Westslope cutthroat trout COSEWIC assessment and status report: chapter 8

Limiting Factors and Threats

A number of factors appear to be limiting the abundance of WCT in Canada. While some of these occur naturally, it is clear that the most imminent and serious threats to cutthroat are of anthropogenic origin; primarily habitat loss, overharvesting, and the introduction of non-native species and/or genotypes through inappropriate stocking practices.

Naturally occurring factors

Cutthroat trout biology

Westslope cutthroat trout possess biological characteristics that make them naturally susceptible to a host of limiting factors. First, the habitat requirements of the subspecies are such that populations typically inhabit coldwater habitat with limited productivity, making them historically subject to thermal and physical isolation (Behnke 2002). Populations appear small and supported by variable numbers of spawners, and so may be subject to stochastic events such as epizootics or catastrophic environmental change (e.g., drought, earthquakes, landslides). The small effective population sizes typical of the species may further predispose them to inbreeding and increased losses of genetic diversity (Amos and Harwood 1998; Vucetich and Waite 2001). Cutthroat trout are subject to significant predation mortality and negative interactions with other salmonids. As well, their well-developed natal philopatry suggests high levels of demographic independence among adjacent populations so that declining populations and extirpated areas are not likely to be recolonized over the short term (see BIOLOGY). The lack of more robust population-specific information has likely contributed to localized declines. Little biological information has been collected for WCT in a consistent and standardized manner over long time periods and no rigorous system is in place to monitor catch/creel results throughout much of the range. Relationships between life history types and their particular habitat requirements are understudied, as is the scope and variation typical of WCT movements, the range and extent of distinct breeding units, or the determinants of substantial population structure in the wild.

 

Anthropogenic factors

The dramatic declines in WCT populations over the last century clearly indicate that the greatest threats to cutthroat trout are the anthropogenic manipulation and degradation of the environment in which it lives (Allendorf and Leary 1988; Liknes and Graham 1988; Nehlsen et al. 1991; Slaney et al. 1996; Johnson et al. 1999; Shepard et al. 2003). Throughout its range, the number and distribution of populations have steadily declined in response to the cumulative effects of habitat loss and degradation, overexploitation, and detrimental interactions with introduced species (i.e., competition, predation, hybridization).

Climate change

It is likely that climate change brought on by global warming may play an important role in further limiting the distribution of WCT in the future. The Canadian climate in 1998 was the warmest on record, in one of the warmest decades on record. This may pose a problem for cutthroat trout, which are a coldwater-adapted species. Westslope cutthroat trout are associated with water temperatures less than 16°C at all life history stages (Behnke 1992; McIntyre and Rieman 1995) and the ‘critical thermal maximum’ for WCT 27°C has been reported to be lower than those estimated for brook trout and rainbow trout:  29.8°C and 31.6° C, respectively (Feldmuth and Eriksen 1978 cited in McIntyre and Rieman 1995).  Increasing water temperatures resulting from global warming may, therefore, give non-native fish a competitive advantage over WCT in marginal habitats.  A summary of available climate change models suggests that mean air temperatures in the Pacific Northwest could increase by 2 - 5°C in the next 50-100 years (Neitzel et al. 1991). In the Rocky Mountain region, one study estimated that an increase of as little as 1°C in mean July air temperatures would reduce the geographic area of suitable salmonid habitat by 16.8%, and a 5°C increase in mean air temperature would reduce the amount of habitat by 71.8% (Keleher and Rahel 1996).  In particular, a recent trend analysis of daily average temperatures found that from 1895 to 1995, the Southern Interior Mountain region (which contains the core of WCT in BC) has increased in average summer temperatures by 1.2oC (BC Ministry of Environment 2006). Increasing temperatures are thought to be at least in part responsible for the massive infestations of mountain pine beetle (Dendroctonus ponderosae) much of BC is currently experiencing.  These infestations are expected to affect stream bed substrate composition (including sedimentation), channel morphology, large woody debris presence and water temperatures over time -- all key features affecting WCT habitat suitability.

The potential impacts of climate change are not trivial as they will affect the related aspects of precipitation pattern changes, hydrologic changes, stream morphology changes and loss of glaciers which provide summer flows in many important WCT streams such as the Bull, White and Upper Kootenay Rivers.

Habitat loss

As noted above (Habitat trends), timber extraction, mining, and hydro-electrical developments have been responsible for loss and degradation of WCT habitat and decline of several populations (e.g., Joseph Creek, Spray and Kananaskis rivers). The road networks associated with primary resource extraction have encroached upon untold numbers of streams, causing many to be culverted or otherwise altered. As well, it has led to an explosion of access points for angling and recreational activities (off-roading, ATV use) which further serve to degrade sensitive habitats. Protected areas exist within the range of WCT in Canada, but they are often small and do not necessarily encompass all the habitats required by the various life history forms within an area (particularly migratory forms). It is apparent that in the absence of more rigorous protection, required habitat will continue to be degraded and populations increasingly fragmented.

