, 705710. Sprules, W.G., Casselman, J.M. and Shuter, B.J. (1983). Size distributions of pelagic particles in lakes. Canadian Journal of Fisheries and Aquatic Sciences 40,1761-1765. Steadman, D.R. (1995). Prehistoric extinctions of Pacific Island birds: biodiversity meets zooarchaeology. Science 267, 1123-1 131. Stearns, S.C. (1976). Life-history tactics: a review of the ideas. Quarterly Review of Biology 51, 3 4 7 . Stearns, S.C. and Crandall, R.E. (1984). Plasticity for age and size at sexual maturity: a life history response to unavoidable stress. In “Fish Reproduction: Strategies and Tactics” (G.W. Potts and R.J. Wootton, eds), pp. 13-34. Academic Press, London. Steele, J.H. (1974). “The Structure of Marine Ecosystems”. Blackwell Scientific Publications, Oxford. Steele, J.H. and Henderson, E.W. (1984). Modeling long-term fluctuations in fish stocks. Science 224,98>986. Stephenson, R.L. and Kornfield, I. (1990). Reappearance of spawning Atlantic herring (Clupea harengus harengus) on Georges Bank - population resurgence not recolonisation. Canadian Journal of Fisheries and Aquatic Sciences 47, 1060-1 064. Stillman, R.A., Goss-Custard, JD., McGrorty, S., West, A.D., Durell, S.E.A. le V. dit, Clarke, R.T., Caldow, R.W.G., Norris, K.J., Johnstone, I.G., Ens, B.J., Bunskoeke, E.J., Merwe, A.v.d., van der Meer, J., Triplet, P., Odoni, N., Swinfen, R. and Cayford, J.T. (1996). Models of shellfish populations and shorebirds. Report to Commission of the European Communities by the Institute of Terrestrial Ecology, Furzebrook PEM/93/03. Stockton, W.L. and DeLaca, T.E. (1982). Food falls in the deep sea: occurrence, quality and significance. Deep-sea Research 29, 157-169. Stokes, T.K. and Blythe, S.P. (1993). Size selective harvesting and age at maturity 2. Real populations and management options. In “The Exploitation of Evolving Resources” (T.K. Stokes, J.M. McGlade and R. Law, eds), pp. 232-247. Springer-Verlag, Berlin.

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

349

Stokes, T.K., McGlade, J.M. and Law, R. (eds) (1993). “The Exploitation of Evolving Resources”. Springer-Verlag, Berlin. Sugihara, G., Garcia, S., Gulland, J.A., Lawton, J.H., Maske, H., Paine, R.T., Platt, T., Rachor, E., Rothschild, B.J., Ursin, E.A. and Zeitzschel, B.F.K. (1984). Ecosystems dynamics. In “Exploitation of Marine Communities” (R.M. May, ed.), pp. 131-153. Springer-Verlag, Berlin. Sutcliffe, W.H., Drinkwater, K. and Muir, B.S. (1977). Correlations of fish catch and environmental factors in the Gulf of Maine. Journal of the Fisheries Research Board of Canada 34, 19-30. Sweatman, H.P.A. (1995). A field study of fish predation on juvenile crown-of-thorns starfish. Coral Reefs 14, 47-53. Thiel, H. (1992). Deep-sea environmental disturbance and recovery potential. Internationale revue der Gesamten Hydrobiologie 77, 33 1-339. Thiel, H. and Schriever, G. (1990). Deep-sea mining, environmental impact and the DISCOL project. Ambio 19, 245-250. Thiebaux, M.L. and Dickie, L.M. (1992). Models of aquatic biomass size spectra and the common structure of their solutions. Journal of Theoretical Biology 159, 147-1 6 1. Thiebaux, M.L. and Dickie, L.M. (1993). Structure of the body size spectrum of the biomass in aquatic ecosystems: a consequence of allometry in predator-prey interactions. Canadian Journal of Fisheries and Aquatic Sciences 50, 1308-1 3 17. Thomas, C.J. and Cahoon, L.B. (1993). Stable isotope analyses differentiate between different trophic pathways supporting rocky-reef fishes. Marine Ecology Progress Series 95, 19-24. Thompson, D.R., Furness, R.W. and Lewis, S.A. (1995). Diets and long-term changes in ‘’N and I3C values in northern fulmars Fulmaris glacialis from two northeast Atlantic colonies. Marine Ecology Progress Series 125, 3-1 1. Thompson, K.R. (1992). Quantitative analysis of the use of discards from squid trawlers by black-browed albatrosses Diomedea melanophris in the vicinity of the Falkland Islands. Ibis 134, 11-21. Thompson, R. and Munro, J.L. (1983). The biology, ecology and bionomics of the hinds and groupers, Serranidae. ICLARM Studies and Reviews 7, 59-8 1. Thresher, R.E. (1983). Environmental correlates of the distribution of planktivorous fishes in the One Tree Reef Lagoon. Marine Ecology Progress Series 10, 137-145. Thrush, S.F., Hewitt, J.E., Cummings, V.J. and Dayton, P.K. (1995). The impact of habitat disturbance by scallop dredging on marine benthic communities: what can be predicted from the results of experiments? Marine Ecology Progress Series 129, 141-150. Thrush, S.F., Whitlatch, R.B., Pridmore, R.D., Hewitt, J.E., Cummings, V.J. and Wilkinson, M.R. (1996). Scale-dependent recolonization: the role of sediment stability in a dynamic sandflat habitat. Ecology 77, 2472-2487. Tiews, K. (1978). On the disappearance of bluefin tuna in the North Sea and its ecological implications for herring and mackerel. Rapports et Proc&- Verbaux des Re‘unions, Conseil International pour I’Exploration de la Mer 172, 30 1-309. Tilman, D. (1996). Biodiversity - population versus ecosystem sustainability. Ecology 77, 35C363. Tilman, D. and Downing, J.A. (1994). Biodiversity and stability in grasslands. Nature 367, 363-365. Tilman, D., Wedin, D. and Knops, J. (1996). Productivity and sustainability influenced by biodiversity in grassland ecosystems. Nature, London, 6567, 718-720.

