Revision of the Inyo, Sequoia and Sierra National Forests Land Management Plans

Home , Kern brook lamprey, Kern River rainbow trout, Musteloidea, Prionus californicus, Quercus douglasii, Waucoba Mountain

Draft Biological Evaluation for Sensitive Wildlife, Fish and Invertebrate Species

Prepared by: Patricia A. Krueger Regional Threatened and Endangered Species Coordinator

for: the Inyo, Sequoia and Sierra National Forests

May 25, 2016

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Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Contents

Purpose ...... 1 Current management direction ...... 2 Description of Proposed Action ...... 2 Species of Conservation Concern ...... 4 Willow Flycatcher (All Three National Forests) ...... 5 Great Gray Owl (Sequoia and Sierra National Forests) ...... 5 Sierra Marten (All Three National Forests) ...... 5 Bats (All Three National Forests) ...... 6 Invasive Species ...... 6 Climate Change ...... 6 Assumptions ...... 6 Existing Environment ...... 6 Sensitive Species that are also Species of Conservation Concern ...... 7 Willow Flycatcher (Empidonax traillii) ...... 8 Sage-grouse (Centrocercus urophasianus) ...... 14 Bald Eagle (Haliaeetus leucocephalus) ...... 21 California Spotted Owl (Strix occidentalis occidentalis) ...... 27 Great Gray Owl (Strix nebulosa) ...... 40 Sierra Marten (Martes caurina) ...... 45 Sierra Nevada Red Fox (Vulpes vulpes necator) ...... 54 Townsend’s Big-eared Bat (Corynorhinus townsendii townsendii) ...... 61 Pacific Fringed-tailed Myotis (Myotis thysanodes) ...... 74 Pacific Fisher (Martes pennanti) ...... 86 Inyo Mountains Slender Salamander (Batrachoseps campi)...... 131 Relictual Slender Salamander (Batrachoseps relictus) ...... 137 Kern Canyon Slender Salamander (Batrachoseps simatus) ...... 142 Fairview Slender Salamander (Batrachoseps bramei) ...... 146 Kings River Slender Salamander (Batrachoseps regius) ...... 150 Black Toad (Anaxyrus exsul) ...... 154 Limestone Salamander (Hydromantes brunus) ...... 160 Foothill Yellow-legged Frog (Rana boylii) ...... 164 California Golden Trout (Oncoryhnchus mykiss aguabonita)...... 172 Hardhead Minnow (Mylopharodon conocephalus) ...... 176 Kern brook lamprey (Lampetra hubbsi) ...... 178 Kern River Rainbow Trout (Oncorhynchus mykiss gilberti) ...... 186 Tehachapi fritillary butterfly (Speyeria egleis tehachapina) ...... 192 Apache Silverspot Butterfly (Speyeria nokomis apacheana) ...... 195 Western Pearlshell (Margaritifera falcata) ...... 198 Owens Valley Springsnail (Pyrgulopsis owensensis) ...... 207 Wong’s Springsnail (Pyrgulopsis wongi) ...... 210 Sensitive Species not selected as Species of Conservation Concern ...... 214 Northern Goshawk (Accipiter gentilis)...... 214 Pallid Bat (Antrozous pallidus) ...... 222 Pygmy Rabbit (Brachylagus idahoensis) ...... 234 Wolverine (Gulo gulo) ...... 242 Western Pond Turtle (Actinemys marmorata) ...... 255 Panamint Alligator Lizard (Elgaria panamintina) ...... 262 California Legless Lizard (Anniella pulchra) ...... 267

i Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Contributers: ...... 272

Tables

Table 1. List of wildlife, fish and invertebrate species considered and forests with status for evaluation by national forest ...... 1 Table 2. Acres of both sage-grouse priority habitat and population management unit (PMU) on the Inyo National Forest ...... 16 Table 3. Active livestock grazing allotments within sage-grouse priority habitat on the Inyo NF 18 Table 4. Sierra Nevada red fox habitat (CWHR) associations in the winter ...... 57 Table 5. Sierra Nevada red fox habitat (CWHR) associations in the summer ...... 57 Table 6. Proportion of sites occupied in the Sequoia and Sierra National Forests ...... 88 Table 7. CWHR 2.1 high and moderate capability habitat for fisher (CWHR 2008 as modified by Davis et al. 2007 [CWHR2] and applied to southern Sierra Nevada forest types [CWHR2.1]) ...... 91 Table 8. Estimates of the southern Sierra Nevada fisher population occurring across the Sequoia and Sierra National Forests, Mountain Home State Park, tribal lands, Yosemite and Sequoia and Kings Canyon National Parks ...... 93 Table 9. Average fisher home range sizes in the Sequoia National Forest ...... 98 Table 10. Average fisher home range sizes in the Sierra National Forest ...... 98 Table 11. Natal and maternal den means for female fishers in Kings River Project area of Sierra National Forest through 2010 ...... 103 Table 12. Tree measures for natal and maternal dens with surrounding habitat for female fishers on the Sequoia National Forest from 1992-1996 ...... 104 Table 13. Mean of tree measures for natal and maternal dens with surrounding habitat for female fishers on the Sequoia National Forest from 1992-1996 ...... 104 Table 14. Tree species, size and height of live trees and snags used by fisher in the Sierra Nevada Adaptive Management Project Fisher Study Area for reproduction ...... 105 Table 15. Mean of tree size and height of live trees and snags used by fisher in the Sierra Nevada Adaptive Management Project Fisher Study Area for reproduction ...... 105 Table 16. Diameter at breast height (dbh) in inches of rest trees used by fishers on the Sequoia NF 1992 -1996. Derived from Truex et al. (1998) ...... 107 Table 17. Mean values for fisher rest trees and snags in the Sierra National Forest, 1999-2001 108 Table 18. Live resting tree species used by fishers compared to the number of available large trees (greater than 30 inches) in 2.47 acre plots surrounding them on the Sierra National Forest 1999-2001 ...... 108 Table 19. Female fisher home range CWHR forest type composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons ...... 110 Table 20. Female fisher home range CWHR size class composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons ...... 110 Table 21. Female fisher home range CWHR canopy closure composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons ...... 110

ii Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Figures

Figure 1. Map of willow flycatcher locations from the NRIS Wildlife Database, 2016 ...... 9 Figure 2. Map of bald eagle locations in the NRIS Wildlife Database, 2016 ...... 22 Figure 3. Map of California spotted owl locations from NRIS Wildlife Database, 2016 ...... 29 Figure 4. Map of great gray owl locations from the NRIS Wildlife Database, 2016 ...... 41 Figure 5. Map of Sierra marten locations from the NRIS Wildlife Database, 2016 ...... 46 Figure 6. Map of Sierra Nevada red fox locations from NRIS Wildlife Database, 2016 ...... 56 Figure 7. Map of Townsend’s big-eared bat locations from NRIS Wildlife Database, 2016 ...... 63 Figure 8. Map of Pacific fringe-tailed myotis locations ...... 76 Figure 9. Map of Pacific fisher locations from the NRIS Wildlife Database, 2016 ...... 90 Figure 10. Map of Inyo Mountains slender salamander locations...... 132 Figure 11. Map of Inyo Mountains slender salamander from NRIS databases, 2016 ...... 133 Figure 12. Map of relictual slender salamander locations from multiple sources, 2010 ...... 138 Figure 13. Map of relictual Slender Salamander locations from the NRIS databases, 2016 ...... 139 Figure 14. Map of Kern Canyon slender salamander from multiple sources, 2010 ...... 143 Figure 15. Map of Fairview slender salmander locations from multiple sources, 2010 ...... 147 Figure 16. Map of Kings River slender salamander locations from multiple sources, 2010...... 151 Figure 17. Distribution map of black toad from 2010 from multiple sources ...... 155 Figure 18. Map of black toad from NRIS Databases, 2016 ...... 156 Figure 19. Map of limestone salamander based on multiple sources, 2010 ...... 161 Figure 20. Map of foothill yellow-legged frog locations in the southern range ...... 165 Figure 21. Map of foothill yellow-legged frog locations in the central range ...... 166 Figure 22. Map of foothill yellow-legged frog in the NRIS Databases, 2016 ...... 167 Figure 23. Map of California golden trout from NRIS Aquatic Survey Database, 2016 ...... 173 Figure 24. Distribution of the Great Basin fritillary, the nominate species for Tehachapi fritillary, in the western United States ...... 193 Figure 25. Distribution of the Nokois fritillary, the nominate species for Apache silverspot butterfly, in the western United States...... 195 Figure 26. Map of 113 historical freshwater mussel sites in California ...... 199 Figure 27. Map of 105 sites recently surveyed or compiled by Howard (2010) ...... 200 Figure 28. Distribution of western pearlshell in North America 201 Figure 29. Areas where Owens Valley springsnail occurs ...... 208 Figure 30. Map of Wong’s springsnail from the NRIS Aquatic Survey Database, 2016 ...... 211 Figure 31. The distribution of Wong’s springsnail...... 212 Figure 32. Map of northern goshawk records within the NRIS Wildlife Database, 2016...... 215 Figure 33. Map of pallid bat locations from the NRIS Wildlife Database, 2016 ...... 224 Figure 34. Pygmy rabbit distribution map ...... 235 Figure 35. Pygmy rabbit occurrences on or adjacent to the Inyo National Forest ...... 236 Figure 36. Historic and current range of the wolverine ...... 243 Figure 37. Map of wolverine in the NRIS Wildlife Database, 2016 ...... 245 Figure 38. Modeled wolverine habitat in the western United States ...... 247 Figure 39. Map of western pond turtle from multiple sources, 2010 ...... 256 Figure 40. Map of Panamint alligator lizard from multiple sources, 2010 ...... 263 Figure 41. Map of Panamint alligator lizard from NRIS Databases, 2016 ...... 264 Figure 42. Map of California legless lizard for multiple sources, 2010 ...... 268

