Biological diversity: discovery, science, and management in this issue

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Figure 2. The map shows the locations of catchments (i.e., watersheds) in Yellowstone and Grand Teton National Parks that are used for long-term monitoring of amphibians. The maximum number of breeding amphibian species observed in a catchment (species richness) is shown by the outer circles, with the proportion of dry wetlands (proportion dry) indicated by the inner circles. The circles summarize results from surveys conducted from 2006 to 2013. Red circles indicate amphibian “hotspots,” where four amphibian species have been documented as breeding in a catchment.


Species summary

The boreal chorus frog was the most common species encountered during this eight-year period and breeding was detected in an average of 23 (range 19–26) catchments annually. The Columbia spotted frog was also widely distributed and breeding was detected in 20 (18–22) catchments each year. The western tiger salamander and boreal toad were less widespread and breeding was detected in 10 (7–14) and 4 (3–6) catchments annually, respectively. No northern leopard frogs or spadefoots were observed.

Amphibian breeding hotspots

Just fewer than half (15 of 31) of all catchments surveyed contained three breeding amphibian species for at least one year of monitoring (fig. 2). Surprisingly, all four widely distributed species were documented only in four catchments (fig. 2, shown in red). Three of these catchments were located in Yellowstone’s Northern Range and one was in Grand Teton’s Snake River Valley. Across all years, higher-elevation regions (>2,500 m or 8,200 ft) had the lowest amphibian richness. In general, we discerned a weak but significant inverse relationship (r = -0.373, P = 0.030) between elevation and average annual amphibian breeding richness.

Variations in amphibian breeding richness through time

Amphibian hotspots (catchments with four species breeding in at least one year) in Yellowstone’s Northern Range fluctuated from two to four breeding amphibian species during this eight-year period (fig. 3). A synchronous drop in breeding richness occurred in 2007 at these hotspots; however, the identity of species that did not breed varied by catchment. The 2007 drop was followed by a synchronous increase in 2008. After 2008, breeding species rich ness varied annually but lacked synchrony among these hotspots. The only hotspot in Grand Teton varied from two to four breeding amphibian species. In this catchment, breeding richness declined to two species in 2007, returned to three species in 2008, and increased to four species in 2012 when boreal toad breeding was detected for the first time (fig. 3).

Wetland desiccation

The proportion of dry wetlands ranged from 0 to 1 in monitored catchments. The median proportion of dry wetlands within a catchment was 0.40, indicating 40% of available wetlands within that catchment were dry at least once in the 2006–2013 period. Catchments in the Northern Range contained few wetlands altogether (6.0 ± 0.9; mean ± 1 SD), and four of six Northern Range catchments had a high proportion of dry wetlands (=0.57; fig. 2). In contrast, catchments in lower-elevation regions (<2,250 m or 7,380 ft) of Grand Teton generally contained more wetlands (five of six catchments contained =14 wetlands) and a much lower proportion of dry wetlands (=0.36). Catchment elevation and the proportion of dry wetlands were not correlated (r = −0.097, P = 0.591).

Amphibian breeding hotspots and wetland desiccation

The proportion of dry wetlands in documented amphibian hotspots ranged from 0.17 to 0.83. The catchments that had the highest proportion of dry wetlands (0.57 and 0.83) also exhibited the most frequent fluctuations from two to four breeding species (figs. 3B and 3C).

[(Four graphs) The graphs show the annual variation in amphibian breeding richness (left vertical axis) for four amphibian hotspots (colored lines). Hotspots are long-term monitoring catchments that contained four breeding amphibian species in at least one year (see fig. 2). Also shown is the proportion of dry wetlands (right vertical axis) in each catchment summarized by year (gray bars). Catchments summarized in panels A, B, and C are located in Yellowstone’s Northern Range. The catchment shown in panel D is located in the Snake River Valley of Grand Teton National Park. New beaver activity was documented in this catchment in 2012.]

Figure 3. The graphs show the annual variation in amphibian breeding richness (left vertical axis) for four amphibian hotspots (colored lines). Hotspots are long-term monitoring catchments that contained four breeding amphibian species in at least one year (see fig. 2). Also shown is the proportion of dry wetlands (right vertical axis) in each catchment summarized by year (gray bars). Catchments summarized in panels A, B, and C are located in Yellowstone’s Northern Range. The catchment shown in panel D is located in the Snake River Valley of Grand Teton National Park. New beaver activity was documented in this catchment in 2012.