While the exact nature of their movements is relatively unknown for many populations in Canada, it seems likely that WCT are adapted to move during moderate to high flow events. Their movements often coincide with the rising limb and peak of the hydrograph, allowing them to negotiate seasonal barriers within streams where increased flows may be necessary to gain access (Brown and MacKay 1995a; Schmetterling 2001). This, of course, has obvious implications for landscape planning, road building, and long-term population viability (Hilderbrand and Kershner 2000b). While it is apparent that WCT can and do move significant distances to find required habitat, migration of this type is dependent on the preservation of suitable migration corridors between habitat types. Unfortunately, many culverts may not be designed to accommodate fish passage at high flows. Culverted crossings of spawning tributaries must allow for fish passage under a range of different flow conditions. The dramatic decline of anadromous and fluvial populations throughout the lower Columbia River and parts of Alberta attests to the profound influence of migration barriers on those systems (e.g., Nehlsen et al. 1991). The loss of these migratory forms may be particularly egregious, as it tends to limit the recolonization potential of areas with locally extirpated resident populations. Because many such populations appear to be demographically independent, local declines or extirpations are not likely to be reversed by immigration from even nearby populations.

The main causes of habitat degradation cited above are forest harvesting, mining and urbanization. While it is true that forest harvesting, urbanization and mining cause major impact, the current pressure from development of rural lands and foreshore areas (lakes and rivers) by the recreation industry – golf courses, ski hills, resort and summer home development, marinas, docks, retaining walls, beach development and other foreshore development and increasing boat use must not be overlooked. This industry extracts significant volumes of surface water, hardens shorelines, removes riparian vegetation and degrades water quality. It should be noted that there is no end in sight to this development pressure and the impacts are expanding from site-specific to watershed level (Bruce Macdonald, Habitat and Enforcement Branch, Department of Fisheries and Oceans, Nelson, BC, personal communication 2006).

In addition there is increasing and extensive mining exploration and the development of new mines particularly coal mines. The impacts of coal mines were reported above (see Habitat trends), but the scale of the present and future impacts of mining development in the upper Elk River was not fully discussed.  In some tributaries of the Elk River entire headwater reaches have been annihilated and the populations fragmented by rock drains. It is probable that these headwater reaches contain valuable genetically pure populations. Large areas of the Flathead River and Elk River basins, two of the main WCT river systems in BC, are now being considered for coal bed methane development (Macdonald, pers. comm. 2006).

Transportation infrastructure, roads and railways, is expanding and in many cases parallels WCT streams resulting in hardening of the stream banks to protect the infrastructure. One example of this increasing development is that there have been several derailments of coal trains in the last year all causing some impact to streams and fish habitat (Macdonald, pers. comm. 2006).

Agriculture development is one of the most significant users of surface water in the WCT range. Agriculture has also been responsible for extensive removal of riparian habitat along WCT streams.  Also, the possible effects of an expanding independent power production industry should not be overlooked.  Many of these proposals are for headwater systems, which may appear to be removed from main fish streams but in fact propose to dewater higher elevation and headwater stream reaches (Macdonald, pers. comm. 2006).

Overharvesting

Cutthroat trout are a popular sport fish in western Canada, perhaps second only to rainbow trout in terms of angler interest. Like many other sport fish in Canada, angling pressure is likely a significant factor limiting natural production (reviewed by Post et al. 2002). The overfishing of native fish stocks began with European settlement of western Canada in the 1880s and many CT populations declined during a hundred-year period of liberal fishing regulations in the region (Mayhood 2000). The sometimes voracious feeding habits of cutthroat trout and their accessibility in smaller systems make them particularly susceptible to over harvesting. As early as the 1950s, significant declines were noted in Canadian populations and were most pronounced near urban areas where human population densities are greatest. Harvest rates in Alberta during the 1980s averaged ~11 million fish per year (some 7 million kg) with a large proportion of that being trout (Nelson and Paetz 1992). Since that time, more restrictive catch regulations have been implemented in Alberta (as they have in BC); however, this may be an indication of the kind of pressures faced by native fish populations throughout western Canada as human populations continue to grow and harvest pressures increase.  While catch-and-release fisheries have been implemented in particularly sensitive areas and have stemmed declines in some cases (Baxter, pers. comm. 2004), hooking mortality following release may have a significant impact on populations which have already been marginalized by habitat loss or the introduction of non-native fishes (Marnell and Hunsaker 1970; Nehlsen et al. 1991; Slaney et al. 1996). Even when not directly targeted by fishing efforts, WCT may often be subject to significant bycatch mortality in other fisheries.

