Genetic Problems of Hatchery-Reared Progeny Released into the Wild ...

BULLETIN OF MARINE SCIENCE, 62(2): 623–640, 1998

GENETIC PROBLEMS OF HATCHERY-REARED PROGENY RELEASED INTO THE WILD, AND HOW TO DEAL WITH THEM Fred Utter ABSTRACT Long-overdue concerns about the genetic effects of releases of hatchery-reared progeny on wild populations have been increasingly addressed during the past 20 yrs. Because of an extensive history of translocation and culture, much attention in both of these categories has focused on salmonid species. Native salmonids have been indirectly affected by translocations through such activities as induced overharvest in stock mixtures, disease introductions, and displacements resulting in fragmented populations with reduced numbers and, commonly, localized extinctions. Direct genetic effects through introgressive hybridizations of translocated salmonids have resulted in replacement of many native populations by hybrid swarms. Genetic effects from salmonid culture have led to losses of variability, both within the cultured populations and between the total cultured and wild populations, and to adaptive divergences distinguishing cultured and wild populations. Examples are provided that show that the overall salmonid experience is directly relevant to marine culture as well, although special considerations deal with the lower among-population diversity and greater fecundity often found in marine species. As marine enhancement programs inevitably expand, identifying and preserving natural populations are concerns common to harvest, management, and conservation interests. Coordinated rather than polarized approaches to these concerns best serve all parties as well as the irreplaceable resources.

Concerns about releases of cultured aquatic organisms into the environment have accelerated over the past 25 yrs, following centuries of largely unregulated activity. This shift from a benign to a more questioning outlook arose through the convergence of multiple factors. In salmonid fishes, for example, increased awareness of the intrinsic value of indigenous species and populations that were threatened or displaced by exotic ones (Utter, 1981; Ryman, 1983; Waples, 1991a) coincided in many areas with continuing declines in overall productivity in spite of increased releases of cultured fish (Hilborn, 1992; Hilborn and Winton, 1993). During this period, the development of single-gene (i.e., Mendelian) markers through emerging tools of molecular biology facilitated understanding and dealing with some of the genetic issues underlying these concerns (Utter, 1991; Carvalho and Pitcher, 1994). This paper probes genetic consequences resulting from releases of cultured fish. The primary focus is on the level of genetic variability where gene exchange may occur among individuals, as it commonly does within species. First, effects of exogenous introductions are considered on the basis of examples from a long history of culture and release of salmonid fishes. Variable genetic results that may arise from differing cultural practices within a common natural population are then examined. The following section focuses on cultured releases in general, and particularly on marine species, and is based on the lessons that have accumulated through releases of cultured salmonids. Courses of action are outlined that are generally necessary to increase the efficiency of enhancement efforts while perpetuating the genetic variation existing within and among natural populations.

