Successful stock restoration or enhancement requires harvest controls but also demands and attention to human impacts on the habitat. Reducing exploitation alone on the stock being restored will not be effective if critical habitat has disappeared. Figure 6.1 points to the role of habitat as an important sequential constraint. The use of habitat restoration to improve yields of marine coastal species does occur locally, but often has to be searched for in the literature on marine ecology, the role of marine parks, etc. Habitat restoration has on occasions been considered more feasible than stock enhancement per se. Hilborn (1998) for example, points to the low success rate in economic terms of most stock enhancement exercises, although most cases he refers to relate to finfish. Nonetheless, we agree with his suggestion that stock enhancement should be compared with alternatives such as habitat protection, fishery regulation and stricter enforcement before embarking on potentially costly and uneconomic large scale operations without prior experimentation. In this section we consider the first of these options. Performing a formal cost-benefit analysis on habitat restoration is feasible where artificial structures are added to the water to improve holding capacity, but basic pilot stage applications that allow us to establish feasibility, would seem required first.
Figure 6.1 Showing how different stages of the life history (in this case of a crustacean such as a lobster) may be carried out within different habitats and depths.
Perhaps the Multispecies Virtual Population Analysis (MSVPA) experiment carried out for North Sea fish species (Sparholt, Larsen and Nielsen, 2002) provides the most eloquent argument against automatically assuming that juvenile enhancement programmes, sea ranching or restocking programmes will necessarily be effective. This extensive programme of stomach content analysis revealed that juvenile fish (as for juvenile shellfish) have a generally high vulnerability to predation, and extremely high juvenile mortalities apply in the first 1-3 years of life. While this experiment suggests that any reduction in juvenile mortality of commercially important species could pay major dividends, evidence will be presented that there may be a need to mitigate the impacts of habitat-mediated bottlenecks.
In the United States, the Magnuson-Stevens Act requires fisheries managers to describe essential fish habitat, and to minimize the impacts on this by fishing. "Essential Fish Habitat" (EFH) is defined as "those waters and substrate necessary for spawning, breeding, feeding or growth to maturity", and requires to be defined for all species under management. Mapping such areas using GIS techniques is now an essential component for deciding on closed areas, leases, enhancement areas, and in assigning priority use within an Integrated Coastal Area Management (ICAM) activity. This may involve remote sensing (see Rubec et al., 1998). Figure 6.2 illustrates the complexity of decision-making in the coastal zone, and the fact that many other activities impinge on shellfish production, and will need to be reconciled in a multi-disciplinary way, involving all "players" or stakeholders in the coastal zone.
Figure 6.2 attempts to map some of the activities, causes and effects of human activities, starting with those associated with runoff from the adjacent freshwater catchment basin (upper left). This has effects listed on the right (e.g. increased red tides, shellfish toxicity etc) which impact the human activities on the right (e.g. aquaculture, fish processing etc), The fishery and fish culture activities in the box with dotted lines is thus seen as small "downstream" components vulnerable to these non-fishing anthropogenic effects.
Limitations of habitat may be important bottlenecks for some invertebrate populations, thus Scheding et al. (2001) found that the presence of sandy bottom used by ovigerous females to bury themselves in aggregations during egg incubation, may be a limiting factor for adult Dungeness crab abundance in Alaska. Experience suggests that benthic and demersal organisms of a wide range of taxonomic groups and species often pass through primary and secondary habitats in the course of their life histories, thus, American lobsters may first occupy crevices in cobble bottom areas in a completely cryptic life stage until 5-10 cm in length (Wahle and Steneck 1991), when they migrate to burrows under stones or construct burrows on more sedimentary bottom. Similarly, Palinurid lobsters settle first in finely branching substrates such as red algae or mangrove roots, before migrating to sea grass beds and later to caves under coral patches further offshore. In both cases, the extent of these critical juvenile habitats may be the bottleneck limiting overall production, given that habitat and food supplies available to older stages may not be limiting.
The giant scallop Placopecten magellanicus and various species of Chlamys have been described as settling on bryozoa before attaching with a byssus to shell gravel, even inside the umbones of dead adult scallops. Similar primary and secondary settlement episodes, and the characteristic type of bottom conditions that identify them, have also been described for Mytilus edulis (see e.g. Bayne, 1964, 1976). Such primary and secondary "nursery" habitats have been described for a wide range of invertebrate species, and their often restricted occurrence makes one wonder if this poses a limit to recruitment, or in other words forms a "bottleneck" in the production process. Various kinds of bottlenecks have been proposed, from the more general term "demographic bottleneck" which covers all forms of restriction on survival of a year class, to more specific "shelter-limited bottlenecks" which imply a structural limit, or "trophic bottlenecks" where food is the limiting variable. In fact Walters and Juanes (1993) show that it is the limitation of food adjacent to cover that requires organisms to be "risk-takers" and feed outside of cover.