We identified amphibian breeding hotspots in Yellowstone’s Northern Range and in the Snake River Valley of Grand Teton. These areas supported breeding populations of boreal chorus frogs, Columbia spotted frogs, western tiger salamanders, and bo real toads. The latter two species had the patchiest distributions, suggesting that breeding hotspots may be tied to special habitat conditions or may be associated with particular biogeographic conditions (e.g., proximity to glacial refugia).

Our eight-year data set on amphibians underscores the importance of multiyear monitoring for making inferences about amphibian status. We found that annual breeding richness variability can be very high and fluctuated by as much as 75% in some years and catchments. Importantly, annual fluctuations in the number of species breeding were common among Northern Range catchments, a region where wetland desiccation has been well documented (McMenamin et al. 2008; Schook and Cooper 2014). Basing inferences on amphibian status on only 2007 data, for example, would provide an underestimate and a potentially incorrect interpretation of amphibian breeding richness. The high annual variability across the region emphasizes the need for multiple years of sampling to accurately describe amphibian richness and, potentially, overall biodiversity.

Climate-driven wetland desiccation has been implicated in changes to amphibian richness in Yellowstone’s Northern Range (McMenamin et al. 2008). In Wyoming, low-elevation wetlands have the greatest desiccation risk because they typically have higher air temperatures and lower precipitation than higher-elevation wetlands (Copeland et al. 2010). We found that wetland desiccation is proportionally high in the Northern Range and is widespread across Yellowstone and Grand Teton (Ray et al. in press), but that elevation alone did not explain differences in the proportion of those that were dry among catchments. This is likely because some wetlands may be connected hydrologically to permanent water sources (e.g., the Snake River) or are made resistant to desiccation by beaver activity, which can impound and store water even during dry years. Interestingly, beaver activity was documented in two catchments since 2012, and in both catchments boreal toad breeding occurred at the newly created or expanded wetlands.


Our amphibian and wetland monitoring efforts indicate that amphibian breeding hotspots in the Yellowstone Northern Range are vulnerable because they occur in a region with few wetlands and high susceptibility to wetland drying. Breeding hotspots in Grand Teton are less vulnerable to wetland drying because they occur in the Snake River Valley, where there are more wetlands per catchment, where some wetlands have a hydrological connection to permanent waters, and where beavers have been active recently. In the Northern Range and other areas that are susceptible to wetland drying, monitoring and vulnerability modeling can be helpful strategies to increase awareness of the potential for climate effects on amphibians and wetlands. In addition, adaptation strategies, including the removal of other stressors in permanent wetlands (e.g., nonnative fish; Ryan et al. 2014), can help increase amphibian resiliency. Another management option that may increase wetland resiliency is protection of beaver dams and, where possible, beaver establishment (see McKinstry and Anderson 1999 for attitudes regarding beaver management). Increasing resiliency and growing awareness are just two of the primary tenets of adaptation planning (Heller and Zavaleta 2009) that can help to conserve some of the most biologically rich yet climate change–vulnerable resources.

Literature cited

Adams, M. J., D. A. W. Miller, E. Muths, P. S. Corn, E. H. Campbell Grant, L. L. Bailey, G. M. Fellers, R. N. Fisher, W. J. Sadinski, H. Waddle, and S. C. Walls. 2013. Trends in amphibian occupancy in the United States. PLoS ONE 8(5):e64347.

Copeland, H. E., S. A. Tessman, E. H. Girvetz, L. Roberts, C. Enquist, A. Orabona, S. Patla, and J. Kiesecker. 2010. A geospatial assessment on the distribution, condition, and vulnerability of Wyoming’s wetlands. Ecological Indicators 10:869–879.

Fancy, S. G., J. E. Gross, and S. L. Carter. 2009. Monitoring the condition of natural resources in U.S. national parks. Environmental Monitoring and Assessment 151:161–174.

Gould, W. R., D. A. Patla, R. Daley, P. S. Corn, B. R. Hossack, R. Bennetts, and C. R. Peterson. 2012. Estimating occupancy in large landscapes: Evaluation of amphibian monitoring in the Greater Yellowstone Ecosystem. Wetlands 32:379–389.

Guzy, J. C., E. D. McCoy, A. C. Deyle, S. M. Gonzalez, N. Halstead, and H. R. Mushinsky. 2012. Urbanization interferes with the use of amphibians as indicators of ecological integrity of wetlands. Journal of Applied Ecology 49:941–952.

