BY AMY L. HAAK and JACK E. WILLIAMS
April 2015 | Volume 21, Number 1
As we celebrate the 50th anniversary of the Wilderness Act in the United States, conservationists across the country can be proud of the 757 areas encompassing nearly 110 million acres (44,515,420 ha) that currently comprise our National Wilderness Preservation System (NWPS). However, in the contiguous United States this represents less than 3% of the land base, and as the number of threatened and endangered species continues to rise, it is important to assess the role and effectiveness of wilderness preservation in species conservation, particularly in an era of rapid environmental change.
A federal designation of wilderness is the highest form of land protection provided to any federal wildland and as such could provide a safety net for the nation’s biodiversity. However, a history of protecting higher-elevation scenic areas while many of our more biologically diverse environments have gone unprotected (Groves et al. 2000; Scott et al. 2001) has diminished the potential effectiveness of the NWPS in biodiversity conservation. While this bias has been recognized for more than a decade, current efforts to rectify the situation are often stymied by divisive politics over wilderness preservation and increasing development pressure on the more productive mid- and lower-elevation public lands. Rapid environmental change and increasing uncertainty due to climate change have added to the complexity of prioritizing land protection for biodiversity conservation while simultaneously increasing the need. Recognition of these challenges has led to a growing emphasis on strategic conservation planning at the landscape scale.
Although there are numerous approaches to systematic conservation planning for biodiversity conservation (Pressey et al. 1993; Margules and Pressey 2000), most are based on terrestrial species and habitats. The protection of freshwater habitats is seldom a strategic driver in landscape-scale prioritization processes, and the development of a parallel approach for aquatic ecosystems has been slow to emerge (Linke et al. 2011; Haak and Williams 2013). The linear nature of rivers and the interconnectedness of drainage systems provide many challenges to the management of protected areas that are typically based on terrestrial features and land ownerships (Williams et al. 2011). Therefore, it is not surprising that nearly 40% of freshwater and diadromous fish species native to North America are at risk of extinction (Jelks et al. 2008), as are nearly half of all freshwater crayfishes (Taylor et al. 2007) and two of every three species of freshwater mussels (Williams et al. 1993), despite expenditures for aquatic threatened and endangered species that exceed their terrestrial counterparts (Williams et al. 2011).
Before looking ahead to the next 50 years of wilderness protection, we first take a look back at how well the previous 50 years have protected freshwater biodiversity. To do this we use the historical range (circa 1850) and current distribution of native trout to provide an overview of how well the existing wilderness system protects aquatic biodiversity. Based on our findings, we propose an integrated approach to conservation planning that explicitly incorporates freshwater habitats into landscape-scale conservation initiatives in a manner that will not only increase the resilience of native aquatic species to climate change but will also improve the aquatic integrity of existing wilderness areas.
Native Trout as Indicators of Aquatic Diversity
We chose native trout as a proxy for freshwater biodiversity in our analysis of the wilderness system for several reasons. First is the simple fact that native trout are in trouble – nearly every species and subspecies is in decline and some are listed pursuant to the Endangered Species Act, while the remainder are considered sensitive species by state and federal agencies. Therefore native trout should be target species for land protection strategies that encompass cold-water habitats. Second, native trout tend to be sensitive to environmental disturbance, especially climate change, and their presence or absence is a good barometer of local and watershed-scale habitat conditions. Third, we lack distribution and status information for many aquatic species, whereas trout are relatively well studied and broadly distributed with detailed spatially explicit population information available from state and federal agencies.
In cold-water habitats, native trout function as a keystone species, serving as both the top predator in the aquatic system as well as prey for a variety of terrestrial species (Koel et al. 2005; Varley and Schullery 1998). Their diverse life history strategies have not only enabled trout to survive and prosper for 10,000 years but also serve an important ecological role. Spawning runs of migratory populations move nutrients upstream from lakes and rich valley bottom habitats to headwaters where they are available for use by local flora and fauna (Tronstad 2008), while their facultative migratory behavior enables them to abandon habitats rendered unsuitable by a disturbance event and subsequently recolonize these habitats once they recover (Rieman and Dunham 2000). The larger size of these migratory fish also contributes to increased fecundity and resistance to nonnative predators (Figure 1).
