Sharp changes in the abundance of targeted and non-target invertebrate species, and also in the relative species composition of exploited communities, have been detected worldwide, as a result of growing fishing intensity (Orensanz, Parma and Hall, 1998). Overfishing with a progressive decrease in stock size commonly occurs at the highest levels of the trophic chain (Bustamante and Castilla, 1987; Steneck et al., 2002). This has led to the recognition that in practice, harvesting affects incidental or intentionally different species within a community (Caddy and Sharp, 1986; Defeo, 1998; Castilla, 1999; Cabrera and Defeo, 2001). Over the last two decades, fishery scientists have taken a more holistic approach to management (Caddy and Sharp, 1986; Walters, Christensen and Pauly, 1997; Fulton, Smith and Johnson, 2003), and thus research has been directed at evaluating the ecological effects of fishing on invertebrate communities (Thrush et al., 1998; Menge et al., 1999; Tegner and Dayton, 2000).
Direct and indirect effects of fishing have been widely reported on for many marine benthic assemblages (Tegner and Dayton, 1999 and references therein). Much of this work has been focused on industrial shellfisheries with emphasis on mobile fishing gears, whereas little is known about the quantitative effect of small scale fishing gears intensively used in coastal shellfisheries. An understanding of the role of fishing and how it affects ecological functions is needed to place enhancement programmes in an integrated management context. In synthesizing the effects of fishing on estuaries and nearshore systems, Blaber et al. (2000) defined eight process-orientated categories according to the nature and extent of the fishing impact: target organisms, non-target organisms, nursery functions, trophic effects, habitat change, reduction of water quality, human environment, and extinctions. The marked decline in commercial catches of top carnivorous species and the general increase in species low in the food chain suggest potential trophic effects of fishing caused by the concurrent increase in fishing technology and effort. This agrees with the "fishing down the food web" hypothesis of Pauly et al. (1998, 2001), but it is also clear that changes in fishing technology have played a crucial role (Caddy, 1999a). A number of examples provided strong support to the sequential depletion hypothesis; i.e.: overexploitation of target species first and incidental ones later on. According to Orensanz et al. (1998) in their extensive study in the Greater Gulf of Alaska, the pattern of collapse in several shellfisheries is not haphazard but proceeds serially, starting with the most valuable resources.
Trophic cascades are defined by Pinnegar et al. (2000) as "predatory interactions involving three trophic levels, whereby primary carnivores, by suppressing herbivores, increase plant abundance" (Figure 5.1). These are also known as "community-level cascades" (sensu Polis et al., 2000), where plant biomass is substantially altered throughout the system as a response of predator removals. Examples involving three or more levels in a food web are shown in Figure 5.2. from Pinnegar et al. (2000), and may not necessarily involve marine plants. Polis et al. (2000) define a "species-level cascade" as a mechanism which can account for changes in a subset of a community, such that changes in predator abundance affect several species, including one or a few plant species. They emphasise that although the definitions of species-level and community-level cascades refer explicitly to the three levels mentioned above, "they also apply to any multilink linear food-web interaction".
Trophic cascades involving shellfish fisheries have been identified mainly on rocky substrates (Estes and Palmisano, 1974, see review in Tegner and Dayton, 2000; see also Steneck et al., 2002), largely because on particulate sediments and in the water column, the linkages are less obvious. One could model quantitatively the linked species groups shown in Figures 5.1 and 5.2 through softwares such as ECOPATH and its derivatives (Walters, Christensen and Pauly, 1997; Pauly, Christensen and Walters, 2000), but for purposes of management it may be sufficient to be aware that such linkages exist, and avoid destabilizing the ecosystem by excessive removal of a species that is maintaining the ecosystem in balance.
