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ICES Journal of Marine Science, 56: 405–409. 1999 Article No. jmsc.1999.0483, available online at http://www.idealibrary.com on Extinction considerations for diadromous fishes B. Jonsson, R. S. Waples, and K. D. Friedland Jonsson, B., Waples, R. S., and Friedland, K. 1999. Extinction considerations for diadromous fishes. – ICES Journal of Marine Science, 56: 405–409. Among diadromous fishes, the sturgeons are particularly threatened with 13 of 15 species considered endangered or vulnerable to extinction worldwide. These species have low intrinsic rate of growth and specialized habitat requirements making them sensitive to environmental changes and human exploitation. Habitat destruction, degradation, and alternations are probably prime causes of extinction in diadromous fishes. The spread of alien species, naturally or transported by humans to new environments, may be the second most important cause. Harvest is a contributing factor to extinction risk for many diadromous populations, although there is no known example where overfishing has led directly to the extinction of species. Population viability analysis is a method which may be used for evaluating extinction risk. However, uncertainties associated with parameter estimation and model choice limit its usefulness. Avoiding short-term extinctions is difficult enough, but restoration and recovery of natural populations present a whole new array of challenges. The understanding of the genetic population structure of the species is essential in this context. Captive rearing has some use as tool to assist stock restoration. However, artificial propagation will not lead to recovery unless the fundamental problems that caused the population declines are addressed. 1999 International Council for the Exploration of the Sea Key words: measuring extinction, population threats, causes of extinction, risk assessment, habitat restoration, population recovery. Received 27 March 1998; accepted 16 February 1999. B. Jonsson: Norwegian Institute for Nature Management, Dronningens gt 13, PO Box 736 Sentrum, N-0105 Oslo, Norway. R. Waples: Conservation Biology Division, Northwest Fisheries Science Center, 2725 Montlake Boulevard East, Seattle, WA 98112, USA. K. Friedland: Population Dynamics Branch, Northeast Fisheries Science Center, 166 Water Street, Woods Hole, MA 02543, USA. Correspondence to B. Jonsson: tel: +47 2335 5000; fax: +47 2335 5101; e-mail: [email protected] Introduction Extinction is a natural and inevitable biological process: the fossil record provides ample evidence that episodes of unusually high extinction rates occurred in this planet’s past and that species persist for finite periods of time. In comparison, documented recent extinctions involve just a small fraction of known animal and plant species (Smith et al., 1993). Nevertheless, there is growing concern regarding the fate of Earth’s biodiversity for two reasons: the number of recorded extinctions has climbed steeply over the past century and growing evidence for human impacts on ecosystems suggests that the extinction rate may increase rapidly in the future. The collection of papers in this issue examine various aspects of extinction risk for diadromous fishes, which by definition must migrate between fresh water and salt 1054–3139/99/040405+05 $30.00/0 water (or vice versa) to complete their life cycle (McDowall, 1999). These papers were part of a special theme session entitled ‘‘Diadromous Fish Extinction: Threats on Local and Global Scales’’ at the annual ICES Science Meeting held in Baltimore, Maryland (USA) in September 1997. Measuring extinction In exceptional cases, the last few members of a species can be monitored closely, and an extinction is recorded when the last individual dies. In general, however, measuring extinction is not a straightforward process. Both absolute and relative estimates of extinction rates are subject to a number of biases related to sampling effort, detectability, taxonomic uncertainty, and inadequate understanding of the organism’s biology (Smith 1999 International Council for the Exploration of the Sea 406 B. Jonsson et al. et al., 1993). A common practice (used, for example, by the World Conservation Monitoring Centre) is to consider a species extinct if it has not been observed for 50 years. This method may be reliable for once common and frequently encountered species, but it will be less so for species that have always been cryptic, seclusive, or otherwise difficult to observe. Grogan and Boreman (1999) showed how to calculate the probability that diadromous fish species have been extirpated, using a method that takes into account the temporal pattern of historic observations of the species. They also pointed out that in evaluating extinction probability for fish species, special care is needed if historical observations of early life stages are used, since fluctuations in abundance over time there are generally much larger than for the adult phase. Threats to diadromous fishes Due to their migratory habits, diadromous fishes may be at generally higher risk of extinction than many other groups and present special conservation challenges. Whereas approximately 5% of all fish species are considered endangered, threatened, vulnerable, rare, or of indeterminate status (Barbault and Sastrapradja, 1995), McDowall (1999) has identified about 18% of diadromous species that are of some conservation concern. One group, the sturgeons (Acipenser spp.), are particularly imperilled, both in Europe and Asia. This concerns both anadromous and freshwater