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RELEASES OF CULTURED SALMONIDS Salmonid fishes are a useful group for examining genetic problems related to releases of cultured fishes into wild populations. Their commercial and sporting value and relative ease of culture have led to a long history of hatchery operations (see Davis, 1956). These attributes, coupled with a diversity of life-history adaptations, have resulted in distribution of salmonid populations far beyond their native ranges in temperate freshwater and marine environments throughout the earth (see MacCrimmon, 1971; Quinn et al., 1996). An extended literature has arisen on the genetic consequences of this extended culture and redistribution, much of which is generally pertinent to questions concerning genetic effects of fish culture on both cultured and native populations. Some of this literature is explored in this section as a basis for subsequent considerations of pertinent generalities. EXOGENOUS POPULATIONS Evaluating the extent to which and the manner by which indigenous populations are genetically affected by releases of cultured fish is commonly complicated by the inability to separate the effects of exotic origins from those of the cultured environment (Campton, 1995). In order to clarify this distinction, I consider only problems related to exotic origins at this point. An overview of some potential indirect and direct genetic effects arising from introductions of nonnative populations (Fig. 1) covers a broad range of possibilities. Indirect Genetic Effects.—Indirect genetic effects are defined here as those inducing genetic changes within or among populations without actual infiltration of exogenous genomes into native groups (i.e., without introgressive hybridization). Induced decreases in numbers within a population threaten its capability to adapt through reduced genetic variability and possibly inbreeding depression and also increase the likelihood of extinction, the ultimate genetic loss. Indirect genetic effects among populations arise from isolation of formerly connected groups from normal patterns of migration and gene exchange. The potential for further coevolution of such disrupted groups has ceased as a consequence of diminished or eliminated reproductive, rearing, or migratory habitats. Consequently, their combined reproductive and adaptive potential has become modified and threatened. A few examples in the context of those outlined in Figure 1 are mentioned here (cf. Hindar et al., 1991; Krueger and May, 1991, for more extended descriptions). The presence of large numbers of cultured fish stimulates excessive harvest pressure on wild populations. This topic has been extensively discussed and documented for anadromous salmon and trout populations of the genus Oncorhynchus of western North America (e.g., Wright, 1981; Fraidenburg and Lincoln, 1985; Lichatowich and McIntyre, 1987). Recognition of this problem has stimulated considerable interagency activity resulting in current use of largely genetically based mixed-stock analysis (MSA) in postseason and real-time harvest management applications aimed at protecting weaker wild populations of Pacific salmon (Shaklee and Phelps, 1990; Utter and Ryman, 1993; Lincoln, 1994). Although such protective efforts are encouraging, mixed-stock fisheries persist as a major contributor to declining natural populations, being associated with 100 of 214 naturally spawning groups of Pacific salmon recently identified as at risk of extinction (Nehlsen et al. 1991). The problem of mixed stock fisheries is not limited to human harvests. For example, attraction of predators by large-scale releases of cultured fish has been implicated in the

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Figure 1. Some indirect and direct genetic effects of releases of exogenous hatchery-reared fish on native populations.

decline of comigrating native populations of coho salmon, O. kisutch (Anonymous, 1987), in Oregon. The possibility of similar predation by squawfish, Ptychocheilus oregonensis, attracted by releases of juvenile salmonids in the Columbia River has been described (Steward and Bjornn, 1990). Reductions of native populations by diseases introduced by exogenous fishes are well documented. A particularly severe problem of current interest concerns the fluke Gyrodactylus salaris, presently infesting susceptible Norwegian populations of Atlantic salmon Salmo salar, which arrived with introduced resistant Baltic fish (Bakke et al., 1990). The problem has been compounded by furunculosis introduced to Norwegian cultured fish from cultured Scottish salmon infected with the bacterium Aeromonas salmonicida. In addition to spreading furunculosis, escaped cultured fish have competed and possibly interbred with the decimated wild populations in spawning areas (e.g., Hindar 1992). The disappearance of Norwegian wild salmon early in the next century has been predicted if these conditions persist (K. Hindar, pers. comm.). Nonintrogressive hybridization (i.e., that going no further than the first generation) between released hatchery and native fish promotes displacement of native fish through wastage of gametes and habitat. An increased incidence of first-generation hybridization between Atlantic salmon and brown trout (see Verspoor and Hammar, 1991) has occurred since escaped netpen-reared salmon have returned to spawn in Scottish rivers (Youngson