Examples shown in Figure 5.2 (after Pinnegar et al. 2000) are: (A) the basic concept of a three-level trophic cascade involving plants, grazers and their predators; and specific examples of trophic cascades for: (B) Northeast Pacific kelp ecosystems, and the sensitivity of algal cover to predation on sea urchins by other food web components; (C) South African mussel and macro-algal equilibria affected by over-harvest of rock lobsters, showing also a reciprocal predator-prey relationship between rock lobsters and whelks (depending on their relative densities); and (D) a typical Caribbean coral reef ecosystem, where overharvesting of reef fish can lead to urchin explosions and decimation of algae, or where either predation on sea urchins or the disease epidemics they are prone to, can lead to coral overgrowth by algae. From all these examples, one can deduce the facility with which overharvesting of a keystone component (Paine, 1994; Power et al., 1996) of a food web can lead to major changes in productivity and holding capacity, if not to a transition to a quite different assemblage of species.
Figure 5.1 An example redrawn from Pinnegar et al. (2000) illustrating how overfishing sea urchins may tip the ecosystem from dominance by large algae to coralline barrens with minimal cover. (+ve implies a reinforcement of the box pointed to).
Figure 5.2 Examples of trophic cascades involving invertebrate resources (modified from Pinnegar et al., 2000). [Key: f - negative effect of fishing on population size pointed to; p- similar effect due to predation; g - similar effect due to grazing; +ve and -ve: positive and negative effects of the factor at the base of the arrow on the box pointed to].
Dye, Lasiak and Gabula (1997) noted the difficulties in achieving recovery of depleted stocks of the brown mussel in South Africa resulting from the ever increasing level of exploitation of the species which resulted from "political and logistic problems stemming from law enforcement". Mussels recruit preferentially to mussel beds, and where these have been depleted, algal cover replaces them. Dye, Lasiak and Gabula (1997) estimate that the rate of natural recruitment of mussels to substrates is very low and recovery may take up to eight years. They suggest that active enhancement be implemented in conjunction with rotational cropping as the best management strategy. Schoeman, McLachland and Dugan (2000) conducted a community-level, short-term and manipulative experiment involving a simulated anthropogenic disturbance, directed to address the potential effects of harvesting Donax serra on the macrofauna community of sandy beaches of South Africa. Some evidence suggested harvesting effects on the structure of the macrofauna, although the impacts appeared temporary because of the impossibility of controlling morphodynamic variables such as beach face slope and tidal rhythms. The high dynamics of the intertidal environment obscures the results of this kind of experiments because the effect of harvesting is not easy to assess over the short-term.
From South American experience in experimental shore ecology, Castilla (1999) notes that humans have high impacts on coastal and shore ecosystems through uncontrolled shellfish gathering, and that these impacts often occur in the sequence: (1) habitat transformation, fragmentation or destruction; (2) introduction of exotics or extinction of native species; (3) resource depletion; (4) food web modifications or "trophic cascades" as a result of 1 to 3. Through the institution of experimental "no-take" areas, he was able to observe and quantify some of these transitions. Often impacts from removal of grazing species led to increase algal cover. The exclusion of humans from rocky shores ("Reserves") in Chile allowed effects of handpicking and fishing (diving) on shellfish abundance and community elasticity to be tested (Moreno, Sutherland and Jara, 1984; Moreno, Lunecke and Lépez, 1986; Moreno et al., 1987; Castilla and Durán, 1985; Oliva and Castilla, 1986; Castilla and Bustamante, 1989; Durán and Castilla, 1989). Experiments in Central and Southern Chile demonstrated that humans, as specialized top predators, constitute the key factor ("capstone" species, sensu Castilla, 1993), altering exploited and unexploited benthic coastal populations, and generating ecological cascading effects that affect the structure and functioning of communities (Castilla, 1999, 2000). Varying extraction intensity on species of different trophic levels may translate into different community structures, thus enhancing the identification of linkages and showing the strengths of ecological interactions. This information was used to understand system elasticity and to translate ecological knowledge into management strategies (Castilla and Defeo, 2001).