iii Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

iv Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Purpose Forest Service Manual 2672.41 specifies that a biological evaluation be prepared to determine if a project may affect any USDA Forest Service (FS) sensitive species. The purpose of this biological evaluation is to review the Forest Plans for the Inyo, Sequoia and Sierra National Forests in sufficient detail to determine to what extent the proposed action may affect any sensitive wildlife, fish and invertebrate species of record for the project area.

Table 1 lists the scientific and common name for the species considered in this document and indicates which national forest they are known to occur in and if they are determined to be a species of conservation concern or have other status for the forest plan revision. The list is sorted by the type of animal.

Table 1. List of wildlife, fish and invertebrate species considered and forests with status for evaluation by national forest Type Scientific Name Common Name Inyo NF Sequoia NF Sierra NF Birds Accipiter gentilis Northern goshawk Known Known Known Birds Centrocercus Greater Sage-grouse Known - SCC Not applicable Not applicable urophasianus Birds Empidonax traillii Willow flycatcher Known - SCC Known - SCC Known - SCC Birds Haliaeetus Bald Eagle Known - SCC Known - SCC Known- SCC leucocephalus Birds Strix nebulosa Great gray owl Not applicable Known - SCC Known - SCC Birds Strix occidentalis California spotted Not applicable Known - SCC Known - SCC occidentalis owl Mammals Antrozous pallidus Pallid bat Known Known Known Mammals Corynorhinus Townsend's western Known - SCC Known - SCC Known - SCC townsendii townsendii big-eared bat Mammals Gulo gulo Wolverine Potential Potential Potential Mammals Martes caurina Marten (combined Known – SCC Known – SCC Known – SCC Humboldt and Sierra) for M.c. sierra for M.c. sierra for M.c. sierra Mammals Pekania pennanti Fisher Not applicable Known - SCC Known - SCC Mammals Vulpes vulpes necator Sierra Nevada Red Candidate Candidate Candidate Fox Mammals Brachylagus idahoensis Pygmy rabbit Known Not applicable Not applicable Mammals Myotis thysanodes Fringe-tailed myotis Known – SCC Known – SCC Known – SCC for M.t. for M.t. for M.t. vespertinus vespertinus vespertinus Amphibians Batrachoseps campi Inyo Mountain Known - SCC Not applicable Not applicable salamander Amphibians Batrachoseps relictus Relictual slender Not applicable Known - SCC Not applicable salamander Amphibians Batrachoseps simatus Kern Canyon slender Not applicable Known - SCC Not applicable salamander Amphibians Hydromantes brunus Limestone Not applicable Not applicable Known - SCC salamander