Introductions

One of the greatest threats facing native populations of WCT in Canada is the harmful effect of introductions, especially of hatchery-origin salmonids. The natural fecundity of fish (and the relative ease with which their reproductive cycle can be manipulated) has made the hatchery production of salmonids a common response to declining fish populations and the desire to provide fishing opportunities. However, it is becoming apparent that hatchery fish have been routinely stocked without an understanding of the effectiveness of the transfer, the fate of the released fish, or the impacts on wild populations. In the United States, the introduction of non-native species is believed to be the primary cause for the declines of several inland species of cutthroat trout.  The introduction of hatchery-origin salmonids can result in both genetic (e.g., hybridization, outbreeding depression), and ecological impacts (e.g., displacement, competition, disease) on native cutthroat trout populations, depending on the species introduced.

Hybridization and Introgression

Rainbow trout and Yellowstone cutthroat trout introductions have resulted in significant levels of introgressive hybridization throughout the historic range of WCT (Leary et al. 1984; Leary et al. 1987; Allendorf and Leary 1988, Hitt et al. 2003). Less than 29% of occupied habitats in the United States are believed to support populations at or near the habitat’s potential capacity. Genetic testing suggests that WCT populations may be genetically unaltered in less than 8% of its historical range in the US (Shepard et al. 2003). Hybrid swarms between RBT and WCT have now been documented in both Alberta and British Columbia (Rubidge 2003; Janowicz 2004) and levels of introgression appear to be spreading rapidly among streams and upstream from mainstem rivers (Hitt et al. 2003; Rubidge 2003; Weigel et al. 2002; Janowicz 2004) [see also POPULATION SIZES AND TRENDS section]. The factors influencing the spread of this introgressive hybridization are poorly understood at this time. It has been suggested that since RBT are unlikely to successfully colonize and exploit colder, high elevation habitats (e.g., Paul and Post 2001), the spread of hybridization into high elevation sites may be impeded by natural physical or ecological barriers (Weigel et al. 2002). Considering the widespread history of stocking in Canada and the fact that RBT and non-native salmonids continue to be introduced throughout the native range of WCT in Canada, it is likely that many genetically pure WCT populations are at risk and will increasingly be restricted to isolated headwater streams.

(1) British Columbia: In the upper Kootenay River drainage, it is apparent that introgressive hybridization between introduced RBT and native WCT has increased in the last 15 years. Leary et al. (1987) detected approximately 5% hybridization between WCT and RBT within the White River watershed (a large tributary of the upper Kootenay River). More recent surveys indicated that hybridization has increased since 1986 and has spread to the lower reaches of seven other tributaries including Wild Horse, Mather, Skookumchuk, and Gold creeks, as well as the Elk, St. Mary and Lussier rivers (Rubidge 2003). In the US, Hitt et al. (2003) reported a similar increase in the number of introgressed populations in the upper Flathead drainage (24 of 42 sites (57%); seven more than a previous study in 1984). In both cases, the spread of hybridization appears to be spreading in an upstream direction from the site of most RBT introductions: Lake Koocanusa and Flathead Lake, respectively. Evidence from these areas suggests that the spread of hybridization may not be prevented by the ecological gradients (other than, perhaps, impassible upstream barriers) but is instead related to the distance from the nearest site of stocking.  Table 2 summarizes the numbers of systems reported to contain WCT that have been stocked at least once with either rainbow trout or coastal cutthroat trout in BC.  However, spread to non-stocked systems has undoubtedly occurred where stocking occurs in ‘open’ systems.

(2) Alberta: An early study of hybridization in Alberta found limited evidence for hybridization between native WCT and introduced species. McAllister et al. (1981), examined morphological and biochemical variation in WCT from Banff National Park (10 lakes), Kootenay National Park (Floe Lake), Waterton Lakes National Park (Sofa Creek), as well as a sample from the Connor Lakes in British Columbia. The authors found that 10 of the 13 sites contained pure WCT; two sites were found to contain WCT x Yellowstone CT hybrids (Baker Lake (BNP) and Sofa Lake (WLNP)) and a third, Taylor Lake (BNP), was found to contain a pure introduced population of Yellowstone cutthroat trout. No hybridization with RBT was indicated. However, the morphological comparisons and allozyme markers used in the study appear to have had limited resolution to detect RBT introgression. For example, half of the populations showed no genetic variability at the 10 allozyme loci used and 4 of the 10 loci showed no species-specific diagnostic bands and were unable to distinguish between WCT and RBT.  Furthermore, all samples were collected from alpine systems (elevation > 2000m) and were chosen with the expectation that they would contain pure WCT populations. The stocking of non-native species in the sampled populations was believed to be minimal or non-existent.