Tegner and Dayton (2000) reviewed the effects of fishing in kelp forest communities organized around the primary productivity and physical structure provided by members of the Laminariales. This ecosystem supports a variety of fisheries, including harvesting of the kelp itself for alginates. The authors showed that sea-urchin grazing affects the abundance of both urchins and kelps, and the association of exploitation of various urchin predators with destructive levels of urchin grazing usually leads to cascading implications for other species dependent on the productivity and habitat provided by the kelps. Competition between abalones and sea urchins also affects some kelp communities. These ecosystem-structuring processes are also impacted through the ecosystem effects of fisheries for predators, abalones, sea urchins, and kelps. The authors suggested that no-take MPAs may be the only way to determine the true ecosystem effects of fishing.
Steneck et al. (2002) also reviewed the conditions in which kelp forests develop globally; and where, why and at what rate they become "deforested", as well as describing trophic cascades affecting different members of these communities. Overfishing of highly valued vertebrate top predators often lead to increases in herbivore populations and consequent kelp deforestation. This has had profound and lasting impacts, leading to species-depauperate systems in e.g. Alaska and the western North Atlantic. Continued fishing down of coastal food webs has resulted in shifting harvesting targets from apex predators to their invertebrate prey, including kelp-grazing herbivores. The recent global expansion of sea urchin harvesting (Andrew et al., 2002) has allowed kelp forest biomass to increase significantly. Shifts from fish- to crab-dominance caused by the absence of top predators in some places have occurred in coastal zones of the United Kingdom and Japan. Fishing impacts on kelp forest systems have been both profound and much longer in duration than previously thought. In other places, the large-scale removal of predators for export markets increased sea urchin abundances and promoted the decline of kelp forests over vast areas. The authors concluded that management should focus on minimizing fishing impacts and restoring populations of functionally important species in these systems.
Shears and Babcock (2003) reported that between 1978 and 2001 benthic communities in the Leigh Marine Reserve shifted from being dominated by sea urchins Evechinus chloroticus to being dominated by macroalgae, as a result of a trophic cascade resulting from increased predator abundance at reserve sites. Reserve sites had lower urchin E. chloroticus densities and a reduced extent of urchin barrens habitat, with higher biomass of the two dominant algal species (Ecklonia radiata and Caipophyllum maschalocarpum). At reserve sites, E. chloroticus was completely absent by 2001. Predation of gastropods and limpets on sea urchins at reserve sites are thought to be at least partially responsive for changes in community structure.
Control of predators has been frequently mentioned as part of shellfish management programmes, and could be considered in the development of an enhancement plan. One of the suggested implications for enhancement programmes is that an unutilized niche in the food web or pyramid may exist, that can be filled by enhancement experiments (Figure 5.3). Experimental removal of competitors and predators might then be used to promote stock enhancement and allow time for adaptation if this is feasible and proves cost effective, and could diminish predation effects. The response of the benthic community will depend on the strength of interaction between species, as well as on the functional role of the species to be enhanced. Keeping in mind this multispecies framework, the returns from enhancement operations in theory could be maximized by deliberately overexploiting potential predators/competitors of the target species.
Contrary to this perspective, although control of predators or culling of a dominant competitive species has often been suggested as a way to enhance a targeted population, this strategy has not been a clear success in many cases, though circumstantial evidence from fisheries landings does suggest that in some cases, a decline in predators has had some positive impact on prey landings. The benefits from this operation have usually been moderate however, due to compensatory increases in other predators or competitors. Contrary to this perspective, it has been argued that these experiments be performed to promote stock enhancement under the hypothesis that increasing the availability of limiting resources (e.g. food and space) could promote competitive release or a diminution of predation effects on an economic valuable species (Carr and Reed, 1993). Experimental removal of competitors and predators could be used however, to evaluate the strength of, and better understand the ecological indirect and direct effects on the population dynamics and structure of a harvested population (i.e. side effects and cascade effects).