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et al., 1993), indicating a potential threat of such escapes to native populations of both species. This process also appears to be contributing to the displacement of endangered bull trout (Salvelinus confluentus) by introduced brook trout (S. fontinalis) in the western United States (Leary et al., 1993, 1995). However mediated (e.g., by mixed-stock fisheries, disease, competition, nonintrogressive hybridization), displacements by nonnative populations or species have contributed to fragmentation of native populations in diverse marine and freshwater habitats (see Ryman et al., 1995b). Such disruptions inevitably reduce the amount of contiguous habitat available for population maintenance. The resulting restrictions in gene flow and rearing and reproductive space reduce the numbers of breeders, promote short- and long-term decreases in adaptive capability through genetic degradation, and increase the likelihood of local extinctions. The threat to survival of species that depend on maintenance of substantial contiguous habitat, such as bull trout (Rieman and McIntyre, 1995), grows as fragmentation increases. Modified and restricted habitats inevitably impose different selective pressures on populations that survive disruptions inflicted by exogenous introductions, although measurement of such changes may be difficult (Waples, 1991b). For example, the appearance of riverine sockeye salmon, O. nerka, (an atypical life-history adaptation circumventing the normal need for a lake for juvenile rearing and adult maturation) in tributaries of the upper Columbia River is correlated with major habitat disruptions and relocations within this region over the past 60 yrs (Chapman et al., 1995a). Direct Genetic Effects.—Direct genetic effects are restricted here to any outcome resulting from introgressive hybridization of hatchery releases with an indigenous population. Introgressions break down genetic distinctions between populations and species that are important in maintaining the long-term viability of species and subspecific taxa. In addition, evidence is accumulating for introgressive erosion of population fitnesses through matings of genetically divergent individuals, i.e., outbreeding depression (Waples, 1995; Allendorf and Waples, 1996). The abundance of Mendelian markers revealed through the advent of molecular genetics (e.g., Utter, 1991, 1994) has facilitated the detection of introgression (Campton, 1987, 1990; Verspoor and Hammar, 1991). Fixed differences (i.e., absence of shared alleles) at one or more loci commonly distinguish congeneric species (Shaklee, 1983; Campton, 1987, 1990). Unambiguous evidence for introgressive hybridizations in native populations resulting from numerous releases of hatchery salmonids has been obtained through the existence of such fixed differences (e.g., Leary et al., 1995). Detecting introgression is much more problematical between groups distinguished by differing frequencies of the same alleles, a prevalent situation among conspecific populations (Campton, 1987, 1990). As subsequently noted, conspecific populations of marine fishes are typically less differentiated than those of freshwater fishes (Utter and Ryman, 1993). Introgressions at this level, although undoubtedly common, are less clearly documented than those identified through fixed differences. The difference in discriminatory power between fixed and quantitative allelic differences is illustrated through two examples based on released hatchery salmonids in the northwestern USA (Fig. 2). Different levels of introgression of released rainbow trout with native westslope cutthroat trout are clearly identified in four collections based on fixed differences at two loci (Fig. 2A). Conclusions based on quantitative allelic differences at two loci distinguishing coastal and inland populations of rainbow trout are far


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Figure 2. Contrast between fixed allelic differences (no common alleles) and quantitative differences (varying frequencies of the same alleles) for measuring introgressive hybridization. A. (From Leary et al., 1984.) Identification of hybrid swarms between native westslope cutthroat trout, Oncorhynchus clarki clarki, and introduced rainbow trout, O. mykiss, in four tributaries of the Clark Fork River based on fixed differences between taxa (*100 alleles characterizing rainbow trout are absent in westslope cutthroat trout). B. Mean values and ranges of allele frequencies of collections of resident and anadromous (steelhead) rainbow trout having different frequencies of the same alleles at two polymorphic loci. Locations and life histories include four coastal steelhead (A1), seven inland steelhead (A3), and, respectively, five Yakima River resident (A2), two Yakima River steelhead (B), and eight steelhead from western drainages of Mount Adams, Washington (C). Sources of data include A, Campton and Johnston (1985); B, Schreck et al. (1986); C, Phelps et al. (1994).

less conclusive (Fig. 2B). Nonoverlapping allele frequencies of native coastal and inland groups, coupled with intermediate allele frequencies in the inland Yakima River drainage that had been heavily stocked with coastal-type fish, initially suggested introgressive hybridization among self-perpetuating resident populations (Campton and Johnston, 1985). Presumably, migratory (i.e., steelhead) native inland populations were trapped by downstream migratory blockages and ultimately hybridized with the introduced fish. However, intermediate allele frequencies were also detected in subsequent samplings of steelhead populations in the Yakima River and adjacent drainages. Because these drainages lie near the boundary separating native coastal and inland populations, an alternative