Unfortunately, the broader ecological value of native trout is less commonly realized today because of habitat fragmentation and management practices that isolate remaining populations from the larger native fish community. It is a common conservation trade-off for native trout to be isolated in small headwater streams as a means to protect them from invading nonnative fishes in downstream areas (Fausch et al. 2009). Hatchery-based stocking programs for native and nonnative trout also obfuscate the conservation status of populations. Nonetheless, the presence of large, self-sustaining populations of native trout throughout an interconnected stream network is indicative of aquatic systems that are more likely to be ecologically intact than those without these populations.
We use six subspecies of inland cutthroat trout (Oncorhynchus clarkii sspp.) to assess the efficacy of the wilderness system in preserving aquatic diversity. We look first at the question of representation as it pertains to the protection of distinct aquatic communities within the wilderness system of the Intermountain West. This is followed by an evaluation of how well designated areas have retained their “wilderness character” as it pertains to aquatic ecosystems. The results of these analyses are used to inform a broader discussion of the management implications and role of wilderness in the conservation of aquatic diversity.
What Trout Can tell Us About Aquatic Diversity: Representation of Aquatic Diversity within the NWPS
The principle of complementarity is the foundation of many conservation-planning approaches that seek to balance the larger preservation portfolio with the protection of underrepresented elements of diversity. In order to identify what is missing, it is first necessary to assess what has been captured. Historically, the seasonal migration of cutthroat trout from cold headwaters downstream into warmer waters defined a basin-specific assemblage of coevolving aquatic species, including sculpins (Cottus spp.), suckers (Catostomus spp.), chubs (Gila spp.), and dace (Rhinichthys spp.), among others similarly isolated in these river basins. Therefore, the historical distribution of extant cutthroat subspecies within the interior West (Figure 2) provides a reasonable proxy for distinct aquatic communities that have been evolving in isolation from one another since the last glacial retreat in North America more than 10,000 years ago (Behnke 2002).
Table 1 summarizes the percentage of stream habitat historically occupied by each cutthroat subspecies that is currently within the existing NWPS. (Yellowstone, Teton, and Glacier National Parks were also included in the wilderness column due to the wilderness management
emphasis of the National Park Service for each of these parks). As Table 1 shows, habitat associated with Westslope cutthroat is the most prevalent within the wilderness system, both in terms of actual miles protected (7,080 [11,394 km]) as well as a percentage of the total historical habitat (13%). In contrast, Bonneville cutthroat trout (BCT) habitat has very little representation within the current wilderness system (less than 1%), and so it may then be surmised that habitat for other native fishes that evolved with BCT in the Bonneville basin is also lacking wilderness protection.
Preserving the “Wilderness Character”
The Wilderness Act’s mandate to preserve the “wilderness character” of designated areas provides the foundation for management policies that restrict resource development on wilderness lands and seeks to maintain the values that made an area worthy of designation. By definition, a wilderness area is an “area of undeveloped Federal land retaining its primeval character and influence … which is protected and managed so as to preserve its natural conditions” (the Wilderness Act, 1964, Section 2(c)). Therefore, we can generally assume that aquatic habitat within a designated wilderness area is high quality and less likely a limiting factor for resident species. However, the interconnected nature of aquatic ecosystems makes them vulnerable to both upstream and downstream impacts outside of the wilderness boundary. This is particularly problematic for highly mobile species such as trout. The presence or absence of native trout within their historical habitat is a good indicator of the health of the larger aquatic ecosystem both within and beyond the wilderness boundary.
Invasive aquatic species, particularly nonnative fishes, complicate this picture. Nonnative trout have been widely introduced throughout the western United States. Most introductions occur outside of wilderness areas, but the interconnected nature of riverine habitats facilitates upstream invasion into protected zones. In the past, there have been many introductions of nonnative trout in historically fishless wilderness lakes, but concerns over amphibians and other species that have been negatively impacted by past introductions (Knapp et al. 2001) have curtailed this practice in recent years.