Figure 6.2 Showing a framework for policy development in the coastal zone, the socio-economic context for integrated policy formulation, and interactions of a variety of environmental factors. The interactions between different social users of the coastal environment, including shellfish production are shown. (after Caddy and Bakun, 1995).
A feature first observed by Morse et al. (1985) as a consequence of the fractal structure of most complex surfaces in nature, is that on a complex surface such as that of aquatic vegetation, there are "more surfaces available for small (arthropods) than large ones". Caddy and Stamatopoulos (1990) generalized this observation to show that the fractal nature of habitat structure results in small crevice sizes being much more common than large ones, hence migration or mortality by predation are the two options that await those juveniles displaced from crevices of the appropriate size while searching for much less abundant larger crevices.
Mya arenaria was introduced to Grays Harbor (Washington) during the 1880s. Palacios, Orensanz and Armstrong (1994, 2000) showed that the maximum size of members of the current population is much smaller than of clams found dead in situ from preceding decades, and which nowadays form extensive "death assemblages". Palacios, Orensanz and Armstrong (1994) concluded that extinct clams grew faster and lived longer because they occupied the best habitats available. After an extensive mass-mortality episode between 1895 and 1897 that resulted in the formation of the deposits, the population has never rebounded into its prime habitat, in spite of potential seeds being regularly available. They also showed that Dungeness crab larvae settle preferentially in these shell deposits, where the abundance of 0 + age juveniles is orders of magnitude higher than on the adjacent flats. They hypothesize that predation by juvenile crabs is the main factor that limits clam recruitment (see also Iribarne et al., 1992).
The point which can only be touched on here is that where such critical habitats occur, their holding capacity will to a significant extent determine the size of the new recruiting age class to the population. Thus there will be no point attempting stock enhancement if the existing population size will subsequently be restricted by some form of bottleneck. As noted, another implication of this type of phenomenon is that even if there may be adequate food for many more adult animals than are actually present, these population levels are unlikely to be realized if there is a bottleneck at or following recruitment. This emphasizes the importance of ensuring that primary and secondary habitat types and cover characteristics are not degraded by environmental influences, and that suitable habitat is provided in extensive culture or where enhancement of depleted wild populations is underway, especially in the presence of predators for the species in question.
There have been many criticisms of hatchery rearing and release programmes aimed at stock enhancement, and as we have noted, the record of success is rather spotty, especially for finfish, but should not prevent careful consideration of this mechanism in certain well-studied situations. Despite the five conferences held on the issue in recent years (Grimes, 1998), there has been a relatively minor focus on related marine habitat issues, and in particular, cover and habitat complexity. A more general consideration is the widespread occurrence of density-dependent mortality in natural systems, which ensures that increases of population size above some carrying capacity will be rapidly reduced. In this sense, Heppell and Crowder (1998) suggest that before considering stock enhancement, the existence of habitat constraints should be checked for. Does the environment contain sufficient critical habitat needed for the life history stages introduced for example? They also stress the need for harvest controls to be in place. An example for a finfish in Florida nearshore waters, the gag grouper (Mycteroperca microlepis), showed that the existence of adequate sea grass beds is a precondition for restoration (Koenig and Coleman, 1998). For another finfish species important to anglers, the red drum, Grimes (1998) notes there is little point in attempting stock enhancement if density-dependent processes in the early life history result in high loss rates of juveniles. At the same time (and this defines one particular focus for any enhancement programme), in this case, recovery of as few as one percent of stocked fish need to be recaptured by high value sports fisheries for enhancement to be considered successful. Achieving an increase in recruitment in the range of five and ten percent by either stock or habitat enhancement would then seem to be a worthwhile and possibly achievable target. Whether this would be cost-effective as questioned by Hilborn (1998) still remains an important question, but in the case of restoration of a stock that has disappeared from its former range, such considerations must take second place to first establishing feasibility, and second, deciding what the restoration of a train of utility values into the indefinite future is worth? If the issue is to restore a train of benefits that has been lost, or otherwise would not continue into the future, and if the population would then be self-sustaining, restoration may be a feasible objective.
Habitat replacement or rehabilitation are less ambitious interventions than aiming for complete ecosystem restoration or rehabilitation, which are goals that are likely to be difficult or in some cases, impossible to achieve, and certainly more costly, and would involve whole system manipulation. Replacement or reclamation of damaged or degraded ecosystems constitutes interventions aimed at restoring economic productivity to a habitat that, due to human intervention, is currently unproductive. In this sense, a stock enhancement or replacement exercise may be a component of a more general environmental intervention aimed at improving habitat "quality", where quality is defined in the sense that it contributes to human welfare. Finally mitigation, the least ambitious type of intervention, aims to reduce the losses incurred due to ecosystem damage, and here again, shellfish enhancement schemes may play a part in a broader context, and is likely to be more feasible than, for example, restoring fish populations. Some definitions for these terms are provided in Table 6.1.