Figure 5.3 One of the suggested justifications for enhancement programmes is that an unutilized niche in the food web or pyramid may exist, that can be filled by enhancement experiments. This diagram (modified from Masuda and Tsukamoto, 1998) illustrates the concept, and may be valid where overexploitation has removed a keystone species from the ecosystem, which risks radical change in its absence. In most cases however, the results of such interventions on destabilized trophic pyramids are likely to be unpredictable.
Experimental and manipulative approaches to ecosystem management have so far not had a high level of success, but seem worth pursuing further where the ecosystem is sufficiently simple to allow some predictions to be made. Such approaches would certainly seem to provide a more reliable means of guiding community structure to a desired state than basing management decisions solely on the predictions of a group of separately viewed single species models (Sainsbury, 1988: p. 361). However, experimental manipulations must be executed first in small experimental units, because the complex dynamics in multispecies assemblages generally precludes a firm forecast of the ecological outcome of a single manipulation of species abundance (Sainsbury et al., 1997; Castilla, 2000). Even in rocky intertidal communities where trophic linkages are among the best understood of any assemblage, relatively little is known about the effects of massive removals.
These alternatives are risky in nature, because ecosystem linkages are not always predictable. Often informal experiments performed on small experimental units are useful, but the complex dynamics in multispecies assemblages usually precludes a synthetic forecast of the ecological outcome of a single manipulation of species abundance. Nevertheless, some positive results were obtained both for shellfish and finfish concerning competitor and predator control. These provide some cautious support for predator control as an ancillary management tool in some circumstances as an aid to enhancing stocks (see Cowx, 1994; also Northcote, 1995 for examples from finfish populations).
In general, starfishes are commonly identified as keystone species in intertidal invertebrate communities where artisanal shellfisheries take place, because small changes in abundance (growth and survival) of this predator may lead to disproportionately large impacts on community structure. It is thus expected that massive removals of these key species could be useful for stock enhancement. In an experimental and manipulative approach intended to diminish predation rates and favour enhancement. Kristensen and Hoffmann (1991) showed that predation by the starfish Asteria rubens on Mytilus edulis reduced the density of this mollusc from 3 500 to 1 800 ind·m-2. Even though the predation was not concentrated in this case on the culture plots, it was clearly a potentially important source of mortality during enhancement operations (see also Orensanz, Parma and Iribarne, 1991). Natural restocking experiments conducted in Chile (Castilla, 1994), and the implementation of natural reserves and the concession of exclusive use of some grounds to fishers, led to predator control of subtidal populations of those organisms which are considered as competitors or predators of those species with the highest market economic value (Castilla et al., 1993). These authors reported the removal of almost seven tonnes of starfishes and omnivorous grazers (e.g. the black sea urchin Tetrapygus niger) in a local cove, where they were perceived by fishers as dangerous for early stages of the muricid gastropod Concholepas concholepas; one of the most important exploited invertebrates in Chile. This kind of massive removal should be approached with caution however, due to the possibility of other trophic linkages, (such as the release of predatory control on competitors of shellfish for example). The dramatic effects of this kind of disturbances can be unpredictable and even negative for the targeted species in the long-term (see Foerster, 1954 for a well-illustrated case for salmonid populations). Carefully designed field and laboratory experiments are required to rigorously test hypotheses about the effects of predators under realistic conditions, and also to predict how species interaction strengths will change in response to the magnitude of fishing effort exerted (see Castilla et al., 1993 for examples). The same conclusion was reached by Schielbling (1996), who evaluated the effects of predators on the distribution, abundance and behaviour green sea urchin Strongylocentrotus droebachiensis: the dominant grazer in the rocky subtidal zone in eastern Canada, whose abundance largely determines the structure and dynamics of the coastal ecosystem. This author also evaluated the potential for predatory control, and concluded that the lack of knowledge about predator-prey interactions in this system precludes any generalizations about the role of predation in regulating sea urchin populations.