Table 2 provides the percentage of historical stream habitat within designated wilderness areas that is currently occupied by subspecies of native cutthroat. Here we see that although Yellowstone cutthroat (YCT) and Colorado River cutthroat (CRCT) have a similar amount of historical stream habitat within wilderness areas (more than 2,000 miles [3,000 km]), there is a significant difference in the amount of that habitat that is still occupied. Only 14% of the CRCT historical habitat is still occupied, while 97% of the YCT historical habitat continues to support populations of native cutthroat. Although not included in our summary of stream and river habitat, large wilderness lakes, such as Yellowstone Lake, also provided habitat for YCT and Westslope cutthroat, but CRCT and most other cutthroat subspecies lacked these large lake habitats. Very little of the protected historical habitat for Rio Grande and Lahontan cutthroat is currently occupied. The loss of these native fishes from their historical habitat in protected areas is likely the result of displacement by nonnative species, downstream barriers that prevent fish from accessing the wilderness waters, or degraded conditions external to the wilderness that have rendered the entire watershed unsuitable for cutthroat habitat. Although nonnative trout (e.g. rainbow, brook, brown trout) may now occupy these streams, they do not serve the same ecological function as the native species and may jeopardize the long-term viability of the larger native assemblage that evolved with the cutthroat.
While the presence of native trout is a good indicator of habitat quality and watershed conditions, it is not necessarily indicative of natural ecological processes that are also important to the preservation of an area’s wilderness character. Aquatic ecosystems that support small isolated resident populations rather than the historically migratory populations have been diminished in several ways. The loss of the large spawning runs has eliminated an important source of seasonal food for a diverse array of aquatic and predators and reduced the delivery of nutrients to the typically nutrient-poor headwater streams. The remaining small isolated populations have little resiliency to environmental disturbances and are at increased risk of extirpation from wildfire, flood, and drought (Haak and Williams 2012). Historically, migratory populations would have been able to vacate an area rendered unsuitable and thus survive the disturbance. Populations could also have recolonized the habitat once it recovered. The ability to repopulate disturbed sites is particularly important for wilderness areas where disturbances are increasing and recovery is driven by natural processes rather than management intervention. Without migratory populations and interconnected habitat, the aquatic diversity of disturbed watersheds may not be restored.
Populations of native trout that occupy high quality interconnected stream habitat are considered stronghold populations. These populations are less vulnerable to climate change than small isolated populations because populations are larger, occupied habitats are more diverse, and fish are able to move in response to changing environmental conditions. For the purposes of this analysis we refer to Haak and Williams (2012) and classify those populations occupying at least 17.25 miles (27.8 km) of interconnected stream habitat in a patch of at least 24,700 acres (10,0000 ha) as strongholds.
We use the presence of stronghold populations as a proxy for aquatic systems that are most likely to have retained some of the ecological role of native trout within the broader ecosystem, a factor that is important to retaining the “wilderness character” of protected areas as discussed earlier. Table 3 provides the percentage of the occupied stream habitat within designated wilderness areas that supports at least some portion of a stronghold population. Here we find that, with the exception of Westslope and Yellowstone cutthroat, the presence of native trout strongholds within the NWPS is negligible. This is consistent with previous analyses that found few strongholds rangewide among extant populations of cutthroat trout (Haak and Williams 2013). Trout populations that are isolated in small stream habitats above barriers lack resilience to environmental change and have a diminished ecological role. In contrast, a stronghold population, even one that is not completely contained within a wilderness, still provides ecological services to the wilderness area as it utilizes different life history strategies and moves between varieties of habitats within the larger drainage network. As a result, it is less likely to be extirpated due to a wildfire or flood and can recolonize the disturbed site once it has recovered.
Management Implications for the Next 50 Years
The results of our analysis highlight some shortcomings within the existing NWPS as it relates to the preservation of aquatic biodiversity and the integrity of aquatic ecosystems within wilderness areas.
• A lack of representation exists for some unique aquatic ecosystems such as the Bonneville basin where less than 1% of the historical range for Bonneville cutthroat trout is contained within a wilderness area (Table 1). The Great Basin is also significantly underrepresented with just 6% of the historical habitat for Lahontan cutthroat. Future additions to the NWPS should encompass these unique and underrepresented aquatic communities. Although the remaining habitat is often fragmented and large-scale land protection opportunities may be limited, at a minimum the protection of important headwater streams can secure a clean source of cold water for resident fish as well as accrue benefits to species in the downstream nonwilderness reaches.