Table 6.1 Some human interventions on depleted or degraded ecosystems (after Bradshaw, 1996 and Cairns, 1994).
Rehabilitation or restoration
The action of restoring a thing to a previous condition or status
To bring back to a proper state (which may not be the original one)
To procure a satisfactory substitute in place of the original
To rectify or make good, places the emphasis on the process rather than the end point reached
To reduce the negative effects of change (especially habitat destruction) where these cannot be eliminated (or to soften the loss of a particular ecosystem
The importance of habitat for management of wild resources is made evident in Figure 6.3, based on Marshall (1996). This emphasizes the importance of characterizing the biological suitability of the habitat and the biological community within which the species in question is being enhanced. If the habitat has deteriorated, for example through anthropological stresses, and is biologically complex, there has to be less expectation of a successful result. Probably it will be necessary first to restore the ecosystem, to the extent possible, to a productive situation before beginning an enhancement procedure. Better still is to choose another area where these stresses or impacts are less pronounced.
Figure 6.3 An approach to assessing habitat suitability (after Marshall, 1996).
As an example of a rehabilitation or mitigation exercise, we note that included in the goals of the US National Research Council (1992) for national aquatic ecosystems between now and 2010, is to restore ten million acres of wetlands out of the 117 million impaired or destroyed since 1800. Such wetland areas are frequently very productive, and shellfish form key components of these ecosystems and should be harvested in moderation. It is worth noting here that wholesale raking of oyster beds in the Chesapeake Bay system removed these higher relief shell or cutch banks which formerly were "islands" of hard bottom above the Bay floor, and were of ecological importance to a range of species.
Cairns (1994) notes that: "although precise replication of predisturbance conditions will rarely be possible, achieving a naturalistic assemblage of plants and animals" (at the landscape level) "of similar structure and function should be possible in most cases", and comments that: "It is a sine qua non that ecological repair is preferable to neglect of damaged ecosystems". He remarks that it is not excluded that this restoration could be (in the case of wetlands) in areas which did not have them in the first place, as a way of building "ecological capital". He goes on to say: "Necessarily, local societies must have accepted the goal of restoration and cooperate, which in turn requires greater ecological literacy". If self-maintenance is to be achieved, the scale of restoration increases considerably, and ecological restoration is seen as "buying more time for human society to develop less threatening life styles".
Langton et al. (1996) suggested a prioritization of research questions as shown in Figure 6.4, which can equally be used when contemplating an enhancement programme. If the questions in the list of "prioritized research questions" cannot be answered in advance, it is important that priority work be done on resolving them before full scale enhancement begins.
Figure 6.4 A list of research questions proposed by Langton et al. (1996) prior to beginning any intervention involving natural systems.
The experience with shellfish enhancement procedures as implied earlier, has not been uniformly positive, and it is instructive to consider why. Figure 6.5 places enhancement at the summit of a sequence of human ecosystem interventions, which are perhaps more realizable from the bottom of the triangle than the top, and this especially applies to marine finfish stocks.
Figure 6.5 Illustrating a sequence of human ecosystem interventions, arranged from the bottom of the inverted triangle to the top according to their likely feasibility and cost. Shellfish enhancement may play a role in all of these activities, but the implication is that this is likely to be more feasible and relevant for uncontaminated ecosystems that have been otherwise impacted by human activities.
Figure 6.6 needs to be kept in mind when restoring a stock through stock enhancement, which is unlikely to be successful if the critical habitat that limits life history stages is not present, and if effort control does not allow parental stocks to build up so that natural population replenishment can occur. Trophic conditions are also important in providing overall basic requirements for food, but this alone is unlikely to be adequate in ensuring population build-up if the other two factors are not given priority. Mitigation of negative impacts is less demanding than reclamation or replacement, which in turn are less problematic than full restoration of the original ecosystem.
Enhancement, which in the strictest sense implies "improving on nature" would imply even more costs since an "enhanced" system is implied to be an improbable state differing from the equilibrium situation often characterized by the term "virgin conditions". In fact however, enhancement becomes more feasible when it is considered in the context of restoring an ecosystem after serious stock depletion, and is also appropriate for systems which are recovering from disturbance, where for example an unfilled ecological niche may be temporarily present. It is mainly in this sense that the term is used in this paper, so the apical position of enhancement in Figure 6.6 is probably anomalous.
Figure 6.6 Factors that need to be considered when restoring shellfish populations through stock enhancement include basic trophic requirements and special habitat requirements, the latter more difficult to satisfy, while full restoration will require an adequate spawning stock.
In some cases (e.g. for penaeid shrimp stocks, Figure 6.7), river outflow patterns have a major role in determining shrimp production, and drainage of wet lands, cutting of mangrove forests and the use of herbicides in coastal areas can negatively affect shrimp production.