Removal of predators has been reported as very successful in protecting restocked juveniles of the Pacific scallop Platinopecten yessoensis in Japan (Ventilla, 1982). This strategy has also used in a large experimental enhancement trial in France with Pecten maximus, but requires further investigation (Lake, Jones and Paul, 1987). Removal of predators has been documented by Brand et al. (1991) to enhance pectinid populations in the Isle of Man. Potting of lobsters and crabs is permitted within small experimental areas in which the stock is being enhanced by transplantation of spats or cultivated juveniles, since these are potential predators of scallops.
Although proving a direct effect of predator removal is problematic from fisheries landing data, a number of circumstantial cases have been documented where prey species increased in productivity as groundfish have declined. Caddy (1981) documented the increase in octopus landings off West Africa after sparid stocks were fished down, and Caddy and Rodhouse (1998) suggested similar mechanisms for a number of squid resources following groundfish catch declines. Although a specific linkage has not been documented, the unprecedented increases in lobster landings in the Northwest Atlantic seem to coincide with the collapse of groundfish stocks there. This case is still ambiguous despite a growing amount of circumstantial evidence. Elsewhere, for example, Kruse and Zheng (1999) found little evidence to suggest groundfish predation or competition provides an overriding explanation for crab fluctuations in the eastern Bering Sea.
When considering the effect of the collapse of Canadian groundfish stocks early in the 1990s as it affected Newfoundland waters, and especially invertebrate fisheries, Power and Newlands (1999) found a slow decline in mean trophic level of catches since 1900, but a most marked collapse between a mean trophic level of 3.5 and 2.8 between 1988 and 1995, accompanied by a shifts in targeting from groundfish to a higher proportion of invertebrates in catches. This may in part have been a result of changes in market demand, but the reduction of predation on the invertebrate catch increases observed cannot be discounted. The net result for eastern Canada from collapse of groundfish stocks is that a considerable degree of the resulting shortfall in landed values has been made up by increased landings and the high unit values of invertebrate resources.
Many municipalities in New England, USA, maintain public shellfish stocks, in part by using hatchery seed. A survey of 68 municipal managers responsible for these programmes (Walton and Walton, 2001) estimated annual seed loss averaged 44 percent, but survival to market was between 25 and 49 percent. Predation was viewed as the major source of loss, with the green crab, Carcinus maenas, the main threat, but there was a division of opinion on the effectiveness of predator trapping to reduce losses, although further studies were called for. Mark-recapture experiments were used by Shepherd (1998) to estimate survival of ages 2-4 year cohorts of two Australian abalone species. Natural mortality at 2-8 months was density-dependent, but from eight months to four years, was lower, and independent of density. Predators were mostly crabs and wrasses. A related issue is the importance of shellfish for sea birds, and Jennings, Kaiser and Reynolds (2001) described the effect of disturbance by shellfish harvesters on mortality of sea birds such as oystercatchers.
The large number of overfished shellfish stocks, as well as the indirect negative effects of fishing gear on marine ecosystems, confirm that management has often failed to achieve sustainability. A more holistic approach incorporating interspecific interactions and physical environmental influences would contribute to restoring shellfish populations (Botsford et al., 1993). Potential side effects of introducing hatchery stock to the benthic community may include multispecies interactions for example. Hatchery-raised shellfish, when transplanted to natural habitats, may be more susceptible to predation and show different behaviour patterns from native juveniles (Schiel and Welden, 1987; Schiel, 1993). Increased mortality rates may thus occur before the organisms become established in the natural habitat.