Hamilton, A. J. 2005. Species diversity or biodiversity? Journal of Environmental Management 75:89–92.

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 142:14–32.

Koch, E. D., and C. R. Peterson. 1995. Amphibians and reptiles of Yellowstone and Grand Teton National Parks. University of Utah Press, Salt Lake City, Utah, USA.

Matthews, J. H., W. C. Funk, and C. K. Ghalambor. 2013. Demographic approaches to assessing climate change impact: An application to pond-breeding frogs and shifting hydropatterns. Pages 58–85 in J. F. Brodie, E. Post, and D. Doak, editors. Wildlife conservation in a changing climate. University of Chicago Press, Chicago, Illinois, USA.

McKinstry, M. C., and S. H. Anderson. 1999. Attitudes of private- and public-land managers in Wyoming, USA, toward beaver. Environmental Management 23:95–101.

McMenamin, S. K., E. A. Hadly, and C. K. Wright. 2008. Climatic change and wetland desiccation cause amphibian decline in Yellowstone National Park. Proceedings of the National Academy of Sciences 105:16,988–16,993.

Ray, A., A. Sepulveda, B. Hossack, D. Patla, D. Thoma, and R. Al- Chokhachy. Monitoring Yellowstone’s wetlands: Can long-term monitoring help us understand their future? Yellowstone Science, in press.

Ryan, M. E., W. J. Palen, M. J. Adams, and R. M. Rochefort. 2014. Amphibians in the climate vice: Loss and restoration of resilience of montane wetland ecosystems of the American West. Frontiers in Ecology and the Environment 12:232–240.

Schook, D. M., and D. J. Cooper. 2014. Climatic and hydrologic processes leading to wetland losses in Yellowstone National Park, USA. Journal of Hydrology 510:340–352.

Sergio, F., and P. Pedrini. 2007. Biodiversity gradients in the Alps: The overriding importance of elevation. Biodiversity and Conservation 16:3243–3254.

About the authors

Andrew Ray ( is an aquatic ecologist and Kristin Legg is program manager; both are with the NPS Greater Yellowstone Inventory and Monitoring Network in Bozeman, Montana. Adam Sepulveda is a biologist with the U.S. Geological Survey, Northern Rocky Mountain Science Center, in Bozeman. Blake Hossack is a research zoologist with the U.S. Geological Survey, Aldo Leopold Wilderness Research Institute, in Missoula, Montana. Debra Patla is field coordinator for the amphibian monitoring program with the Northern Rockies Conservation Cooperative in Jackson, Wyoming.

Environmental DNA: Can it improve our understanding of biodiversity on NPS lands?

By Andrew Ray, Adam Sepulveda, Blake Hossack, Debra Patla, and Kristin Legg

Traditional biodiversity monitoring approaches require large investments in field time, are based largely on visual observations, and require significant taxonomic expertise. New survey techniques using DNA collected from aquatic habitats may provide a cost-effective, repeatable approach to sampling a large number of sites for many taxonomic groups (Thomsen et al. 2012b; Bohmann et al. in press).

Environmental DNA (eDNA) monitoring enables the detection of organisms from DNA present and collected in water samples (Darling and Blum 2007; Darling and Mahon 2011). Detection of organisms can be confirmed because aquatic and semiaquatic organisms release DNA contained in sloughed, damaged, or partially decomposed tissue, gametes, and waste products into the water. In fact, recent evidence suggests that DNA survey techniques may be considerably more sensitive than traditional surveys for rare species (Jerde et al. 2011; Dejean et al. 2012; Pilliod et al. 2013a) and offer the ability to identify multiple species simultaneously (Minamoto et al. 2012; Thomsen et al. 2012b; Thomsen et al. 2012a) from individual water samples.

For these reasons, the Greater Yellowstone Inventory and Monitoring Network is partnering with university and agency scientists to begin testing whether eDNA monitoring can be integrated with ongoing amphibian monitoring in Grand Teton and Yellowstone National Parks. Our monitoring program is uniquely suited to evaluate the use of eDNA for amphibian richness monitoring across Grand Teton and Yellowstone for multiple reasons. First, visual encounter surveys are completed each year at approximately 250 long-term monitoring wetlands and will provide a means of testing the efficacy (i.e., determine if it is accurate and repeatable) of eDNA monitoring and potentially develop protocols for its incorporation into long-term monitoring. Additionally, these parks had two native species (e.g., spadefoot and northern leopard frog) that have not been detected in eight years of surveying. The ability to detect species at low densities with eDNA monitoring therefore offers greater potential for detecting these secretive, rare, or now-defunct species. Finally, our work and that of others suggest that some of the most biologically rich wetlands in the region occur at lower elevations; these same wetlands may be at risk for changes in climate. Cataloging the amphibian, mammalian, avian, and invertebrate assemblages or their use of these wetlands using eDNA techniques may help to more fully characterize the biodiversity of these threatened habitats (see Bohmann et al. in press).