Figure 6.7 Catches of the shrimp Penaeus notialis in Casamance river estuary as a function of rainfall index (after Le Reste, 1992).
Determining the effects of land usage on intertidal and estuarine resources is an important precondition to producing safe shellfish products. White et al. (2000) mention a multi-agency project that was carried out in an Integrated Coastal Area Management (ICAM) context, to determine the effects land use in the adjacent watershed had on shellfish closures. Bacterial data were monitored and indicated increasing loadings in runoff water, with especially high levels during storm events. Dye studies confirmed that bacteria would move through the watershed over a brief time period with negligible mortality. Low levels of bacteria were found during dry weather. Most contamination came from a nearby residential area with some malfunctioning septic tanks, but also from pets and wildlife. A mitigation programme was designed using GIS, and included education, and restoration of wetlands, automated storm water monitoring, and DNA tracking of faecal sources.
A general discussion of shellfish restoration activities in the context of ecological functions is provided by Coen and Luckenbach (2000), who mention some of the "ecosystem services" provided by shellfish beds, whose value is usually underestimated. These include contributions to the filtering capacity of the water column, benthic-pelagic coupling, a role in nutrient dynamics, increased suspended sediment deposition, and stablilization of bottom sediments. Very dense shellfish beds may have negative impacts due to fouling by pseudofaeces, and Kaiser et al. (1998) review some of the impacts caused by dense shellfish culture. Castel et al. (1989) find that densely stocked oyster beds elevated organic carbon levels in adjacent sediments, sometimes producing hypoxic conditions. Site selection preferably in areas of current flow is therefore important. However, considering that shellfish beds were certainly denser before human harvesting, such negative impacts of shellfish populations are in themselves, local effects; and under natural conditions would be less evident. One reason being that through accumulation of shell material, native shellfish beds were often raised above the sedimentary level of the estuary floor and hence cleaned by stronger local currents. Sparsis, Lin and Hagood (2001) even evaluated the feasibility of using juvenile giant clams to remove dissolved inorganic phosphate and nitrate from holding tanks of ornamental or food fish. The nutrients were removed by xooxanthellae in the mantle of giant clams in lighted periods, and all species tested except Tridacna gigas survived, and some grew faster in effluent than elsewhere, suggesting the possibility of using giant clams in polyculture.
Restoring whole ecosystems where the environment has changed due to human intervention or climate change is a much less certain prospect, and in this case, actions to mitigate the damage may be in order. An example from other than shellfish enhancement comes from the North American Great Lakes (Regier et al., 1988), where eutrophication and overfishing had destroyed the valuable deep water lake trout fishery, but subsequently established an abundant but low value resource of small forage fish. The introduction of a west coast salmon species which could prey on these forage fish completed the transformation of the system into an "exotic" pelagic ecosystem (where effectively no economic resources of surface waters existed before) and this now has high economic value for sports anglers. This example is not an apologia for habitat degradation, just that restoring the Great Lakes to their pristine condition was probably infeasible, and would anyway have meant radical changes to the life of several million inhabitants within the catchment basin of the Great Lakes, hence restoring a productive, if alternative, system appears the most feasible option available. For this reason, fully restoring the native lake trout population is a costly option compared with the current sports fishery for introduced coho salmon. This example of a successful mitigation of human impacts shows that the more general objectives of restoring habitat quality for a range of purposes (recreation, quality of life issues etc), will usually have to precede the restoration of a population of organisms of value to humans, and these may be different from those originally present.
Another example can be cited in the case of the Black Sea. At the start of the twentieth century this was a mesotrophic body of water with oxygenated waters on shelf areas, and a rich fauna of indigenous species (see Caddy and Griffiths, 1995). The eutrophication of the basin led to a cascade of ecological effects described in Zaitsev (1993). Seasonal anoxia of shelf areas and the littoral in summer led rapidly to the elimination of many indigenous species. Eutrophication coincided with the likely accidental introduction of two exotic species, the clam Mya arenaria, and a gastropod Rapana sp. both adapted to eutrophic conditions. This latter species, a predatory snail introduced from Japan with oyster imports, has acclimatized, and now supports a fishery of some 40 000 t/year. So far Mya is unexploited but is believed to support a very large biomass. The fate of a large subtidal population of the mussel, Mytilus galloprovincialis in the Northwest of the Black Sea is also instructive. This was the target of a Russian fishery, but subsequent events suggest it may have played a major role in controlling turbidity of shelf waters (Sorokin, 1993). With seasonal hypoxia, this deeper water mussel population has largely collapsed, and coinci-dentally, a subtidal population of a red alga, Phyllophora sp. in the same area disappeared, presumably due to poor light penetration caused by algal blooms. Again Sorokin (1993) deduced that both of these declines played a significant role in oxygenation of shallow shelf waters and reduction of suspended material, thus improving light penetration. This example illustrates that the services provided by bottom fauna and flora, including populations of filter feeders, have a significant role, in this case in controlling suspended sediments and phytoplankton and coastal eutrophication, apart from their importance for human food.