As noted in Chapter 3, drastic declines of the surf clam Mesodesma donacium in Peru after an ENSO event enabled the increase in abundance of subordinate competitors for food and space, the suspension feeders Donax peruvianus and Emerita analoga. This suggests potential interspecific interactions because of competitive release of resources by formerly dominant members of the faunal community becoming depleted or sub-dominant. The fishery closure for 32 consecutive months, from April 1987 to November 1989, in the yellow clam Mesodesma mactroides of Uruguay (see Chapter 3) was also used to investigate the effects of fishing activities on the demography of the yellow clam and on the sympatric suspension feeder, the wedge clam Donax hanleyanus. Markedly different effort levels over the long term generated major changes in the abundance and population dynamics of the wedge and yellow clams beyond the effects of exploitation; thus highlighting the ecological impacts of humans in the ecosystem, as extractors and as a source of physical disturbance. Abundance of the sympatric unharvested wedge clam D. hanleyanus rose steadily throughout the fishery closure. However, this could be explained as a monotonically decreasing exponential function of yellow clam density, i.e. what was effectively an interspecific stock-recruitment relationship, or in other words, an exclusion effect. In 1989 and 1990, during and immediately after fishery closure, both species occurred at their highest observed densities, suggesting that wedge clam recruitment was affected by the amount of fishing on yellow clam (Defeo and de Alava, 1995; Defeo, 1998). The spatio-temporal abundance of wedge clams was inversely correlated with fishing intensity over M. mactroides, suggesting that incidental damage (broken shells) and physical stress produced by sediment disturbance during harvesting were significant causes of mortality. Later information (1993-2002) confirmed the above trends, with an inverse relationship in abundance emerging between the two suspension feeders. In fact, anthropogenic disturbance in soft-sediments can lead to changes in substrate penetrability (Probert, 1984; Peterson and Black, 1988; Wynberg and Branch, 1992, 1994), which restricts the movement of burrowing organisms and increases mortality (Peterson, 1985). Hypotheses tested on the population dynamics and demography of exploited and unexploited bivalves by manipulating fishing effort, showed that human activities and endogenous density-dependent factors play important roles for exploited sandy beach molluscs (Brazeiro and Defeo, 1999; Lima, Brazeiro and Defeo, 2000).
Figure 5.4 Long-term fluctuations in abundance (ind per stripe
transect: ind·m-1) of adults of M. mactroides
(
) and D. hanleyanus
(
) in Uruguay.
Positive interactions of associated species on the target species may also occur, and commensalism, mutualism and facilitation were indicated as important processes in soft sediments. Ahn, Malouf and Lopez (1993) showed that dense assemblages of the gem clam Gemma gemma enhanced settlement of four to five-day old hatchery-reared larvae of the commercially important bivalve Mercenaria mercenaria. A positive interspecific adult-larval interaction was demonstrated, and even the presence of empty shells of G. gemma enhanced settlement of hard clam spat. Peterson and Black (1993) manipulated local densities of two bivalves of Katelysia in a Western Australia lagoon characterized by exceptionally high bivalve abundances. They showed that competition occurs only weakly, and that the addition of sufficient numbers of clams and dead shells filled with sand and implanted in the living position in the sediment mitigated mortality at low density. Thus, this particular positive interspecific interaction could be used to enhance settlement by simply scattering empty shells on the sediment to increase the quality of the settlement habitat. Indeed, as a possible alternative to costly restocking/seeding, habitat enhancement using artificial or natural materials added to the habitat may provide additional space for settlement and increase or concentrate stocks, possibly at a lower cost than rearing juveniles onshore (see Chapter 6).
Introduction of exotics has been criticized from the strict perspective of conservation of ecosystems, and there has been little practical evidence that negative consequences of anthropogenic effects can be reversed, or that the introduction of exotic species, for example in ships ballast, can be easily controlled. The accidental introduction with seed or cultch of a variety of species of exotic predators and troublesome commensal organisms could however have been avoided by careful quarantine procedures, but these procedures have rarely been implemented. An appropriate precautionary approach to species introductions is described in FAO (1996), based in part on the ICES "Code of Practice on the Introduction and Transfer of Marine Organisms 1994", and is described in Annex A to the above-cited FAO report. Although this text does not enter into detail on this topic, which is dealt with in the above FAO publication, inevitably we touch upon the implications of these processes in enhancement programmes.