350

S. JENNINGS AND M.J. KAISER

Trippel, E.A. (1995). Age at maturity as a stress indicator in fishes. BioScience 45, 759-77 1. Turner, S.J. (1994). The biology and population outbreaks of the corallivorous gastropod Drupella on Indo-Pacific reefs. Oceanography and Marine Biology, an Annual Review 32, 461-530. Twichell, D.C., McClennen, C.E. and Butman, B. (1981). Morphology and processes associated with the accumulation of fine grain sediment deposit on the southern New England shelf. Journal of Sedimentary Petrology 51, 269-280. Ulanowicz, R.E. (1986). “Growth and Development, Ecosystems Phenomenology”. Springer-Verlag, New York. Ursin, E. (1982). Stability and variability in the marine ecosystem. Dana 2, 51-67. Vader, W., Barrett, R.T., Erikstad, K.E. and Strann, K.B. (1990). Differential responses of common and thick-billed murres to a crash in the capelin stock in the southern Barents Sea. Studies in Avian Biology 14, 175-180. Van Dolah, R.F., Wendt, P.H. and Levisen, M.V. (1991). A study of the effects of shrimp trawling on benthic communities in two South Carolina sounds. Fisheries Research, London 12, 139-156. Vane-Wright, R.I. (1996). Identifying priorities for the conservation of biodiversity: systematic biological criteria within a socio-political framework. In “Biodiversity: a Biology of Numbers and Difference” (K.J. Gaston, ed), pp. 309-344. Blackwell Science, Oxford. Vince, M.R. (1991). Stock identity in spurdog (Squalus acanthias L.) around the British Isles. Fisheries Research 12, 341-354. Vitousek, P.M., Ehrlich, P.R., Ehrlich, A.H. and Matson, P.A. (1986). Human appropriation of the products of photosynthesis. Bioscience 36, 368-373. Vivien, M.L. (1973). Contribution a la connaissance de l’kthologie alimentaire de l’icthyofauna du platier interne des rkifs coralliens de TulCar (Madagascar). Tethys Supplement 5, 221-308. Vivien, M.L. and Peyrot-Clausade, M. (1974). A comparative study of the feeding behaviour of three coral reef fishes (Holocentridae), with special reference to the polychaetes of the reef cryptofauna as prey. Proceedings of the Second International Coral Reef Symposium 1, 179-192. Volterra, V. (1926). Variazioni e fluttuazioni del numero d’individui in specie animali conviventi. Memoire della Regia Accademia Nazionale dei Lincei 2, 31-1 13. Von Blaricom, G.R. (1982). Experimental analysis of structural regulation in a marine sand community exposed to oceanic swell. Ecological Monographs 52, 283-305. Wainwright, S.C., Fogarty, M.J., Greenfield, R.C. and Fry, B. (1993). Long term changes in the Georges Bank food web - trends in stable isotopic composition of fish scales. Marine Biology 115, 481-493. Walker, B.H. (1992). Biodiversity and ecological redundancy. Conservation Biology 6, 18-23. Walker, P.A. and Heessen, H.J.L. (1996). Long-term changes in ray populations in the North Sea. ICES Journal of Marine Science 53, 1085-1093. Walsh, J.J. (1975). A spatial simulation model of the Peru upwelling ecosystem. Deep Sea Research 22, 1-13. Walters, C.J. (1986). “Adaptive Management of Renewable Resources”. Macmillan, New York. Walters, C.J. and Juanes, F. (1993). Recruitment limitation as a consequence of natural selection for use of restricted feeding habitats and predation risk taking