Recent genetic testing in Alberta suggested that hybridization is widespread in the eastern slopes of the Rocky Mountains. Janowicz (2004) detected hybridized populations in 13/14 watersheds sampled (see Figure 8). Degrees of hybridization within watersheds ranged from 100% of sampled creeks in Ram River (North Saskatchewan drainage) and Sheep River (South Saskatchewan drainage) to 22% in the Kananaskis River. The Elbow River watershed was the only system in which hybridization was not detected. The severity of hybridization within individual streams varied considerably from one or a few hybrid individuals to those where in excess of 80% of all individuals appear to be of hybrid origin. Many populations, in fact, exhibited highly mixed genotypes (more than 50% with heterospecific alleles) indicating that hybridization was advanced and progressing towards hybrid swarms in these creeks. Hybrid swarms present a great danger to the persistence of native species as the unique genotypes typical of pure parental populations are lost once randomly mating hybrid swarms are formed (Leary et al. 1995).

It should be noted that the low number of reference WCT populations in the study might lead to overestimation of the number of hybrids observed in the sample. Only three reference WCT populations were included so that the number of diagnostic WCT alleles was low (averaging ~ 3.3 per microsatellite locus). As noted earlier, WCT populations are often characterized by unique alleles or those that, while locally abundant, are uncommon over a larger geographic area. For example, Taylor et al. (2003) found the total number of alleles across 29 WCT populations in British Columbia averaged ~ 13 per locus while the average in any one population was less than ~ 4. It may be that WCT alleles present in non-reference populations have been misidentified as RBT alleles when in fact they were not. Countering this potential upwards bias, however, is the fact that many of the sampled streams were not chosen randomly, but with the belief that they contained pure WCT populations. While hybridized populations may be of some importance in terms of fisheries opportunities, their ecological and taxonomic status remains largely unresolved (e.g., US Federal Register 1996; Allendorf et al. 2003). It is clear, however, that extensively hybridized populations are of little conservation value for efforts to preserve pure WCT. As such, every effort should be made to identify and determine the conservation status of the remaining pure unstocked populations in Alberta and to halt any further spread of hybridization.

Outbreeding

Hatchery WCT have been stocked within the native range of WCT in both British Columbia and Alberta to ‘supplement’ native production usually for angling purposes (see Table 2 for numbers of systems stocked with WCT in BC).  However, locally adapted biodiversity has not been considered, and no effort to use local stocks has been made as is evident by the very limited source populations used in hatchery production.  For example, British Columbia has relied on a single source (Connor Lake) of WCT for all stockings in the past three decades.  In other western salmonid species, such programs have resulted in increased straying and homogenization of genetic population structure, as well as genetic swamping and outbreeding depression (reviewed by Rhymer and Simberloff 1996, Allendorf et al. 2001).  Since significant genetic substructuring exists for this species, even greater impacts in terms of homogenization and outbreeding depression might be expected. However, no such evaluations have been done for WCT, and there is very little information available to determine how many native populations have been supplemented with hatchery WCT.  Thus the degree to which this impact might affect WCT populations in Canada is unknown.

Ecological impacts

While it is unclear whether other species of introduced salmonids actively displace native cutthroat or simply replace WCT populations depressed by other factors, it is clear that introductions of non-native brook trout have typically resulted in range constriction or elimination of cutthroat trout from large portions of their native habitat (Donald 1987; Fausch 1989; Griffith 1988).  Non-native brook trout have been stocked throughout much of the WCT native range in British Columbia and Alberta.  Brook trout appear to effectively displace or replace WCT, particularly at low elevation locations in Alberta (Paul and Post 2001, see examples under POPULATION SIZES AND TRENDS), contributing to the present restriction of WCT to mainly isolated higher elevation headwaters here.  Similar patterns have not been recorded in British Columbia but may be present in some systems.

Finally, a number of other non-salmonid species have been introduced via authorized and unauthorized means in both provinces. In particular, walleye (Stizostedion vitreum), smallmouth (Micropterus dolomieui) and largemouth bass (M. salmoides), yellow perch (Perca flavens), and northern pike (Esox lucius) have been documented in a number of systems within the native WCT range (Pollard, pers. comm.).  These species are all predatory and most have been implicated in salmonid declines in inland waters of the US (Fuller et al. 1999).

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