• Many of the wilderness areas that do capture important cold-water habitat have not preserved their “wilderness character” due to threats, such as nonnative species that have undermined the integrity of native trout populations, but are frequently outside the purview of wilderness managers to address. The percentage of historical habitat within wilderness areas that is currently occupied by native cutthroat ranges from a low of 10% for Lahontan cutthroat to a high of 97% for Yellowstone cutthroat (Table 2). Colorado River cutthroat have more than 2,000 miles (3,000 km) of historical habitat in wilderness areas, yet only 313 miles (505 km) are currently occupied by the native trout. Much of the remaining habitat, although of high quality, is likely occupied by nonnative trout that have displaced CRCT. Where nonnative species can be controlled, wilderness management should include opportunities to extend and reconnect isolated populations as well as restore native cutthroat to their historical habitat.
• The challenge of preserving ecological functions becomes more problematic in smaller protected areas. The large protected landscapes in Idaho and Wyoming at the core of Yellowstone cutthroat and Westslope cutthroat habitat have thus far enabled these fish to retain their migratory life history and ecological role with 87% and 95%, respectively, of their wilderness habitat supporting stronghold populations (Table 3). In contrast, just 31% of the occupied CRCT habitat within wilderness areas is associated with a stronghold population, and the number drops to only 8% for Rio Grande cutthroat. Although largely undocumented, spawning runs of native trout from lakes and main stem river systems into high mountain wilderness areas may have been a significant source of nutrients and productivity for higher-elevation stream and riparian areas. The loss of this ecological function and the increased vulnerability of small populations to climate change threaten the integrity of aquatic systems in wilderness areas. Increasing the resiliency and restoring the ecological role of native trout within many wilderness areas will require protecting and restoring stronghold populations that extend beyond the wilderness boundary.
Addressing the shortcomings described here necessitates an integrated landscape-scale approach to conservation planning that extends beyond existing and proposed wilderness boundaries. Many wilderness areas are too small to adequately protect an area’s biological diversity totally within the wilderness boundary, and opportunities to protect large landscapes today are limited. The linear nature of rivers and connectivity of drainage systems compounds these challenges for the protection of aquatic biodiversity.
Rewilding, as first described by Soulé and Noss in 1998, promotes not only the protection of core reserves such as wilderness areas but also stresses the importance of connectivity among reserves for preserving the ecological role of large predators as they move among core areas. Although Soulé and Noss’s focus was on terrestrial species, the concept is also applicable to aquatic ecosystems and provides a useful framework for aquatic conservation planning. Restoring native trout populations by reconnecting habitat in a large enough landscape to allow for the full expression of their life history diversity may be the key to conserving aquatic biodiversity in a future characterized by climate change and increasing disturbances such as wildfire, floods, and drought (Heller and Zavaleta 2009; Lawler 2009).
Large wilderness areas such as those found in central Idaho and western Wyoming are expansive enough to support migratory populations of Westslope and Yellowstone cutthroat wholly within the wilderness boundary. However, there have been few opportunities to protect landscapes at this scale, so populations are either small and isolated within wilderness areas or they extend beyond the wilderness boundary where they are vulnerable to anthropogenic impacts (e.g., roads, diversions, dams). The High Uintas Wilderness in northeastern Utah is illustrative of this point and underscores the importance of connectivity beyond reserves as advocated by Soulé and Noss.
The High Uintas is typical of many wilderness areas in that it captures the top of the mountain range, truncating the watersheds that drain from the high peaks. The wilderness area’s 453,500 acres (183,525 ha) encompasses the headwater tributaries of 12 populations of Colorado River cutthroat trout and two populations of Bonneville cutthroat trout (Figure 3). Although the entire historical habitat for BCT is currently occupied, CRCT are only found in 118 miles (190 km) of their 280 miles (450 km) of historical habitat. Most of the unoccupied habitat drains the south side of the mountain range and could provide restoration opportunities if the limiting factors outside of the wilderness boundary can be addressed. The 825,000 acres (333,866 ha) of roadless lands surrounding the High Uintas provides additional high quality habitat and more management flexibility for addressing threats posed by nonnative species and altered habitat.