Hypoxic conditions can of course be detrimental to shellfish enhancement operations, depending on species tolerances. Diaz and Rosenberg (1995) found that commercial species varied considerably in their resistance to hypoxia; thus, bivalves Arctica islandica and Mytilus edulis were most resistant, the clams Mercenaria mercenaria and Spisula solidissima were intermediate, while benthic crustaceans, Nephrops norvegicus, Crangon crangon, Carcinus maenas, and Spisula solida, a clam typical of clean sand, were the most sensitive.
The effects of bottom-water hypoxia on the population density of the clam Macoma balthica was estimated using a survival-based approach by Borsuk, Powers and Peterson (2002). They used a Bayesian parameter estimation to fit a survival model to times-to-death corresponding to multiple dissolved oxygen (DO) concentrations assessed by scientific experts, and combined the survival model with a model describing the time dependence of DO. Under current conditions, the mean summer survival rate was predicted to be only 11 percent. However, if sediment oxygen demand (SOD) is reduced as a result of nutrient management, survival rates increased, reaching 23 percent with a 25 percent reduction in SOD and 46 percent with a 50 percent SOD reduction (Borsuk, Powers and Peterson, 2002).
Lenihan et al. (2001) tested the hypothesis that mobile consumers have the potential to cause a cascading of habitat degradation beyond the region that is directly stressed, by concentrating in refuges where they intensify biological interactions and can deplete prey resources. They worked on structurally complex, species-rich biogenic reefs created by the eastern oyster, Crassostrea virginica, in the Neuse River estuary, North Carolina. Bottom-water hypoxia and fishery-caused degradation of reef habitat induced mass emigration of fish, thus modifying community composition in refuges across an estuarine seascape. Moreover, oyster dredging reduced reef height and exposed the reefs located in deep water to hypoxia/anoxia for more than two weeks, killing reef-associated invertebrate prey and forcing mobile fishes into refuge habitats. High-density accumulations of refugee fishes on reefs in oxygenated shallow water depleted epibenthic crustacean prey populations. Thus, the interaction of reef habitat degradation through fishery disturbance and extended bottom-water hypoxia/anoxia caused oyster mortality and influenced the abundance and distribution of fish and invertebrates that use this reef habitat (see also Lenihan and Peterson, 1998). The authors concluded that physical disturbances can impact remote, undisturbed refuge habitats through the movement and abnormal concentration of refugee organisms that have subsequent strong trophic impacts. In this context, they highlight the implications of MPAs as critical refuges.
The upper Adriatic Sea, an area of predominantly fine bottom sediments acts as an "outer estuary" by receiving some thousand tonnes a year of nutrients from the very eutrophic Po river. Italian workers (e.g. Bombaci, Fabi and Fiorentini, 2000) have focussed on use of artificial reefs colonized densely by Mytilus as a way of making use of these high productivities and precipitating suspended material from the water in the pseudofaeces of mussels. The potential role of mytiliculture here is not only to provide considerable economic add-on food value, but also to act as a depurator of estuarine discharges and the precipitation of sediments and algae from the water column: a function of importance to bathing resorts in the Adriatic. These kinds of ecosystem functions that may be achieved through a shellfish enhancement programme deserve further economic analysis. The dramatic increase in shellfish production in the Mediterranean shown from FAO statistics (Figure 6.8), especially in the upper Adriatic and Gulf of Lions under the influence of the Po and Rhone river outflows, needs attention. As noted by Caddy (1993a) and de Leiva Moreno et al. (2000), European inland seas such as the Baltic, Adriatic, Mediterranean and southern North Sea, depending on their degree of enclosure and the extent of the catchment areas feeding them, have, become eutrophic to different degrees, under anthropogenic influences from adjacent catchment basins. Molluscan shellfish production to a significant extent has benefited from this situation, although issues of depuration and the control of transmission of disease vectors through untreated shellfish have also become important. Thus, macroenvironmental trends in an area provide an important context for the shellfish enhancement activities we have been discussing here. It may be noted that in terms of the production of animal protein, shellfish cultivation does not depend on resources of fish meal or agricultural products as for most (carnivorous) species used in marine finfish culture, and as we have noted, if used strategically, mollusc shellfish stocks can play an important role in locally enhancing water transparency and hence in restoring aquatic vegetation.
Figure 6.8 Illustrating the role of nutrients from river catchment areas and river plumes in enriching nearshore shellfish fisheries (after Caddy, 2000b).
The selection of adequate habitats is crucial for the development of enhancement programmes. Recognition of gradients in habitat quality is important in defining the extent of the area available for seeding. A range of habitat sites must be analysed to evaluate likely differences in seeding success according to habitat suitability. Consider for example that a seeding programme for an intertidal soft-bottom bivalve is started with the intention to colonize new areas and develop a new fishery. In this case, some critical variables detailed below will give useful insights as to which sites would be optimal.