1 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Type Scientific Name Common Name Inyo NF Sequoia NF Sierra NF Amphibians Rana boylii Foothill yellow- Not applicable Known - SCC Known - SCC legged frog Amphibians Batrachoseps bramei Fairview slender Not applicable Known - SCC Not applicable salamander Amphibians Batrachoseps regius Kings River slender Not applicable SCC – Known Known - SCC salamander on Monument area only Amphibians Anaxyrus exsul Black toad Known - SCC Not applicable Not applicable Reptiles Actinemys marmorata Western pond turtle Not applicable Known Known Reptiles Elgaria panamintina Panamint alligator Known Not applicable Not applicable lizard Reptiles Anniella pulchra California legless Not applicable Potential Not applicable lizard Fish Oncorhynchus mykiss California golden Known - SCC Known - SCC Not applicable aguabonita trout Fish Mylopharodon Hardhead Not applicable Known - SCC Known - SCC conocephalus Fish Lampetra hubbsi Kern brook lamprey Not applicable Known – on Known - SCC Monument area only Fish Oncorhynchus mykiss Kern River rainbow Not applicable Known - SCC Not applicable gilberti trout Invertebrates Speyeria egleis Tehachapi fritillary Not applicable Known - SCC Not applicable tehachapina butterfly Invertebrates Speyeria nokomis Apache silverspot Known - SCC Not applicable Not applicable apacheana Invertebrates Pyrgulopsis owensensis Owen's Valley Known - SCC Not applicable Not applicable springsnail Invertebrates Pyrgulopsis wongi Wong's springsnail Known - SCC Not applicable Not applicable Invertebrates Margaritifera falcata freshwater pearlshell Known - SCC Known - SCC Not applicable mussel

Current management direction The forests are currently being managed under the 1988 Inyo National Forest Land and Resource Management Plan, the 1988 Sequoia National Forest Land and Resource Management Plan, and the 1992 Sierra National Forest Land and Resource Management Plan, plus amendments to each of these plans, including the 2001 and 2004 Sierra Nevada Forest Plan Amendments. These plans include management prescriptions, standards and guidelines, and other plan components that apply to all Forest activities.

Description of Proposed Action The Inyo, Sequoia and Sierra National Forests encompass nearly 4.6 million acres of National Forest System (NFS) lands located at the southernmost extent of the Sierra Nevada mountain range. Every national forest managed by the Forest Service is required to have a land management plan (forest plan) that is consistent with the National Forest Management Act of 1976. The three Forests are currently managed under land management plans. The three southern Sierra Nevada national forests began efforts

2 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests to revise their forest plans in 2012 as part of a set of “early adopters” of the newly approved 2012 planning regulations (36 CFR 219, 2012). The descriptions of the alternatives are summarized below, the EIS contains a complete description of each alternative.

Four alternatives are analyzed in this environmental impact statement: the no-action alternative (alternative A), which represents the existing plans (as amended), and three “action” alternatives: the draft forest plans (alternative B; the preferred alternative) and two additional action alternatives (alternatives C and D), which respond to the needs for change and issues identified from public involvement.

The public generally agreed with the proposed action except for the following issues related to it:

1. The amount, type and location of thinning to improve ecosystem resilience may not provide adequate habitat for wildlife species that use forests with large trees and dense canopy cover; 2. The limitations on effectively treating enough areas to reduce the density of trees because of concerns for wildlife habitats will leave too much of the forest at risk of loss or unacceptable damage from wildfires or insect attacks during droughts; 3. The amount of prescribed fire and managed wildfire used to meet resource objectives may not be achievable without reducing existing fuels before treatment; 4. The amount of watershed restoration in the proposed revised plan may not keep pace with the increased stresses to aquatic and riparian systems from drought; 5. The proposed revised plan may not adequately protect areas of high aquatic species diversity; 6. Recommending additional wilderness areas in the proposed revised plan might unnecessarily prohibit and further geographically constrain management activities and uses, including tribal uses that would otherwise be allowed; 7. Increasing the amount of prescribed burning would produce more smoke that might impact human health and affect the tourism-based and resource-based economies of counties and rural communities; and 8. The amount of forest management activities and forest product outputs may not adequately contribute to sustaining local and regional industry infrastructure. Alternative C was developed to address Issues 1 and 5, emphasizing prescribed fire as a management tool, rather than mechanical thinning and harvest. Management focus would be on treating small-diameter trees using mechanical and hand treatment methods instead of removing trees across a range of tree diameters, and focuses on follow-up prescribed burning within treated areas.

Mechanical treatments in alternative C would emphasize vegetation and fuel reduction treatments in the wildland-urban intermix defense zone to minimize the threat of large high-intensity wildfires to communities and there would be less fuel reduction treatment in wildland areas. The treatments in the wildland-urban intermix defense zone would focus on implementing and maintaining a pattern and intensity of effective fuel reductions to lower the intensity of wildfires immediately adjacent to communities.