THE EFFECTS OF FISHING ON MARINE ECOSYSTEMS

35 1

by juvenile fishes. Canadian Journal of Fisheries and Aquatic Sciences 50, 20582070. Walters, C.J., Stocker, M., Tyler, A.V. and Westrheim, S.J. (1986). Interaction between Pacific cod (Gadus macrocephalus) and herring (Clupea harengus pallasz] in the Hecate Strait, British Columbia. Canadian Journal of Fisheries and Aquatic Sciences 43, 830-837. Walters, C.J., Christensen, V. and Pauly, D. (1997). Structuring dynamic models of exploited ecosystems from trophic mass-balance assessments. Reviews in Fish Biology and Fisheries 7 , 139-1 72. Wanink, J.H. (1991). Survival in a perturbed environment: the effects of Nile perch introduction on the zooplanktivorous fish community of Lake Victoria. In “Terrestrial and Aquatic Ecosystems: Perturbation and Recovery” (0. Ravera, ed), pp. 269-275. Ellis-Horwood, New York. Wanless, S. and Harris, M.P. (1992). Activity budgets, diet and breeding success of kittiwakes Rissa tridactyla on the Isle of May. Bird Study 39, 145-154. Warren, P.H. (1990). Variation in food web structure - the determinants of connectance. The American Naturalist 136, 689-700. Warren, P.H. (1994). Making connections in food webs. Trends in Ecology and Evolution 9, 136141. Wassenberg, T.J. and Hill, B.J. (1990). Partitioning of material discarded by prawn trawlers in Moreton Bay. Australian Journal of Marine and Freshwater Research 41, 27-36. Watson, M. and Ormond, R.F.G. (1994). Effects of an artisanal fishery on the fish and urchin populations of a Kenyan coral reef. Marine Ecology Progress Series 109, 115-129. Welleman, H. (1989). “De verspreiding van een aantal cacrobenthos soorten in de Noordzee”. NIOZ, Den Burg, Texel. Werren, J.H. and Charnov, E.L. (1978). Facultative sex ratios and population dynamics. Nature 171, 349-350. Westman, W.E. (1977). How much are nature’s services worth? Science, N. Y. 197, 960-964. Williams, D.M. (1986). Temporal variation in the structure of reef slope fish communities (central Great Barrier Reef): short term effects of Acanthaster planci infestation. Marine Ecology Progress Series 28. Witbaard, R. and Klein, R. (1994). Long-term trends on the effects of the southern North Sea beamtrawl fishery on the bivalve mollusc Arctica islandica L. (Mollusca, bivalvia). ICES Journal of Marine Science 51, 99-105. Witman, J.D. (1988). Effects of predation by the fireworm Hermodice carunculata on milleporid corals. Bulletin of Marine Science 42, 4 4 M 5 8 . Wright, P.J. (1996). Is there a conflict between sandeel fisheries and seabirds? A case study at Shetland. In “Aquatic Predators and their Prey” (S.P.R. Greenstreet and M. Tasker, eds), pp. 154-165. Blackwell Scientific Publications, Oxford. Wright, P.J. and Bailey, M.C. (1996). Timing of hatching in Ammodytes marinus from Shetland waters and its significance to early growth and survivorship. Marine Biology 126, 143-152. Yoshioka, Y., Wada, E. and Hayashi, H. (1994). A stable isotope study on seasonal food web dynamics in a eutrophic lake. Ecology 75, 835-846. Zann, L., Brodie, J. and Vuki, V. (1990). History and dynamics of the crown-ofthorns starfish Acanthaster planci (L.)in the Suva area, Fiji. Coral Reefs 9, 135-144.

352

S. JENNINGS AND M.J. KAISER

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