Exposure (exposed-sheltered). The choice of an adequate habitat for seeding should be a trade-off between different factors acting simultaneously. For example, when considering intertidal oysters, the level of the intertidal chosen to seed is critical, because lower tidal levels are more susceptible to predation. Alternatively, higher tidal levels commonly have a higher degree of silting which increases mortality and lower growth rates. Quantity and quality of food might depend on exposure and the degree of turbidity. Some exposed sandy coasts could constitute semi-closed ecosystems in which high concentrations of surf phytoplankton occur in waters rich in nutrients and oxygen. These sites could be useful for enhancement operations of intertidal suspension feeders (including passive restocking, see Defeo, 1993b, 1996a). However, wave action and speed of currents could act as negative forces that could preclude spat settlement. Once again, a trade-off between these different factors must be evaluated. If a sandy beach mollusc is selected, the definition of beach morphodynamics will be critical, i.e. if it is reflective or dissipative (McLachlan et al., 1996).
Grain size preferences (fine - coarse). Settlement rates could be accelerated in the presence of a suitable substrate. For example, Tong, Moss and Illingworth (1987) reported that juveniles of the abalone Haliotis iris tend to settle in almost all cases associated with the encrusting algae Lithothamnium. A careful selection of sites for seeding can help reduce mortality rates. In some cases (e.g. oysters), ground selection according to consistency of the bottom can also reduce silting mortality. Shifts in habitats, especially burial by sand, in the abalone Haliotis iris (Schiel, 1993), led to high mortality rates and negative rates of return in the enhancement operation.
Other sediment features, such as face slope, substratum penetrability, sediment water content, texture and roughness, are also important agents defining settlement variation among sites. Space availability in proportion to the amount of spats to be seeded should also be assessed.
Knowledge of sediment preferences at different life stages helps the shellfish farmer use adaptive behaviour of shellfish in protecting them from abiotic (e.g. hydrodynamic factors, exposure) and biotic (e.g. predation) factors. By carefully choosing sites, within-site sources and levels of natural mortality due to abiotic and biotic factors can be minimized. Historically productive fishing grounds, which have high probability of recolonization and generally low mortality levels, could serve as potential sources for seed replenishment (Caddy, 1989b). The quality and quantity of food present, or added as supplements, can be critical if economic losses are to be avoided due to density-dependent processes. Food availability often depends on habitat quality, and mortalities will occur due to starvation if seeding is conducted in the wrong place (see e.g. Tegner, 1989; Kristensen and Hoffmann, 1991).
Substrate preferences for many crustacean and molluscan larvae are often rather specific: thus Stevens (2003) found that late larval king crab demonstrated preferences for structurally complex substrates and a low preference for sand where mortalities were higher. This illustrates the importance of leaving "biological oases" where bottom contact fishing gear such as dredges and trawls are prohibited. Tegner et al. (2001) tied declines in abalone production in Southern California not only to overfishing but to the cessation of growth of the alga Macrocystis pyrifera which provides food through drift of debris under the canopy, as well as providing habitat for abalones; such cessation coinciding with the warm, nutrient-poor waters associated with El Niño events.
Hydrodynamic factors act to generate spatio-temporal settlement patterns in shellfish populations. As population patterns and processes in shellfishes are scale-dependent, depending on the stage of the life history involved (Orensanz, 1986; Thrush, 1991; Defeo, 1996a, b), analysis of the physical-biological coupling at different scales is useful in the context of enhancement programmes. Peterson, Summerson and Luettich Jr. (1996) showed for the scallop Argopecten irradians that larvae larval settlement drops off sharply as a result of physical transport of their short-lived pelagic larvae. This has important implications in regulating population size in the system, as well as in developing adequate strategies for enhancement.
Oceanographic factors lead to site-specific variations in the local abundance of larvae and subsequent successful settlement, and thus determine the optimum times and sites for seed collection. For example, lack of suitable hydrographic conditions for supplying and retaining large numbers of larvae in the vicinity of collectors could lead to the failure of seed collection. Often, large settlements occur in semi-enclosed bodies of water or enclosed bays, which can also be good sites for early survival. These places are recurrent sites for successful settlement since they avoid mortalities due to flushing of water masses in the area. This is an important consideration when considering metapopulations, in which the capacity of larval dispersal over the fishing grounds is often dependent on the intensity and direction of wind-driven currents. In this context, self-recruiting, "source" areas (Carr and Reed, 1993) should be differentiated from "sink habitats" (see Chapter 3). As each ground has its distinct regime of primary production, nutrients, food availability, predation and disturbance, in theory, between-ground differences in these features could be quantified.