resident populations. According to the IUCN Red List of Threatened Animals (Baillie and Groombridge, 1996), one species is critically endangered, facing an extremely high risk of extinction in the wild in the immediate future; eight other species are endangered and four species are vulnerable. Only two of the 15 anadromous sturgeons are low-risk species. The sturgeons are particularly vulnerable because they are long-lived with low intrinsic rates of population growth and their specialized spawning requirements make them sensitive to environmental changes as well as human exploitation. Other taxa of particular concern include the Prototroctidae (southern graylings) and the Alosinae (shads) (McDowall, 1999). Of the two southern grayling species, one is extinct and the other threatened. Reasons for their decline are probably combined effects of deforestation and trout introductions (McDowall, 1990). One shad in the Gulf of Mexico is endangered, and the two shads in the East Atlantic are vulnerable. For anadromous salmonids entire species appear not to be in any immediate danger of extinction, but many local spawning populations have been extirpated, and many more face significant risk of extinction (Kellogg, 1999). Causes of extinction Although human impacts have greatly increased species risks over the past century, extinctions rarely are due to one factor alone but instead involve the interplay of anthropogenic effects, natural factors, chance events, and biological attributes of the species. Threatened species are often characterized by one or more of the following: large body size, high trophic level, small population size, restricted geographic distribution, poor dispersal and colonizing abilities, colonial breeding habits, dependence on specialized habitats or ecosystems, migratory life history, dependence on unreliable resources, and inability to respond to environmental change or disturbances (Reid and Miller, 1989). Some of these risk factors apply to diadromous species, all of which are migratory and many of which are relatively large and near the top of the food web. Most diadromous species are widely distributed, but local populations typically occupy restricted areas and may be strongly isolated. Spawning populations range from relatively small (as is usually the case for Atlantic salmon, Salmo salar) to quite large units [as in many populations of pink salmon (Oncorhynchus gorbuscha) and sockeye salmon (O. nerka) in the Pacific]. Probably the most pervasive anthropogenic influence on extinction risk for diadromous fishes has been destruction, degradation, or alteration of habitat (Nehlsen et al., 1991; McDowall, 1999; Lichatowich et al., 1999; McKinnell and Karlström, 1999). Destruction and modification of habitat occurs through a wide variety of human activities, including forestry, agriculture, ranching, mining, hydropower development, and urbanization. Effects range from short term, localized modifications of habitat structure and function to alteration of ecosystem processes and structure (Gregory and Bisson, 1997). For example, the Atlantic salmon populations in southern Norway have been lost due to acidification of the freshwater habitats (Hesthagen and Hansen, 1991), and more than 90% of the Baltic salmon populations have been lost due to river regulation for hydropower purposes (Ackefors et al., 1991). Although we are not aware of any case in which overfishing has led directly to species extinction, harvest is a contributing factor to extinction risk for many populations of diadromous fishes. The predominant paradigm in harvest management for salmon has been based on taking a fixed percentage of the population. Cass and Riddell (1999) used a stochastic life cycle model to evaluate the interplay of harvest and natural variability in abundance for some British Columbia populations of chinook salmon (O. tshawytscha). Their results indicate that current fixed-rate harvest policies are not risk averse, nor are they likely to result in maximum long-term economic benefits. An alternative strategy which includes a minimum escapement floor Extinction considerations for diadromous fishes before any harvest is allowed provides much greater protection from extinction and greater long-term economic benefits than the current strategy. A modification of the threshold strategy to allow harvet rates to increase with increasing stock biomass could help alleviate some of the concerns for the economic and social effects of prohibiting all harvests in years of low abundance. Another cause of extinction is interaction with alien species, spread naturally or, more frequently, transported by humans to new environments. Some introductions are relatively benign, but others cause major disruptions to ecosystems and may play a significant role in elevating extinction risk. The spread of nonindigenous species is second only to habitat destruction in harming native communities and causing population extinction (Barbault and Sastrapradja, 1995). Humans have also enhanced the spread of diseases and parasites, which may be more responsible for population declines than is commonly recognized (McCallum and Dobson, 1995). For example, the monogenous parasite Gyrodactylus salaris was first introduced into a Norwegian river in the mid 1970s, and by 1991 it could be found in 35 rivers. This parasite causes a catastrophic mortality of Atlantic salmon parr; typically, parr are nearly absent from a river 2 years after Gyrodactylus is first observed (Heggberget et al., 1993). It has been estimated that Gyrodactylus may be responsible for a 50% reduction in the natural