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hypothesis, that resident and migratory populations in this transitional area arose through more ancient introgression during postglacial repopulation of this area, has been suggested (Chapman et al., 1994a). Clarification of these hypotheses is important as biologists and managers increasingly focus on identifying and preserving native populations (Waples, 1991b) and diminishing or eliminating hybridized ones (Leary et al., 1995). Such clarification can be facilitated through information from maternally inherited mitochondrial DNA, as has been done with native and introduced rainbow trout populations of the Snake River (Williams et al., 1996). In spite of the limitations of quantitative allelic variations, consistent patterns have proved valuable for distinguishing major ancestral intraspecific subgroups (see Ståhl, 1987; Utter et al., 1989). Once established, these distinctions permit monitoring for temporal changes that may indicate the occurrence of introgressions. For example, records for six polymorphic loci were maintained for from 10 to 14 yrs (depending on the locus) for two genetically distinct fall-run chinook salmon (O. tschawytsha) populations of the Columbia River belonging to a common major subgroup. A small population returning to the Snake River is presently protected under the U.S. Endangered Species Act, and a much larger and unprotected stock returns to the mid-Columbia River upstream from the Snake River (Matthews and Waples, 1991; Bugert et al., 1995). A convergence of Snake River allele frequencies toward those of the mid-Columbia River stock through 1989 comported with straying rates from the mid-Columbia River of 43% (based on codedwire tag returns) in 1989. The Snake River hatchery stock was subsequently returned to allele frequencies similar to those preceding the years of excessive straying through use of only tagged fish with Snake River codes in its maintenance. Conversely, major ancestral regional patterns persisted for more distantly related conspecific groups of Columbia River chinook salmon in spite of extensive recorded introductions at this level over many decades (Utter et al., 1995). Thus, large-scale introgressions appear to have been resisted at this greater level of divergence. However, the abovenoted cautions pertain to this conclusion. Limited introgressions cannot be excluded, particularly those based on alleles common to immigrant and native groups. In addition, introgressions resulting from introductions within major ancestral groups have almost certainly occurred but remain undetected because the necessary monitoring, such as that noted in the above example, has not occurred. Nevertheless, such resistance to widespread within-species introgression in the presence of extensive translocations, though unpredictable, is not uncommon. For example, unexpected persistence of apparently local populations of anadromous (steelhead) rainbow trout was found in many areas within the state of Washington having long histories of stocking with exogenous fish (Phelps et al., 1994). Clearly, presumed absence of native populations resulting from an extensive stocking history requires confirmation from reliable genetic information. INDIGENOUS CULTURED AND WILD POPULATIONS In addition to those related to released exogenous hatchery and native fish, genetic problems arise entirely within cultured local populations. These generally take the form either of losses of genetic variation resulting from restricted effective numbers of breeders or of adaptive changes that may result from differences between culture and natural conditions. Such genetic changes are examined through different types of hatchery operations that might arise from a hypothetical natural population (Fig. 3). Unless otherwise


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Figure 3. Some different patterns of migration between a natural population and derivative hatchery populations.

stated, sampling of the natural population (NA) occurs proportionately to the numbers of fish maturing over the entire spawning season. Each type of operation presents a distinct set of genetic concerns and opportunities related to the hatchery operation itself and its effect on population NA. Population S is intended to increase the size of NA and is derived entirely from a single generation of cultured progeny of a subsample of NA. If successful, this supplementation increases the survival of cultured over naturally reproduced progeny. Ryman and Laikre (1991) noted that this elevated survival (which they termed supportive breeding) reduces the effective size (Ne) of the overall population (NA + S) because of the increased variance of family size (Wright, 1931). If population NA is already weakened by low numbers, supportive breeding would magnify the threat of inbreeding. Subsequently, Waples and Do (1994) demonstrated that inbreeding was least (although potentially significant) when the enhanced population remained large in subsequent generations, i.e., the enhancement was successful. They cautioned that causes of decline must be understood and corrected before programs of supportive breeding are attempted, so as to maximize the probability of successful enhancement, and that maintenance of existing numbers would serve only to accelerate levels of inbreeding. They noted that such precautions have been incorporated into an ongoing enhancement project for spring-run chinook salmon of the Dungeness River on the Strait of Juan de Fuca, where the success presently is undetermined.