Bohmann, K., A. Evans, M. T. P. Gilbert, G. R. Carvalho, S. Creer, M. Knapp, D. W. Yu, and M. de Bruyn. Environmental DNA for wildlife biology and biodiversity monitoring. Trends in Ecology and Evolution, in press.

Darling, J. A., and M. J. Blum. 2007. DNA-based methods for monitoring invasive species: A review and prospectus. Biological Invasions 9:751–765.

Darling, J. A., and A. R. Mahon. 2011. From molecules to management: Adopting DNA-based methods for monitoring biological invasions in aquatic environments. Environmental Research 111:978–988.

Dejean, T., A. Valentini, C. Miquel, P. Taberlet, E. Bellemain, and C. Miaud. 2012. Improved detection of an alien invasive species through environmental DNA barcoding: The example of the American bullfrog Lithobates catesbeianus. Journal of Applied Ecology 49:953–959.

Jerde, C. L., A. R. Mahon, W. L. Chadderton, and D. M. Lodge. 2011. Sight unseen: Detection of rare aquatic species using environmental DNA. Conservation Letters 4:150–157.

Minamoto, T., H. Yamanaka, T. Takahara, M. N. Honjo, and Z. Kawabata. 2012. Surveillance of fish species composition using environmental DNA. Limnology 13:193–197.

Pilliod, D. S., C. S. Goldberg, R. S. Arkle, and L. P. Waits. 2013a. Estimating occupancy and abundance of stream amphibians using environmental DNA from filtered water samples. Canadian Journal of Fisheries and Aquatic Sciences 70:1123–1130.

Pilliod, D. S., C. S. Goldberg, M. B. Laramie, and L. P. Waits. 2013b. Application of environmental DNA for inventory and monitoring of aquatic species. USGS Fact Sheet 2012-3146. U.S. Geological Survey, Forest and Rangeland Ecosystem Science Center, Corvallis, Oregon, USA.

Thomsen, P. F., J. Kielgast, L. L. Iversen, P. R. Møller, M. Rasmussen, and E. Willerslev. 2012a. Detection of a diverse marine fish fauna using environmental DNA from seawater samples. PLoS ONE 7(8):e41732.

Thomsen, P. F., J. Kielgast, L. L. Iversen, C. Wiuf, M. Rasmussen, M. T. P. Gilbert, L. Orlando, and E. Willerslev. 2012b. Monitoring endangered freshwater biodiversity using environmental DNA. Molecular Ecology 21:2565–2573.

About the authors

Andrew Ray ( is an aquatic ecologist and Kristin Legg is program manager; both are with the NPS Greater Yellowstone Inventory and Monitoring Network in Bozeman, Montana. Adam Sepulveda is a biologist with the U.S. Geological Survey, Northern Rocky Mountain Science Center, in Bozeman. Blake Hossack is a research zoologist with the U.S. Geological Survey, Aldo Leopold Wilderness Research Institute, in Missoula, Montana. Debra Patla is field coordinator for the amphibian monitoring program with the Northern Rockies Conservation Cooperative in Jackson, Wyoming.

Restoring biodiversity in Ireland’s national parks

By Daniel Sarr, Cameron Clotworthy, and Robbie Millar

Although national park conservation arose in the wilderness landscapes of the American West, it is rooted in a strong preservation ethos and its worldwide adoption has since brought it into many long-settled lands. Such human-dominated landscapes often contain novel albeit anthropic ecosystems, distinctive biodiversity, and compelling questions for conservation science (Palmer et al. 2004). Of particular interest are historically de graded landscapes that have lost essential elements of biodiversity in the past, but through changes in land use and ecological restoration may be recovering their natural and cultural heritage. Ireland, an ancient cultural landscape, provides fascinating examples of the roles that national parks can play in biodiversity conservation and restoration, particularly in highly modified landscapes (fig. 1).