production of salmon in Norway (Johnsen and Jensen, 1991). A swimbladder parasite of the Japanese eel, Anguillicola crassus, was introduced to Europe together with living European eels, Anguilla anguilla, at the beginning of the 1980s. Within 10 years of its introduction it had achieved high rates of infection intensities in populations of the European eel over most of Europe (Hartmann, 1994). This parasite destroys the swimbladder of the host, with the likely consequence that eels with nonfunctional swimbladders will fail to reach their breeding grounds. The recent discovery of its appearance in American eels, Anguilla rostrata, in South Texas and South Carolina (Fries et al., 1996), is further reason for concern. Estimating extinction risk Key to avoiding extinction of a species is to detect which species are particularly sensitive. Evaluating extinction risk involves estimating the probability that a species (or population) will not persist for a given length of time. In the past 15 years, a great deal of progress has been made in developing quantitative models to do so (Boyce, 1992), and some of these have been applied to diadromous species (cf. Wainwright and Kope, 1999). This general approach referred to as population viability and analysis (PVA), can be quite powerful, but uncertainties 407 associated with parameter estimation and model choice have limited their usefulness to date. Practitioners must be cautious of model predictions having such low precision to render them useless. Furthermore, most PVA models are based on demographic and population dynamic processes of individual populations. Although some PVA models have made an attempt to incorporate genetic and/or metapopulation effects, a comprehensive, fully-integrated PVA model that incorporates all significant risk factors remains an illusive goal. Therefore, qualitative approaches, together with professional judgement, are likely to remain important for the foreseeable future. Evaluating extinction risk for anadromous Pacific salmonids poses some complex challenges. A practical list would include: (1) How can extinction risk for conservation units that are composed of many local populations best be evaluated?; (2) On what temporal and spatial scales are populations connected by migration and gene flow?; (3) How does metapopulation structure affect small-population demographic and genetic risks?; (4) How best can the genetic and ecological risks from hatchery production be factored into an overall risk analysis?; (5) When rivers are stocked with hatchery fish, how can sustainability of the natural population be evaluated?; and (6) What is the best way to evaluate the contribution of large-scale climate changes (e.g. decadal-scale cycles in ocean productivity) to extinction risk for salmon populations, which typically exhibit large fluctuations in abundance? Wainwright and Kope (1999) describe an approach that combines quantitative and qualitative elements to arrive at an overall assessment of extinction risk for complex conservation units of Pacific salmon. Restoration and recovery The challenge of avoiding extinction is a difficult one paralleled by the challenges of restoring extirpated or endangered populations. Foremost in any recovery strategy should be understanding the genetic population structure for the species of concern. Nielsen (1999) describes the historical processes that has led to the decline of Pacific salmon (Oncorhynchus spp.) at the southern end of their range in North America and commented on recovery strategies for these populations. Artificial propagation of salmon has been used for over 100 years as an alternative to conservation of natural populations; however, more recently there has been increased interest in its use in stock restoration. Artificial propagation can be used for reintroductions into inaccessible habitat, and it also has the potential to reduce transient extinction risk by rapidly increasing population size. A novel approach toward this latter end is described by Dempson et al. (1999), who captured 408 B. Jonsson et al. out-migrating juveniles from anadromous populations of Atlantic salmon (Salmo salar) in Newfoundland, transported the fish to a salt-water rearing facility, reared them to maturity, and subsequently returned them to the native stream to spawn as adults. The overall result was positive, indicating that captive rearing may have some use as tool to assist stock restoration. In the experiments, however, only 80% of the surviving adults released into a fjord returned to their native river for spawning, whereas 20% strayed to other nearby rivers. To home successfully, salmon is believed to pick up sequential cues during the outward migration for navigational help during the return (Hansen et al., 1993; Hansen and Jonsson, 1994). In the experiment of Dempson et al. (1999) the smolts were captured 0.75 km from the river outlet and the fish were trucked to the rearing facility in the estuary. This migratory disruption may be the cause of the high straying rate (Fleming et al., 1997). Smolt transfers of a few hundred metres have been found to influence the precision and timing of the return migration of Atlantic salmon (Jonsson et al., 1991, 1994), and smolt releases in rivers appear to give the lowest rates of straying to other rivers. Thus, there is also potential for improvements. Artificial propagation will not by itself lead to recovery unless the fundamental problems that caused the declines in the first place (or which are limiting recovery now) are addressed. Even the best-run hatchery programme entails risks for natural