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Hatcheries A, B, and C produce and release fish each year to maintain or increase commercial and sports harvests. They differ in the manner in which their brood fish are derived or their surplus production is distributed. Like population S, both Hatchery A and Hatchery B contribute to supportive breeding of the overall population through contributions to the wild population. Each generation, some wild brood fish are intentionally included for Hatchery A to reduce the possibility of strong adaptations to hatchery conditions. The surplus production released to spawn naturally thus retains much of its wild adaptive capability. As a compromise between entirely cultured and entirely natural brood stock, Hatchery A retards adaptations either to hatchery conditions (i.e., domestication) or to variable natural settings accommodating different subgroups of population NA. The production of summer/fall-run chinook salmon at the Wells Dam Hatchery (on the Columbia River just downstream from the confluence of the Methow and Okanogan rivers) is analogous to that at Hatchery A. Spawners each year include fish returning to the hatchery that were reared and released there as juveniles, strays from other similarly operated hatcheries in that region, and intercepted naturally produced fish migrating past the hatchery. Progeny from Wells Hatchery are released either in nearby tributary streams or at the hatchery (Chapman et al., 1994b). This blending of wild and hatchery fish each generation coincides with genetic homogeneity of wild and hatchery summer/fall-run fish upstream from the confluence of the Columbia and Snake rivers (Utter et al., 1995). Thus both wild and hatchery summer/fall-run chinook salmon populations appear to be in a perpetual flux, which prevents strong adaptations to either hatchery or natural conditions. Although production at Hatchery B is initiated from natural fish, it is perpetuated from fish reared and released from the hatchery. With each succeeding generation, population B diverges increasingly from population NA through adaptations to conditions of artificial breeding and rearing. Like that of Hatchery A, its surplus is released for natural spawning. Hatchery production of fall-run chinook salmon of the lower Columbia River (called “tules”) has generally followed the model of Hatchery B. Considerable genetic homogenization has been observed, apparently reflecting extensive strayings and egg shipments among hatcheries (Simon, 1972; Utter et al., 1989; Marshall et al., 1995). Strays from nearby hatchery populations typically dominate natural spawning fish (Marshall et al., 1995). Not surprisingly, fish like those from Hatchery B tend to be outperformed by native populations in their natural environments (e.g., Reisenbichler and McIntyre, 1977; Chilcote et al., 1986; Campton et al., 1991). In addition to adaptive disadvantages based on habitat and on physiological and behavioral differences (e.g., Fleming and Gross, 1992; Fleming et al., 1994), duration of spawning is often reduced and peak timing altered by conformity to logistical constraints of hatchery operations (see Allendorf and Waples, 1995). Furthermore, restricting brood fish to those of hatchery origin may also lead, in the absence of suitable monitoring, to excessive losses of effective numbers of breeders relative to comparable native populations, as has apparently happened with some coastal chinook salmon hatchery populations in Oregon (Waples and Teel, 1990). Hatchery C differs from Hatchery B only in that its surplus production is confined to the area of the hatchery. Because the natural population is not affected by competition and interbreeding with population-C fish, the two groups are free to evolve independently of one another. Concerns relating to inbreeding remain important for population C, and as with population B, adequate effective numbers of spawners are needed each