Alternative C complies with a 2014 court-ordered settlement agreement to include and analyze an alternative that is consistent with the findings and recommendations set forth in the “Southern Sierra Nevada Fisher Conservation Strategy” (Spencer et al. 2016); the “Draft Interim Recommendations for the Management of California Spotted Owl Habitat on National Forest System Lands” (USDA FS 2015); and that establishes plan components that conserve key characteristics associated with the ecological integrity for post-fire, complex early seral habitat. Alternative C adds the most critical aquatic refuges on all three

3 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests national forests. Direction for riparian conservation areas remains similar to alternative B, but would not include the exception to allow mechanical thinning within riparian conservation areas to facilitate burning there.

Alternative C includes the most area of recommended wilderness of all alternatives including many areas the public identified for consideration. It also includes the most areas that would be managed as critical aquatic refuges, including many areas the public identified for consideration. Alternative C also increases the size of the management area for the Pacific Crest National Scenic Trail to include areas that offer iconic views to better provide for the scenic values of the trail.

Alternative D was developed to address Issues 2, 3 and 8 by increasing the intensity of treatments and the area where fuels are pre-treated using mechanical methods in combination with strategic treatment locations to favor larger landscape prescribed burns. Like alternative B, some of the strategic treatments are concentrated in focus landscapes with more area treated and with a greater restoration toward the desired condition. This approach allows for prescribed burning across larger landscapes and provides more opportunity to manage wildfires to meet resource objectives. It allows removal of more trees, which helps managers to more rapidly address desired conditions to reduce stand density and drought-related stress on residual large and old trees and improve overall resilience of vegetation.

Alternative D emphasizes strategic mechanical thinning and prescribed burning treatments in the community wildfire protection zone, closest to communities, and the general wildfire protection zone, where fires can originate and have a high probability of reaching communities, to minimize the threat of large high-intensity wildfires. Alternative D also treats more area within the wildfire restoration zone increasing the potential to manage wildfires to meet resource objectives.

Alternative D also addresses Issue 6 by not recommending any additional areas for wilderness designation. Critical aquatic refuges are the same as alternative B. Direction for riparian conservation areas remains the same as alternative B.

Issue 4, concern regarding the pace and scale of watershed restoration is addressed by having the greatest amount of stewardship project opportunities related to the increased amount of mechanical fuel reduction. Issue 7, concern regarding the potential of smoke to affect local community health and economic sustainability, is addressed by increasing the amount of mechanical fuel reduction prior to prescribed burning and by increasing the opportunity to manage wildfires to meet resource objectives through the use of strategically located treatments and larger landscape prescribed burning.

Alternatives B, C and D incorporated relevant key findings and recommendations from the Southern Sierra Fisher Conservation Strategy and Interim Recommendations for California spotted owl, complex early seral habitat but to different extents. In addition, these alternatives include the following direction. Species of Conservation Concern The National Forest Management Act requires the Forest Service to “provide for diversity of plant and animal communities based on the suitability and capability of the specific land area in order to meet overall multiple-use objectives.” As such, the 2012 Planning Rule requires the Forest Service to maintain or restore ecological sustainability, integrity and diversity as the primary approach to species conservation. In addition, the rule requires plan components to provide the ecological conditions to maintain a viable population of species of conservation concern. A viable population is defined as one “that continues to persist over the long term with sufficient distribution to be resilient and adaptable to stressors and likely future environments.” Species of conservation concern are those species that are known to occur within the plan area and for which there is a substantial concern about the species’

4 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests capability to persist over the long term in the plan area. As required by the Planning Rule and agency directives, the Regional Forester has identified a list species of conservation concern for each national forest, which do not vary by alternative. These species are listed in the section “Wildlife, Fish and Plants.”

Alternatives B, C and D identify vegetation desired conditions designed to provide overall ecological integrity, including habitat for all associated species, and specifically to ensure they provide the ecological conditions necessary to maintain viable populations of species of conservation concern within the plan area. A guideline was developed to protect trees from removal that are used for nesting, denning, or roosting by at-risk species. This extends to some adjacent trees that provide necessary shade or other important habitat conditions. In addition, a guideline was developed to consider at-risk species early in the environmental planning process.

Alternatives B, C and D identify desired conditions, standards and guidelines for special habitats that represent small-scale habitat or vegetation types that support many at-risk plants and animals. These special habitats have plan direction to increase their consideration in project design and to help maintain and improve key ecological conditions that support several plant species of conservation concern.