Enhancement operations will be increasingly affected by pollution on the coastal zone. Nearshore environments are more and more vulnerable to harmful algal blooms, sewage discharges, oil spills and so on. The quality of the site for seeding must be assessed from these points of view also, because their occurrence implies increasing variable costs and investment.
For shellfish resources, it is common to find some portions of the habitat more densely populated than others as a response to gradients in habitat quality. Such spatial variations might be assessed to determine distributional patterns common to adults and recruits, in order to select a site for restocking. The following mechanisms could explain patchy distribution patterns and should ideally be evaluated: (1) active larval choice, and ability to colonize areas of habitat with greater environmental stability where population growth is maximized; (2) occurrence of higher mortalities operating before and after settlement due to adverse environmental effects (e.g. low salinity); and (3) a major incidence of hydrodynamic forces, which determines passive transport of larvae to a given receiver site.
Population regulation may be habitat-dependent, as demonstrated for shellfish populations by the positive covariance between density, resources and environmental harshness (e.g. salinity, seston, food availability). Density-dependent habitat limitation within the seeding area could greatly reduce the potential benefit of any restocking programme. Shelter/space availability may control the size of many shellfish populations (see e.g. Caddy and Stamatopoulos, 1990; Beck, 1995). An attempt to investigate experimentally whether hatchery-reared animals displace natural stocks should aim at testing the hypothesis of habitat limitation.
The above facts clearly suggest that the optimum individual size for restocking and the carrying capacity of the system will depend on site quality and extent: suitable hydrographic conditions, absence or rarity of predators, food availability and available shelter could be some of the factors affecting habitat quality and thus carrying capacity. Carrying capacity will also differ at different stages of the life cycle (Orensanz, 1986; Fréchette, 1991) and restocking operations must take this into account. For example, if a natural stock is already present, the total biomass of wild plus enhanced stock should not lead to compensatory mortality and depressed growth rates as a result of stocking. Dijkema, Bol and Vrooland (1987) found that high population densities of the cockle Cerastoderma edule in Netherlands determined density-dependent growth rates and that at very high densities, some individuals are pushed out of the sediment and subsequently die. The re-seeding of cockles on an experimental scale demonstrated the major advantages to thinning very dense natural cockle beds in order to improve growth rates. All of these factors affect production and thus the economic viability of the operation (see Schiel, 1993 and Brand et al., 1991 for examples).
Maller (1990) and Blackburn, Lawton and Perry (1992) developed a simple and effective method to determine the slopes of the upper boundary (maximum densities) of the relationship between density and body size. Even though the original procedure was conceived for scaling body size to density, an issue also relevant for stock enhancement initiatives as an indicator of available energy in the ecosystem, the methodology equally applies to any combination of variables in which a Constraint Envelope Pattern (CEP: sensu Marquet, Navarrete and Castilla, 1995) has a real biological meaning. The procedure involves dividing the X-axis into intervals of equal length and recording the maximum value of the response variable on the Y-axis for each X interval (see also Marquet, Navarrete and Castilla, 1995 and Blackburn and Gaston, 2001 for additional theory). Because the value of the slope depends on interval size used in the X-axis, it is suggested to consider a range of interval sizes from, say, 0.1 units of the X variable up to a value that renders at least three values of Y (Marquet, Navarrete and Castilla, 1995). Then, the nature of the relationship defined by the points of the upper boundary is visualized through a simple scatterplot and then the appropriate (linear or non-linear) model is fitted. The upper limit corresponds to optimal combinations of the X and Y variables, whereas values within this "envelope", well below the upper ceiling, represent a wide range of suboptimal conditions. Typically, the CEP has a well-defined upper boundary with a negative slope indicating an inverse relationship between X and Y.
The above methodological approach has been suggested as a simple way of estimating carrying capacity through the use of a scatter diagram of adult and/or recruits density in each sampling unit or quadrat. At this small-scale of spatial resolution, a boundary of carrying capacity for both population components (adults and recruits combined) could be estimated (Orensanz, 1986). This must be done for different degrees of fishing pressure, and by site. Defeo (1996a) estimated carrying capacity with and without fishing activities. However, in the context of stock rebuilding initiatives, this approach is useful for evaluating optimal levels of abundance of the different population components at a given site, and thus is a help to planning the intensity of seeding operations through the estimation of optimal stocking densities (OSD). At a "quadrat scale", Defeo (1996a) showed that highest densities of recruits were never coincident with highest densities of older clams. Maximum densities of recruits per sample core were observed during 1984 and 1985, when they reached between 4 000 and 5 000 ind·m-2; during the same period, maximum adult densities were between 400 and 600 ind·m-2 but not in the same samples where the maximum recruit densities were recorded. These values of adult density, which correspond to the period of active fishery, were far below the maxima recorded after the fishery had been closed; in 1989 they reached 800-900 ind·m-2. It is notable that when adult densities were at least 300 ind·m-2, recruitment was almost absent in the same sample. The negative relation between recruit and adult densities for all years combined is shown in Figure 6.9. The line which defines the upper limits of the relationship represents maximum adult densities for varying levels of maximum recruitment; below the line, the lower values represent a wide range of suboptimal environmental conditions (Maller, 1990). The upper boundary mainly reflected higher densities of recruits during the years 1984-1985, and those of adults inhibiting recruitment during the experiment, i.e. in 1989 (Figure 6.9a). This "envelope" is linear in this case, but could take different forms (e.g. a monotonically decreasing exponential model). The form of this upper boundary should be taken into account when defining the appropriate combination of adult and recruitment densities in a stock-rebuilding experiment.