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generation. However, possible adaptation to hatchery spawning and rearing is viewed positively. The hatchery is maintained as an ecosystem separate and distinct from that of the wild population, and the two groups are permitted to evolve optimally in their respective habitats. Management of spring-run chinook salmon returning to Leavenworth National Fish Hatchery in the drainage of the Wenatchee River (a tributary of the Columbia River) follows the model of Hatchery C. Releases of hatchery fish have been made either at or downstream from the hatchery site on Icicle Creek since large-scale chinook production began in 1971 (Chapman et al., 1995b). Populations of natural spawning areas 20 mi or further upstream in Nason Creek and the Chiwawa and White rivers remain genetically distinguishable from hatchery fish (Utter et al., 1995). A separate Chiwawa River supplementation program (Chapman et al., 1995b) follows the model of population S. RELEVANCE OF SALMONID EXPERIENCE TO MARINE STOCK ENHANCEMENT The examples based on salmonid culture serve to outline general areas of genetic concern for the protection of indigenous populations directly or indirectly affected by cultural activities. The repeatedly expressed basic principles for dealing appropriately with these concerns entail identifying and preserving genetic variation (see FAO/UNEP, 1981; Ryman, 1981; Hindar et al., 1991; Allendorf and Waples, 1996). These general topics are separately considered below, with particular reference to marine populations. IDENTIFYING GENETIC RESOURCES Understanding population structures underlies all other genetic concerns in considering interactions of cultured and wild populations. This critical need has been outlined in a position statement of the Genetics Section of the American Fisheries Society (Utter, 1995) presented at the symposium “Uses and effects of cultured fishes on aquatic ecosystems” (Schramm and Piper, 1995), where requirements included “...field studies to verify the existence and ecological parameters of natural populations, ...laboratory studies to identify the geographic patterns of indigenous ancestral groupings and the presence of exogenous or introgressed populations resulting from introductions, ...determinations of genetic and ecological relationships among hatchery, native and other naturally reproducing populations.” Mismanagement is inevitable without such information because relationships between cultured and wild populations remain conjectural, and monitoring cannot detect genetic changes resulting from introgression, displacement, or excessive inbreeding. Genetic divergence among conspecific marine populations, though quite variable, is typically lower than that observed in freshwater and anadromous fishes, presumably because of lower restrictions to gene flow (Waples, 1987; Utter and Ryman, 1993; Ward et al., 1994). Where genetic differentiation remains indistinct among samples collected over broad geographic ranges, or contrasting environments or life histories (e.g., Atlantic herring, Clupea harengus, Ryman et al., 1984), a source for initial selection of appropriate brood fish may be problematical because of possible undetected divergences. Depending on the cultural goals, such remaining differences (e.g., spawning and migration times, habitat preferences) are often presumed to have some underlying genetic basis and need to be considered (see discussions by Waples, 1991b, of evolutionarily significant units).

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The lower levels of genetic divergence typical of many marine fishes make cultured populations attractive candidates for intentional genetic marking (see Seeb et al., 1986; Utter and Seeb, 1990; Utter and Ryman, 1993). Because the marked population would be particularly conspicuous against the genetically homogenous background of the wild population, the distribution, fishery contribution, and reproduction of releases could be readily measured and monitored. Releases of genetically marked marine fish include cultured cod in Masfjorden, Norway (Jorstad et al., 1994), and red drum in Texas (Ward et al., 1995). PRESERVING GENETIC VARIATION At first glance, the large sizes and opportunities for gene flow of many marine species appear to make them relatively immune to major genetic losses. As discussed by Ryman et al. (1995b), however, concerns about genetic degradations of marine populations are indeed valid for reasons including (1) undetected subdivision, as previously discussed; (2) greater opportunities for selection in large populations; (3) limited numbers and distributions of some marine species; and (4) losses of low-frequency alleles in population crashes. Specifically of concern here is whether or not apprehensions drawn from experiences with salmonid culture realistically extend to marine stock enhancement. The following discussion explores this question as it relates to releases of cultured marine fishes from the perspectives of exogenous populations (Fig. 1) and cultural practices (Fig. 3). Limited empirical data support the generality of the salmonid experience. Exogenous Populations.—Species inhabiting marine estuaries and intertidal areas are particularly susceptible to genetic losses through reduced population sizes and habitat fragmentation. The endangered delta smelt, Hypomesus transpacificus, in the estuary of the Sacramento and San Joaquin rivers in California may be in sympatry with introduced Japanese smelt, H. nipponensis, established in several upstream reservoirs since 1959 (Stanley et al., 1995); although not yet detected, competition and hybridization between introduced and native species remain distinct possibilities. Competition from introduced mosquitofish, Gambusia affinis, has clearly contributed to the decline of endangered Iberian toothcarp, Aphanius iberus, in marine estuaries of the northwestern Mediterranean Sea (Garcia-Marin et al., 1990). Potentially catastrophic effects on natural aquatic populations have been abetted by exogenously introduced bacterial, viral, and parasitic diseases in addition to the devastations of native populations of Atlantic salmon in Norway (see reviews by Stewart, 1991; Sindermann, 1993). Particularly noteworthy in the present context are eel parasites, Anguillicola, apparently introduced to European eels, Anguilla anguilla, through introductions of exogenous eels from Asia and New Zealand. Widespread infestations presently threaten both wild populations and wild-derived cultured fish (Peters and Hartmann, 1986; Moller et al., 1991), and preliminary evidence suggests initial presence of A. crassus in North American eels, Anguilla rostrata (Fries et al., 1996). Similarly, reintroduction of the European oyster Ostrea edulis from colonized North American populations brought a North American parasite, Bonamia, which has subsequently devastated European populations (Chew, 1990). The absence of examples of introgressive hybridizations resulting from translocated marine species or populations may reflect limited opportunities rather than any intrinsic biological bases excluding such invasive gene transfers. Evidence in marine fishes for