populations, and these risks must be balanced against potential benefits to determine whether supplementation can be effective in restoration schemes (Waples, 1999). There is still a great deal we need to learn about this issue. However, as discussed by Reisenbichler and Rubin (1999), the empirical data are sobering: each of the controlled studies that has examined this issue found that progeny of hatchery fish were characterized by substantial reductions in fitness in the wild compared to progeny of wild fish. This raises the possibility that at least some wellintentioned supplementation programmes may do more harm than good, and that healthy populations could be transformed into a state in which they must depend on continued hatchery supplementation for their persistence. Improved techniques of fish culture, release strategies, etc., can help to reduce deleterious effects of fish culture, but they cannot be avoided altogether (Waples, 1999). Finally, Lichatowich et al. (1999) point out that the basic causes of the decline of Pacific salmon (habitat degradation, overfishing, blockage of migratory routes by dams) have been understood for over a century. However, this knowledge has not prevented depletion of the resource, which suggests that institutional processes associated with fishery management have prevented effective conservation measures from being enacted. This simple realization indicates that, if meaningful recovery of diadromous fishery resources is to occur, biologists, managers, and society as a whole must fundamentally change the way species are perceived in the ecosystem and in social contexts. References Ackefors, H., Johansson, N., and Wahlberg, B. 1991. 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A., Furey, G., Bloom, M., Nicholls, T., and Hoskins, G. 1999. Evaluation of an alternate strategy to enhance salmon populations: cage rearing wild Atlantic salmon smolts from Conne River, Newfoundland. ICES Journal of Marine Science, 56: 422–432 (this issue). Fleming, I. A., Lamberg, A., and Jonsson, B. 1997. Effects of early experience on the reproductive performance of Atlantic salmon. Behaviour Ecology, 8: 470–480. Fries, L. T., Williams, D. J., and Johnson, S. K. 1996. Occurrence of Anguillicola crassus, an exotic parasitic swim bladder nematode of eels, in the southeastern United States. Transactions of the American Fisheries Society, 125: 794–797. Gregory, S. V., and Bisson, P. A. 1997. Degradation and loss of anadromous salmonid habitat in the Pacific Northwest. In Pacific salmon and their ecosystems, pp. 277–314. Ed. by D. J. Stouder, P. A. Bisson, and R. J. Naiman. Chapman and Hall, New York. Grogan, C. S., and Boreman, J. 1999. Estimating the probability that historical populations of fish species are extirpated. North American Journal of Fisheries Management (in press). Hansen, L. P., and Jonsson, B. 1994. Homing of Atlantic salmon: effects of juvenile learning on transplanted postspawners. Animal Behaviour, 47: 220–222. Hansen, L. P., Jonsson, N., and Jonsson, B. 1993. Oceanic migration in Atlantic salmon. Animal Behaviour, 45: 927– 941. Hartmann, F. 1994. Untersuchungen zur Biologie, Epidemiologie und Schadwirkung von Anguillicola crassus Kuwahara, Niimi und Ikagaki 1974. (Nematoda), einem blutsaugenden Patasiten in der Schwimmblase des europäische Aals (Anguilla anguilla L.). Reihe Biologie, Verlag Shaker, Aachen. Heggberget, T. G., Johnsen, B. O., Hindar, K., Jonsson, B., Hansen, L. P., Hvidsten, N. A., and Jensen, A. J. 1993. Interactions between wild and cultured Atlantic salmon: a review of the Norwegian experience. Fisheries Research, 18: 123–146. Extinction considerations for diadromous fishes Hesthagen, T., and Hansen, L. P. 1991. Estimates of the annual loss of Atlantic salmon, Salmo salar L., in Norway due to acidification. Aquaculture and Fisheries Management, 22: 85–91. Johnsen, B. O., and Jensen, A. J. 1991. The Gyrodactylus story in Norway. Aquaculture, 98: 289–302. Jonsson, B., Jonsson, N., and Hansen, L. P. 1991. Differences in life history and migratory behaviour between wild and hatchery reared Atlantic salmon in nature. Aquaculture, 98: 69–78. Jonsson, N., Hansen, L. P., and Jonsson, B. 1994. Juvenile experience influences timing of adult river ascent in Atlantic salmon. Animal Behaviour, 48: 740–742. Kellogg, K. 1999. Salmon on the edge. Trends in Ecology and Evolution, 14: 45–46. Lichatowich, J., Mobrand, L., and Lestelle, L. 1999. Depletion and extinction of Pacific salmon (Oncorhynchus spp.): A different perspective. ICES Journal of Marine Science, 56: 467–472 (this issue). McCallum, H., and Dobson, A. 1995. Detecting disease and parasite threat to endangered species and ecosystems. Trends in Ecology and Evolution, 10: 190–194. McDowall, R. M. 1990. New Zealand freshwater fishes: a natural history and guide. Reed, Auckland. McDowall, R. M. 1999. Short communication: Different kinds of diadromy: different kinds of conservation problems. ICES Journal of Marine Science, 56: 410–413 (this issue). 409 McKinnell, S. M., and Karlström, O } 1999. Spatial and temporal covariation in the recruitment and abundance of Atlantic salmon populations in the Baltic Sea. ICES Journal of Marine Science, 56: 433–443 (this issue). Nehlsen, W., Williams, J. E., and Lichatowich, J. A. 1991. Pacific salmon at the crossroads: Stocks at risk from California, Oregon, Idaho, and Washington. Fisheries, 16: 4–21. Nielsen, J. L. 1999. The evaluation history of steelhead (Oncorhynchus mykiss) along the US Pacific Coast: Developing a conservation strategy using genetic diversity. 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