A large number of organisms, some useful (such as the Manila clam, Tapes philippinarum and the Pacific oyster, Crassostrea gigas), but most of them harmful - (such as the slipper limpet Crepidula fornicata, various predatory gastropods, and a variety of seaweeds such as Sargassum weed and Laminaria japonica), are thought to have been introduced into European waters on shells or cultch during exotic shellfish introductions. The development of predictive models of the impacts of species introductions of commercial importance is a major issue for stock enhancement programmes. Introduced species could strongly affect recruitment of other taxa, and these effects varied between spatial scales and sites and among taxa (e.g. Orensanz et al., 2002 and references therein). This implies that development of predictive models of the effects of similar invaders will require detailed knowledge of the responses of individual species that comprise local assemblages (Holloway and Keough, 2002). Changing climate might facilitate invasions by favoring introduced over native species, thus shifting to dominance by nonnative species (Stachowicz et al., 2002). This has shown for sessile invertebrates, which are of major relevance in the context of this paper.
In the marine environment, apart from some estuarine introductions such as striped bass on the west coast of North America, most cases of introduced species that end up being the subject of enhancement activities were originally introduced accidentally. This is true for example of manila clams introduced with Japanese oysters into west coast environments of North America and elsewhere, and the predatory gastropod Rapana sp. which now forms the basis for a significant fishery in the Black Sea. There are relatively few cases where new species have been deliberately introduced, and there is a great reluctance to do so, despite the fact that for some species such introductions could have positive economic repercussions if successful. The repercussions may however be significant: thus the immigration of lizard fish into the Mediterranean from the Red Sea through the Suez Canal has led to a new Mediterranean fishery off Israel for this species, but has constrained the hake stock in the eastern Mediterranean to a more restricted bathymetric range than formerly.
Another paradoxical application, which is the converse of enhancement, is the use of predators to control unwanted invasions of molluscs such as zebra mussels Dreissena polymorpha, which occur in large biomasses and lead to engineering problems in waterworks and pumping infrastructures. Molloy (1998) suggests that microbes represent the most promising biological control agents for these pests, given that predators are often not specific enough in their feeding habits. Practical experience in fact shows that exotic predators introduced for control purposes may consume species of importance to man.
Empty niches generated by kelp deforestation in the western North Atlantic as a result of population increase of herbivore populations were filled by introduced algal competitors, which carpet the substrate and threaten future kelp dominance. Other non-native herbivores and predators became established and are now dominant components of this ecosystem (Steneck et al., 2002).
Harmful algal blooms are more frequent than previously due to high nutrient levels caused by coastal pollution, and while shellfish populations may reduce the density of blooms, the possibility of harmful effects is a real one (see Babaran, Espinosa and Abalos, 1998). In choosing an area for enhancement, any historical records of harmful algal blooms that occurred in the past in the area need to be investigated. Accumulation of toxins resulting from algal blooms and viral diseases may produce massive mortalities that must be addressed. Stock enhancement programmes for beach clams could be constrained by the accumulation of toxins associated with algal bloomsand can cause mass mortalities of clams or render them unsafe for human consumption. The increasing short-term occurrence of this phenomenon makes commercialization and immediate consumption highly risky in areas where these blooms regularly occur (McLachlan et al., 1996). In this context, massive mortalities of suspension feeders in South America almost decimated yellow clam populations all along thousands of km of coast during the 1990s (Fiori and Cazzaniga, 1999). The occurrence of cold atmospheric fronts that accumulated high concentrations of dinoflagellates in the surf zone, have been invoked as a cause of massive clam mortalities along the Brazilian coast (Odebrecht et al., 1995), as well as being associated with specific viral diseases.