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natural hybridizations and possible introgression includes both pleuronectiform and anguillid species (Aron, 1958; Verspoor and Hammar, 1991). Indeed, unchecked expansions of marine enhancement efforts will inexorably be accompanied by increased instances of the circumstances resulting in the above-mentioned possibility for hybridization between native and introduced species of Hypomesus. Amid increased contact between related but previously isolated taxa, altered reproductive schedules, and disturbed habitats, the salmonid experience of more frequent hybridizations resulting from relocations may be destined for repetition in marine species; the typically lower differentiation among conspecific marine populations limits our ability to detect such introgression. A common theme in the preceding paragraphs is the need for restrictions, perhaps even a moratorium, on relocations (see Krueger and May, 1991) in marine enhancement efforts. In summarizing a symposium on ecological and genetic effects of fish introductions, Allendorf (1991) noted: “Purposeful introductions rarely have achieved their objectives. Moreover, both intentional and unintentional introductions usually have been harmful to native fishes and other taxa through predation, competition, hybridization and the introduction of diseases.” More specifically, Ryman et al. (1995b) observed: “The globally increasing aquaculture industry will result in increasing incidental and intentional releases of huge numbers of fish, although those fish will represent a fairly limited number of species and populations. Thus, the threat of hybridization from cultured fish to the genetic integrity of local populations must be considered a major concern.” The short-term expedience of intentionally or accidentally releasing nonnative fish must be carefully balanced by the sobering knowledge of irreversible and unpredictable effects of past introductions. Cultural Practices.—The requirement for adequate numbers of individuals in founding and maintaining cultured populations has recently received considerable attention (e.g., Allendorf and Ryman, 1987; Kapuscinski and Jacobson, 1987). As reviewed by Ryman et al. (1995b), these numbers are commonly derived from the relationship H = 1/ (2Ne), where H is the decrease of heterozygosity (H) per generation and Ne is the genetically effective population size. Ne, commonly considerably less than the actual population size, is particularly reduced by deviations from equal sex ratios and increased variances of family sizes. Such reductions as well as population bottlenecks are not compensated for by increased Ne in subsequent generations. Effective numbers between 50 and 500 individuals per generation have been recommended for avoidance of serious losses of genetic variability in cultured populations over time. As with salmonids, effective numbers of natural and cultured populations of marine species are affected by interactions such as those outlined in Figure 3. Data from two species of sea bream identify the potential for genetic losses within cultured populations of marine species as well as between the total cultured and wild populations. In a situation analogous to that of Hatchery A in Figure 3 for cultured black sea bream, Acanthopagrus schlegeli, in Japan (Taniguchi et al., 1983), estimated effective numbers of parents in two brood stocks were, respectively, 16 and 26, far below the recommended minimum effective numbers; these estimates were accompanied by a 23% reduction in heterozygosity relative to the natural population. Similarly, estimates of effective numbers of parents in cultured populations of red sea bream, Pagrus major, ranged between 8 and 16, and the