For some species of conservation concern, species-specific plan components have been developed or carried forward from the existing plans and are the same across alternatives. These species include great gray owl, Sierra marten and bat species. Species-specific plan components related to willow flycatcher are not carried forward from the existing plans. We have also developed species-specific plan components for California spotted owl, Pacific fisher, greater sage-grouse and Yosemite toad, but these vary across the alternatives and they are described under each alternative description.

Willow Flycatcher (All Three National Forests) Species-specific plan direction for willow flycatcher is not being carried forward into plan revision. Current direction includes survey requirements and livestock grazing direction for occupied sites. However, there is no overlap of occupied sites and livestock grazing; therefore, additional species-specific plan direction is not necessary.

Great Gray Owl (Sequoia and Sierra National Forests) Species-specific plan direction for great gray owl would be similar to existing plan direction but only applies to the Sequoia and Sierra National Forests. Direction includes guidelines that designate a protected activity center, provide a limited operating period during the breeding season, and maintain herbaceous vegetation in meadows near nest sites for prey. It also includes a potential management approach that provide for follow-up surveys. Additional desired conditions and guidelines for forests and woodlands that comprise great gray owl habitat would provide for improved habitat quality and resilience. The great gray owl is not a species of conservation concern for the Inyo National Forest as they are not known to breed there.

Sierra Marten (All Three National Forests) Species-specific plan direction for Sierra marten incorporates recent mapping of combined Pacific fisher and marten core habitat and information from the “Science Synthesis and Climate Adaptation Strategy for the Sierra Nevada.” Much of marten core habitat overlaps with wilderness or inventoried roadless areas and would have limited management. Additional desired conditions and guidelines address management of core habitat to restore and maintain habitat quality and resilience to climate change. Although plan direction related to other species varies by alternative and may also affect marten habitat, alternatives B, C and D include plan direction to incorporate Sierra marten core habitat and conserving the key habitat characteristics.

5 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Bats (All Three National Forests) Species-specific plan direction for bats includes a standard that provides for the installation of bat gates at mines or caves with known breeding or hibernation. Invasive Species The direction for invasive species is primarily focused on noxious weeds in alternative A. For alternatives B, C and D, the direction was updated and expanded to recognize the threats to ecosystem resilience from all non-native invasive aquatic and terrestrial plants and animals likely to cause harm to ecosystems. There is an emphasis on managing invasive species by including objectives that increases the amount of area with non-native invasive plants treated. Climate Change The Forest Service is addressing climate change in a variety of ways from reducing the impacts of the operations of facilities to encouraging reduced impacts from permitted activities. The desired conditions for alternatives B, C and D include adaptations for climate change where relevant. An example is the climate adaptation strategy of providing for habitat connectivity to allow animals to move across the forests easier. Although they are not specifically addressed in the current forest plans, many of these strategies can and are considered during ongoing project development. Assumptions  Law, policy and regulations will be followed when planning or implementing site-specific projects and activities.  Plan components will be followed when planning or implementing site-specific projects and activities.  The planning timeframe for the effects analysis is 10 to 15 years.  Monitoring identified in the plan monitoring program and any broader-scale monitoring will occur and the land management plan will be amended, as needed during the life of the plan.  Funding will be available to implement restoration measures proposed.

Existing Environment The diverse landscapes of the Inyo, Sequoia and Sierra National Forests provide a rich array of ecosystems and habitat types that support thousands of wildlife, fish and plant species. These diverse landscapes include both the east side and west side of the Sierra Nevada, as well as elevations extending from approximately 1,000 feet to 14,494 feet above mean sea level. They include a variety of topography, geology and soils, and are influenced by a wide range of precipitation and temperature regimes. This diversity is also reflected by six major biological provinces present within these three national forests: Sierra Nevada Mountains, San Joaquin Valley, Great Basin Desert, Mohave Desert, Tehachapi Mountains, Great Basin and the Mojave Desert (Long et al. 2014).