Arntz et al. (1987) demonstrated massive mortalities in the recreationally and commercially harvested sandy beach bivalve Mesodesma donacium in Peru, as a response to the strong ENSO climatic phenomenon, and this formerly dominant member of the macrofauna disappeared following ENSO episodes due largely to an increase in sea surface temperature. The intertidal black abalone Haliotis cracherodii has experienced mass mortalities along the coast of California, USA, since the mid-1980s due to infection by a pathogen that leads to a fatal wasting disease called "withering syndrome": the foot of the abalone atrophies until it can no longer adhere to the substratum (Friedman et al., 2000). The presence of the pathogen, and warm water conditions associated with El Niño, may accelerate the development of withering syndrome and the rate of decline of black abalone populations. Anthropogenic disturbances, such as the discharge of heated water or global warming, may thus increase the incidence of this fatal disease (Raimondi et al., 2002).
One additional management problem related to the enhancement of stocks relies on the effect of technological externalities produced by contemporary fisheries for other species, or even on interactions between coastal commercial and sport fisheries in the same area. Once again, legislation must be implemented to avoid confrontational competition between fishers for different target species, or with different gears, probably by some form of zonation by gear type usage, possibly in addition to separate zones for fishers and other users of the coastal zone. This is likely to be especially relevant in cases where enhancement is conducted in coastal nursery areas that are also attractive grounds for other fisheries.
Increasing fishing capacity and introducing technological changes aimed at increasing fishing power may have undesirable effects on the habitat used for stock restoration programmes. These interactions have reduced our ability to calibrate effective fishing effort and mortality, and hence, introduced a major uncertainty into estimating population variables. Complementary studies should be carried out on a case-by-case basis to identify less damaging new technologies that have a reduced impact on the environment. Well-designed experiments provide useful information on the impacts of the technologies on particular habitats, as well as on the species being enhanced and its biological community. For example, maximizing the productivity in many scallop fisheries needs to consider the effects of indirect fishing mortality and disturbance of the seabed through dredging (see example by Bull, 1994). This should lead to a regulation of fishing intensity in these areas, and will contribute to establishing an optimal periodicity of a rotational approach. One of the most important artisanal shellfisheries in Argentina since the 1940s is based on the extraction of the clam Mesodesma mactroides (Olivier et al., 1971). The fishery collapsed because the use of tractors (instead of manual collection as elsewhere in South America). This markedly increased fishing power and negatively affected the physiographic characteristics of the beach substrate (Olivier and Penchaszadeh, 1968).
In sequential fisheries, in which juveniles and adults are spatially segregated and harvested with different techniques, technological and ecological interdependencies occur, affecting both population components and generating externalities (Seijo, Defeo and Alava, 1994). A typical example is the surf clam Mesodesma donacium fishery of the intertidal and shallow subtidal of Chilean sandy beaches (Defeo et al., 1993). Intertidal juveniles are manually harvested during low tides, whereas free and semi-autonomous hookah divers operating from artisanal wooden boats, harvest subtidal adults. In this example, stock enhancement programmes should be carefully planned to minimize potential problems and risks. The most appropriate individual size for restocking should be evaluated, taking into account which population component is easier to handle and transplant and less costly to enhance.
The impact of fishing on the structure of an ecological community might depend also on the number of species that are marketable, and on the corresponding unit prices. For example, the net revenues derived from fishing could be maximized by deliberately overexploiting a less valuable but dominant species, in order to increase production of a subordinate competing species with higher market value. Existing market conditions could also lead to depletion of the most commercially profitable species to such a low population size that the fishery for an interdependent stock becomes economically unfeasible. This suggests that humans behave as generalist predators when closed seasons/areas and other regulatory measures are imposed on harvests of traditional resources (Smith and McKelvey, 1986). This has been frequently observed in shellfisheries (Bustamante and Castilla, 1987; Orensanz et al., 1998; Cabrera and Defeo, 2001). This issue has also been addressed in a text on fisheries bioeconomics published by FAO (Seijo, Defeo and Salas, 1998).