Of the nine populations that drain the north side of the Uintas, five are classified as strongholds (two BCT and three CRCT). These populations extend beyond the wilderness boundary onto the roadless lands as well as multiple use federal land and private land. Protecting these populations downstream of the wilderness boundary, as well as looking for opportunities to establish larger metapopulations by reconnecting existing populations is important for protecting and restoring the integrity and resilience of aquatic ecosystems within the wilderness area.
In reviewing the status of native trout and wilderness conservation, it has become clear that wilderness areas can provide important stronghold habitat in some cases and sources of clean, cold water in others, but the restoration of large, resilient native trout populations must extend well beyond most wilderness boundaries and involve private and public landownerships. Williams et al. (2011) advocate for a management approach that protects entire watersheds and fish communities as Native Fish Conservation Areas (NFCAs). NFCAs are intended to complement existing conservation efforts by protecting and restoring native aquatic communities at the watershed scale while allowing for compatible uses across multiple management jurisdictions. While NFCAs provide management flexibility, ensuring the long-term persistence of intact aquatic communities requires that four critical elements be met:
1. Watershed boundaries should be large enough to maintain the natural processes that shape the aquatic habitat and provide resistance and resilience to disturbances such as wildfire and flood.
2. The NFCA should encompass all habitats necessary to support historical life histories (e.g., fluvial, adfluvial) of native species present as well as to complete a species life cycle (e.g., spawning, overwintering, migration).
3. The NFCA should be large enough to support sufficiently large populations of native species that have a high likelihood of long-term persistence.
4. Management plans and/or agreements should ensure that the NFCA will be managed for the benefit of the aquatic ecosystem and the species it supports in perpetuity.
Dauwalter et al. (2011) apply the NFCA approach to the Upper Colorado River drainage, using the distribution of Colorado River cutthroat trout and three species of warmer water fishes as a means to identify potential NFCAs within the basin. They describe a process of focusing stream restoration and reconnection projects in those remaining watersheds that still contain fragmented populations of genetically pure CRCT in the headwaters as well as warmer water stream fishes farther downstream. The NFCAs identified include the Henrys Fork watershed, which drains the northeastern portion of the High Uintas Wilderness and encompasses two of the CRCT strongholds previously discussed (Figure 3). Effective implementation of the NFCA approach in the Henrys Fork will require coordination with federal land managers and private landowners beyond the wilderness boundary. However, if it is successful, the benefits of restoring an intact aquatic community will accrue throughout the watershed and increase the resilience of the CRCT population within the wilderness.
The next 50 years of wilderness protection promise to be challenging as the human population approaches 9 billion and the impacts of climate change increase. Given the already rapid decline of freshwater species in spite of significant expenditures and more than 100 million acres of protected lands, conservationists must take an integrated approach to land conservation that moves beyond management boundaries and encompasses watersheds and ecological processes. Wilderness can still play an important role in the protection of core refugia and cold-water supplies, but restoring larger interconnected populations of native trout requires a landscape vision that incorporates a variety of strategies. Without this, even our wilderness areas are at risk of losing the species and diversity they were intended to protect.
Behnke, R. J. 2002. Trout and Salmon of North America. New York: The Free Press.
Dauwalter, D. C., J. S. Sanderson, J. E. Williams, and J. R. Sedell. 2011. Identification and implementation of Native Fish Conservation Areas in the Upper Colorado River Basin. Fisheries 36: 278–288.
Fausch, K. D., B. E. Rieman, J. B. Dunham, M. K. Young, and D. P. Peterson. 2009. Invasion versus isolation: Trade-offs in managing native salmonids with barriers to upstream movement. Conservation Biology 23: 859–870.
Groves, C. R., L. S. Kutner, D. M. Stoms, M. P. Murray, J. M. Scott, M. Schafale, A. S. Weakley, and R. L. Pressey. 2000. Owning up to our responsibilities: Who owns lands important for biodiversity? In Precious Heritage: The Status of Biodiversity in the United States, ed. B. A. Stein et al. (pp. 275–300). New York: Oxford University Press.
Haak, A. L., and J. E. Williams. 2012. Spreading the risk: Native trout management in a warmer and less certain future. North American Journal of Fisheries Management 32: 387–401.