The Inyo, Sequoia and Sierra National Forest contain all or portions of 272 watersheds ranging in size from 10,000 to 40,000 acres (see the “Water Quality, Water Quantity and Watershed Condition” section on page 285 [of the draft EIS, volume 1]). All of these watersheds drain into the San Joaquin Valley or terminal Great Basin lakes (Mono Lake, Owens Lake Playa). There is an estimated 1,640 miles of permanent streams are on the Inyo National Forest, an estimated 2,000 miles of permanent streams and

6 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests rivers on the Sierra National Forest, and an estimated 1,280 miles of permanent streams on the Sequoia National Forest. This diversity of habitats supports the following species diversity. Sensitive Species that are also Species of Conservation Concern Forest plans are developed to guide the maintenance or restoration of structure, function, composition, and connectivity of ecosystems to provide ecological conditions that will maintain a diversity of plant and animal communities and support the persistence of most native species in the plan area. This analysis focuses on evaluating the consequences of the plan alternatives on at-risk species.

Forest Service at-risk species include two categories: (1) federally designated species and habitat (species listed as threatened or endangered, species that are proposed or candidates for federal listing, and species with designated critical habitat on the national forests), and (2) Forest Service-designated species of conservation concern. In contrast to categories described above that are derived under the Endangered Species Act, species of conservation concern is a new category developed and used by the Forest Service under the 2012 Planning Rule to describe animal and plant species that are known to occur in the plan area and for which the Regional Forester has determined that the best available scientific information indicates substantial concern about the species' capability to persist over the long-term in the plan area.1 The species of conservation concern list guides forest planning; however, the designation of these species is not a forest plan decision. Just as there is a process for U.S Fish and Wildlife Service to change the federal listing status of a species; the Regional Forester has authority to change species of conservation concern lists to reflect new information.2 The Forest Service “sensitive species” concept is not carried forward as part of the 2012 Planning Rule and is therefore not used in these plans.

The basis for the analysis requires a determination of whether plan components such as desired conditions, objectives, standards, and guidelines provide direction to provide the ecological conditions necessary to contribute to the recovery of federally recognized species and maintain the persistence of species of conservation concern within the plan area. Plan components were developed in an iterative way, which included identifying desired conditions and potential threats to species, and identifying whether proposed plan components are sufficient to address species and their habitat needs (Forest Service Handbook 1909.12 12.52.c-d). It is also recognized that due to circumstances that are neither within the authority of the Forest Service nor consistent within the inherent capability of the land, the plan area may be unable to provide the ecological conditions necessary to maintain a viable population of a particular species of conservation concern. When this occurs, the draft environmental impact statement documents this and where possible, focuses on other efforts that are within the capability and authority of the Forest Service.

The following section provides a brief species account and summary analysis for current Regional Forester Sensitive Species that are also determined to be species of conservation concern for the Inyo, Sequoia and Sierra National Forest plan revision.

1 36 CFR 219.9 2 See Forest Service Handbook 1909.12 chapter 20, section 21.22b

7 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Willow Flycatcher (Empidonax traillii) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account This neotropical migrant species breeds within the contiguous United States, except the southeast, and the southern margins of Canada (Green et al. 2003) and winters from Mexico to northern South America (USDA 2001). Three subspecies occur in California: E. t. extimus (southern California), E. t. brewsteri (north of Fresno County from the Pacific coast to the western slopes of the Sierra Nevada crest), and E. t. adastus (on the eastern slopes of the Sierra Nevada and Cascade ranges) (summarized in USDA 2000 and Greene et al. 2003). The latter subspecies, E. t. adastus, occurs and breeds from May through September (Ibid.) and winters from the Mexican state of Colima to northwestern Venezuela (Unitt 1999 in USDA 2001).

Historically, this species likely occurred in suitable habitats throughout California (Grinnell and Miller 1944) and portions of Nevada including the central coast, Central Valley, Sierra Nevada and Great Basin (summarized in USDA 2001). Willow flycatchers were common in the Sierra Nevada until as recently as 1910 and locally abundant through 1940 (Ibid.). However, this species has declined precipitously in the Sierra Nevada since 1950 (summarized in Green et al. 2003). Urbanization and the draining, channelization, and filling of wetlands, grazing, mining, and pesticide-use are likely responsible for the decline in range and abundance of this species.

In the past three decades, willow flycatchers have undergone substantial population declines in California. Multiple factors likely contributed to the decline including poor quality of meadow habitat, shortened breeding-season length and stochastic weather events, the initial small population size, and low reproduction that influenced dispersal dynamics (Mathewson et al. 2011). Nest predation was the primary cause of nest failure at their study sites. The authors recommend two types of restoration, including: (1) restore meadows currently occupied by willow flycatchers and (2) restore meadows within 5 miles of occupied sites to provide habitat for dispersing flycatchers. Mathewson et al. (2011) suggest that restoration could enhance nest success and recommend increasing riparian shrub cover (e.g., willow) and improving meadow wetness to both increase vegetation and reduce predation rates on nests, fledglings and adults.