[Map showing the national parks of the Republic of Ireland, which are marked with red stars; black triangles are major cities.]

Figure 1. The national parks of the Republic of Ireland are marked with red stars; black triangles are major cities.

Ice, wind, and famine: The rise and fall of Irish biodiversity

After its evolutionary stage was swept largely clean by Pleistocene glaciers, Ireland was colonized by a spare and mobile suite of species and peoples during a relatively brief time when the nation was connected to Great Britain at the tip of the British peninsula (Yalden 1999). With the rise of Holocene seas around 10,000 years ago, it assumed its current island form, and flight, wind, and water became the only routes for species to arrive. The patterns of biological colonization and persistence in Ireland suggest a story of postglacial founders and subsequent invaders that occasionally attained dominance and pushed ancient elements to marginal refugia (Searle et al. 2009). Ireland’s insular setting ensured that native species would be inherently vulnerable to extinction and to the needs of a growing human population. However, the Burren and Killarney National Parks in western and southwestern Ireland contain a number of such relict species.

The environmental history of the British Isles has been well chronicled, especially since the late Middle Ages (Lovegrove 2007). Habitat losses began early. In both Britain and Ireland, largely forested in the early Holocene, major shifts in the fossil forest beetle fauna suggest abrupt habitat changes, most likely deforestation, around 3,000–5,000 years ago (Whitehouse 2006). This parallels the flowering of advanced megalithic cultures on both islands, suggesting that while they built such timeless structures as Stonehenge and Newgrange, late Stone Age cultures started a long trajectory of landscape change. Com pounding habitat losses, persecution of many species in both Britain and Ireland accelerated until the beginning of the 20th century and still occurs in some regions (Lovegrove 2007). Losses of most large carnivores, such as wolves, eagles, and hawks, were complete, or nearly so, by the middle of the 20th century (Hickey 2000; O’Toole et al. 2002). In Ireland these abuses were compounded by the human tragedy of the Great Famine of the mid- 19th century, such that a ravaged fauna became a food source of last resort.

Restoring biodiversity in the national parks of Ireland

The culmination of Ireland’s environmental history was severe biological impoverishment. By the 20th century, only 0.5% of the nation’s land remained in forest, and the last vestiges of wild forests and their inhabitants were often to be found on private estates and hunting lodges. Without a wealth of public lands, the republic was forced to purchase lands incrementally for its parks or to accept land as gifts to the state. Nonetheless, progress has been impressive. In 1970, Ireland had only one national park and no other state-owned conservation areas (Craig 2001). However, by 2000, Ireland had established its current array of six national parks (fig. 1) and other designated conservation areas, constituting approximately 14% of its terrestrial and near-marine areas and putting it in the top half of European Union nations for lands conserved.

In the management of its national parks, Ireland follows the standards set forth by the International Union for the Conservation of Nature (IUCN) in 1969, which recommends that all governments agree to reserve the term “national park” to areas sharing the following characteristics:

• Where one or several ecosystems are not materially altered by human exploitation and occupation; where plant and animal species, geomorphological sites, and habitats are of special scientific, educational, and recreational interest or contain a natural landscape of great beauty.

• Where the highest competent authority of the country has taken steps to prevent or eliminate as soon as possible exploitation or occupation in the whole area and to effectively enforce the respect of ecological, geomorphological, or aesthetic features that have led to its establishment.

• Where visitors are allowed to enter, under special conditions, for inspirational, educational, cultural, and recreational purposes.

The application of this ideal is obviously problematic in a long-settled land, and although they occupy some of the most pristine areas of the country, none of the Irish parks have escaped human impact. Consequently, Ireland’s national parks have become anchors for active and passive restoration (e.g., removal of impacts such as turf cutting and allowing natural recovery, respectively) of native biodiversity. In some cases, this is because they contained remnant forests, for example the birch, oak, and pine woodlands at Glenveagh National Park (fig. 2A). In other cases, national parks have been determined to be places of stable land tenure, where hunting, grazing, and other impacts can be controlled. Nonetheless, major challenges to biodiversity restoration in the Irish parks, as elsewhere, include extirpation of foundational species, habitat change and loss, effects of native and domestic grazers, and invasive plant species.

[(A) Scots pine. Credit: NPS/Daniel Sarr

(B) Red deer. Credit: Con Moriarty

(C) Pontic rhododendron. Credit: Lorcan O’Toole

(D) Golden eagle chick. Credit: Pam Brophy]

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