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reduction of heterozygosity relative to the natural population was 18% (Sugama et al., 1988). Analogous to Hatchery S in Figure 3, some allele frequency differences subsequently detected among collections of local wild fish were attributed to releases of these cultured fish (Taniguchi and Sugama, 1990). The higher fecundity of many marine species gives them a potentially higher susceptibility than salmonids to genetic losses from cultural practices. The millions of eggs produced by a single female cod, as opposed to at most a few thousand from a typical salmonid, provide much greater opportunity for fortuitous survival of progeny from very few matings. Given the finite rearing space of a culture operation, particular care must be given to maintenance of adequate (and more or less equal) numbers of male and female spawners in initiation and perpetuation of brood stocks. These concerns must be particularly considered in previously mentioned programs involving genetic marking of cultured marine fishes, where considerable effort may be required to identify sufficient parents to initiate a marked population (Gharrett and Seeb, 1990). Finally, concerns about adaptation of cultured marine species cannot be ignored. Experience with salmonid culture (e.g., Reisenbichler and McIntyre, 1977) shows that, with or without intentional selection of parents, each successive generation of captive brood stock will favor those parents and progeny that are most amenable to conditions of culture. Under long-term development of captive brood stocks (situations B and C in Fig. 3), the localization of intentional or incidental releases would be needed to minimize competition and interbreedings between cultured and wild populations. Wherever such long-term captive brood stocks are developed, empirical tests for differential adaptations need to be undertaken, and where adaptations are detected, suitable isolation is necessary. CONCLUDING REFLECTIONS A brief review of the value and limitations of molecular genetic information is pertinent at this point, given the widespread and increasing applications of such data to questions about interactions of natural and cultured populations. These data have proven to be enormously valuable in clarifying previously unclear ancestral relationships at various taxonomic levels primarily through differences identified by Mendelian markers and, particularly within species, largely influenced by the evolutionary processes of drift and migration (Avise, 1994). Through such small samplings relative to the entire genome, objective groupings of related organisms have been identified, managed, and protected (Waples, 1991b; Nielsen, 1995), and losses of genetic variation conclusively demonstrated within and among such groupings (Utter, 1991). The novelty and above-mentioned value of molecular genetic identifications tend to mask their limitations. For example, induced interactions between two lineages distinguished by a number of molecular Mendelian markers are inadvisable because of the opportunities for adaptive differences (disease resistance, habitat preference, timing and extent of migration, etc.) to arise during the interval of divergence (Hindar et al., 1991). However, these distinctions yield almost no direct information about such adaptive differences (see Conover, this volume); the vast majority of the genome involved in adaptive processes remains largely unexplored (Hartl, 1994; Allendorf, 1995; Hard, 1995). Given such ignorance, similarities based on molecular markers are largely uninformative with regard to adaptation (Utter et al., 1992, 1993b). Furthermore, even detailed molecular-genetic data provide only limited insights about long-term evolutionary pro-


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Figure 4. Schematic representation of information necessary for a priori predictions of effects on natural gene pools when individuals bred in captivity are released into the wild. From Ryman et al. (1995a).

cesses in hatchery and wild populations (Waples and Teel, 1990; Hedgecock et al., 1992). Consequently, a priori predictability regarding the consequences of specific hatcherywild interactions is precluded (Fig. 4, Ryman et al., 1995a). This limitation requires that actual or potential culture-wild interactions be preceded by preliminary evaluations and risk assessments from empirical observations as outlined in the preceding text. A systematic and cautious approach to marine stock enhancement is therefore essential as empirical data accumulate. An additional reality must also be acknowledged. As the demand for harvest of both natural and cultured fishes expands in proportion with our ever-growing human population, the inherent conflict between harvest and conservation interests also inevitably increases (e.g., see discussion by Ryman et al., 1995b). We must recognize this incompatibility to minimize polarizations and thus to promote the dialog and mutual understandings that must underlie any effective long-term conservation efforts. ACKNOWLEDGMENTS I am grateful for the support of William and Lenore Mote and the Department of Biological Science, Florida State University, in sponsoring this symposium and its proceedings and particularly for the energy, vision, and leadership provided by F. Coleman. This manuscript profited from valuable reviews by C. Coronado, C. Peven, J. Winton, and two anonymous reviewers.

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