Figure 6.9 Scatter diagram of yellow clam recruit density plotted against adult density in each quadrat, for the months when recruitment peaked: see the difference between recruitment densities observed between 1983 and 1988 (·) under low adult densities and high extracting levels (1983-1987), and in 1989-1990 (o), as a result of the closure of the fishery. The dotted line defines the upper limit of the "envelope" between stock and recruitment, representing maximum recruitment densities for varying levels of maximum adult densities determined following Blackburn, Lawton and Perry (1992: see text for details).
Parsons and Dadswell (1992) found an inverse relationship between growth (shell height, meat weight, and whole weight) and stocking density in juvenile giant scallops, Placopecten magellanicus. This could affect OSD estimates, which was also dependent on the overall cultivation strategy type of grow-out technique, and the optimal timing of transfer from the pearl nets. Fréchette, Bergeron and Gagnon (1996) presented a method for estimating OSD via the analysis of the relationship between yield (biomass, B) and population density (N) at harvest, using a B-N diagram (BND). The analysis provided by the authors differs from the usual approach in aquaculture, in which yield is expressed as a function of initial population density, and B and N are analysed separately. Both methods allow estimation of OSD. The BND potentially allows (Fréchette, Bergeron and Gagnon, 1996): (1) assessment of the relative importance of competition-dependent and competition-independent mortality factors; (2) estimation of approximate OSD and maximum yield by extrapolation of results from short-term experiments; and (3) identification of the nature of the factor regulating competition-dependent mortality. They also compared the classical and BND methods using data from mussels grown in suspension cultures, and found that mortality patterns were the same for all stocking densities, and that competition-dependent mortality occurred only at high density. In an experiment designed to test the effect of spat origin (stock effect) on commercial yield, the classical approach suggested that there were no differences in yield and survival, despite differences in growth rate. The biomass-density approach (BN), however, showed that yield was constrained by self-thinning, not by intrinsic properties of the stocks. The BN approach, unlike the classical approach, yielded results consistent with state-of-the-art commercial practice and general knowledge about the stocks tested (Fréchette, Bergeron and Gagnon, 1996). Rheault and Rice (1996) showed that doubling the stocking density from 2.5 to 5.0 kg of oysters Crassostrea virginica per bag resulted in a 20 percent decrease in both the condition index and the growth rate (percent increase in weight). These observations may assist commercial growers determine optimal stocking density for their aquaculture grow-out systems. The variation in food concentration superimposed on the tidal current oscillation leads to massive changes in food flux and the degree of local resource competition.
Fréchette and Bacher (1998) noted that estimating physiological rates of the blue mussel Mytilus edulis in the field as a central part of carrying capacity studies. They also presented a strategy for estimating site-specific physiological rates based on the modelling of a reference growth experiment at a standard site. Growth of mussels was modelled as a function of population density to obtain estimates of biomass-density and production-density curves for the reference experiment. The authors stressed that these curves provide much of the information usually required for managing cultured populations. They concluded that combining the modelling of reference experiments in this way with particle transport models, may prove useful for assessing optimal stocking density in situations where intensive field work programs are not possible.
Intertidal mussels usually form complex multilayered matrices with density-dependent effects on survival and growth, and self-thinning scaling between biomass (B) and density (D) is expected. Guiñez and Castilla (1999) develop a tridimensional model of space-driven self-thinning that in addition to BN, explicitly includes the degree of packing of the mussels, measured as the number of layers (L). The model BNL could be considered as a generalized one in the sense that it encompasses previous bidimensional models (BN) of self-thinning (e.g. Fréchette and Lefaivre, 1990, 1995; Fréchette and Bacher, 1998) as special cases, and enables comparisons between mono- and multilayered populations. Guiñez and Castilla (1999) contrasted the predictions of the bi- and tridimensional models using data obtained from Perumytilus purpuratus mussel beds on the rocky shores of central Chile monitored during a 28-mo period. The B-N-L model suggested that density dependence is much more frequent than hitherto indicated by bidimensional models. The authors also applied their space-driven tridimensional model to other species where spatial overlapping configurations occur, such as the case of tunicate population of Pyura praeputialis in the Antofagasta Bay, northern Chile (Guiñez and Castilla, 2001).