Haak, A. L., and J. E. Williams. 2013. Using native trout restoration to jumpstart freshwater conservation planning in the Interior West. Journal of Conservation Planning 9: 38–52.
Heller, N. E., and E. S. Zavaleta. 2009. Biodiversity management in the face of climate change: A review of 22 years of recommendations. Biological Conservation 142(1): 14–32.
Jelks, H. L., S. J. Walsh, N. M. Burkhead, S. Contreras-Balderas, E. Diaz-Pardo, D. A. Hendrickson, J. Lyons, N. E. Mandrak, F. McCormick, J. S. Nelson, S. P. Platania, B. A. Porter, C. B. Renaud, J. J. Schmitter-Soto, E. B. Taylor, and M. L. Warren, Jr. 2008. Conservation status of imperiled North American freshwater and diadromous fishes. Fisheries 33: 372–407.
Knapp, R. A., P. S. Corn, and D. E. Schindler. 2001. The introduction of nonnative fish into wilderness lakes: Good intentions, conflicting mandates, and unintended consequences. Ecosystems 4: 275–278.
Koel, T. M., P. E. Bigelow, P. D. Doepke, B. D. Ertel, and D. L. Mahony. 2005. Nonnative lake trout result in Yellowstone cutthroat trout decline and impacts to bears and anglers. Fisheries 30: 10–19.
Lawler, J. J. 2009. Climate change adaptation strategies for resource management and conservation planning. Annals of the New York Academy of Sciences 1162: 79–98.
Linke, S., E. Turak, and J. Nel. 2011. Freshwater conservation planning: The case for systematic approaches. Freshwater Biology 56: 6–20.
Margules, C. R., and R. L. Pressey. 2000. Systematic conservation planning. Nature 405: 243–253.
Pressey, R. L., C. J. Humpheries, C. R. Margules, R. I. Vane-Wright, and P. H. Williams. 1993. Beyond opportunism: Key principles for systematic reserve selection. Trends in Ecology and Evolution 8: 124–128.
Rieman, B. E., and J. B. Dunham. 2000. Metapopulations and salmonids: A synthesis of life history patterns and empirical observations. Ecology of Freshwater Fish 9: 51–64.
Scott, J. M., M. Murray, R. G. Wright, B. Csuti, P. Morgan, and R. L. Pressey. 2001. Representation of natural vegetation in protected areas: Capturing the geographic range. Biodiversity and Conservation 10: 1297–1301.
Soulé, M., and R. Noss. 1998. Rewilding and biodiversity: Complementary goals for continental conservation. Wild Earth 8: 18–28.
Taylor, C. A, G. A. Schuster, J. E. Cooper, R. J. DiStefano, A. G. Eversole, P. Hamr, H. H. Hobbs, III., H. W. Robison, C. E. Skelton, and R. F. Thoma. 2007. A reassessment of the conservation status of crayfishes of the United States and Canada after 10+ years of increased awareness. Fisheries 32: 372–389.
Tronstad, L. M. 2008. Ecosystem consequences of declining Yellowstone cutthroat trout in Yellowstone Lake and spawning streams. Unpublished PhD diss., University of Wyoming.
Varley, J. D., and P. Schullery. 1998. Yellowstone Fishes: Ecology, History, and Angling in the Park. Mechanicsburg, PA: Stackpole Books.
Williams, J. D., M. L. Warren, Jr., K. S. Cummings, J. S. Harris, and R. J. Neves. 1993. Conservation status of freshwater mussels of the United States and Canada. Fisheries 18: 6–22.
Williams, J. E., R. N. Williams, R. F. Thurow, L. Elwell, D. P. Philipp, F. A. Harris, J. L. Kershner, P. J. Martinez, D. Miller, G. H. Reeves, C. A. Frissell, and J. R. Sedell. 2011. Native Fish Conservation Areas: A vision for large-scale conservation of native fish communities. Fisheries 36: 267–277.
AMY L. HAAK is a conservation biologist and resource information director with Trout Unlimited’s national science program and also serves on the College of Science Advisory Board at the University of Idaho; email: email@example.com.
JACK E. WILLIAMS is the senior scientist for Trout Unlimited and also serves on the board of directors for Western Rivers Conservancy; email: firstname.lastname@example.org.