Willow flycatchers currently occur and breed in areas where they were thought to have “all but disappeared” (USDA 2001), though at very low densities and with limited reproductive success. The recent extirpation of this species from Yosemite National Park, where suitable habitats are presumably better preserved than those located outside the park suggests that other factors may be contributing to the decline of this species in the Sierra Nevada (Siegel et al. 2008). Siegel et al. (Ibid.) tentatively suggested that severe habitat degradation during the 19th century (due to grazing, which was discontinued in Yosemite National Park decades ago), meadow desiccation (due to global warming and resulting in earlier spring melts and a reduction in site wetness), disrupted meta-population dynamics, or conditions on the wintering grounds or along migration routes may explain the decline in Yosemite National Park.

In the NRIS database, the Inyo NF has 86 records, Sequoia NF has 6 records, and the Sierra NF has 421 records. Figure 1 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

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Figure 1. Map of willow flycatcher locations from the NRIS Wildlife Database, 2016

Habitat Status Suitable habitat (i.e. the combination of resources and environmental conditions required to survive and reproduce) for this species in the Sierra Nevada is defined by site elevation, shrub coverage, foliar density, wetness and meadow size (summarized in Green et al. 2003). Known willow flycatcher sites range in elevation from 1,200 to 9,500 feet, though most (88 percent, 119 of 135) are located between

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4,000 and 8,000 feet (Stefani et al. 2001). Willow flycatchers are closely associated with meadows that have high water tables in the late spring and early summer, and abundant shrubby, deciduous vegetation (especially Salix spp.). Shrubs in these preferred habitats are typically 6.5 to 13 feet in height, with the lower half comprised of dense woody stems. Live foliage density within the shrub layer is moderate to high and uniform from the ground to the shrub canopy (summarized in USDA 2001). Sites are “significantly more likely to support multiple willow flycatchers, and result in successful breeding efforts, as riparian shrub cover in meadows and willow flycatcher territories increases” (Bombay 1999 as cited in USDA 2001).

Within preferred sites, “the herbaceous community is consistent with high water tables and late seral conditions” (Ibid.). Furthermore, this species prefers and is significantly more likely to occupy and defend territories that have standing water or saturated soils during the breeding season, often selecting the wettest portions within meadows (summarized in USDA 2001). Occupied meadows range in size from less than 1.0 acre to 716 acres, averaging approximately 80 acres (USDA 2001). More than 95 percent of breeding meadows are larger than 10 acres, and meadows where multiple territories have fledged young are larger than 15 acres (summarized in Green et al. 2003). This species exhibits some site fidelity; 15 percent of adult birds tarsal-banded in the Sierra Nevada in 1997 and 1998 returned in a subsequent year, compared to 31 percent at the Kern River Preserve (California), and 50 percent at Malheur National Wildlife Refuge in south-eastern Oregon (summarized in Bombay et al. 2003). Between-year site fidelity on wintering grounds in Costa Rica averaged 68 percent (Koronkiewicz et al. 2006).

The CWHR model describes high to moderate capability nesting habitats in the montane riparian vegetation type (high = 2D, 3D, 4M and 4D; moderate = 2M, 3M); high to moderate capability perching habitats in the montane riparian vegetation type (high = greater than 2P; moderate = 2P); and high capability foraging habitat (no moderate capability habitats described) in the montane riparian (all strata except 1 and 2S) and wet meadow (all strata) vegetation types for this species. Similarly, as E. t. adastus nests locally in wet meadows, high and moderate capability perching habitat will include wet meadow (high = all strata) and montane riparian (high = greater than 2P; moderate = 2P) vegetation types. High capability foraging habitat, as described in CWHR (no moderate capability habitats described), will include montane riparian (all strata except 1 and 2S) and wet meadow (all strata).

Sanders and Flett (1989) reported the average territory size for a paired male willow flycatcher as approximately 0.84 acres (range = 0.145 to 2.19) in the central Sierra Nevada. This species typically nests from June 1 to August 31 and fledges young between July 15 and August 31. Fledglings remain in territories for 2 to 3 weeks after fledging (USDA 2004). However, these dates vary due to factors such as when willow flycatchers arrive on the breeding grounds, snow pack, late spring and summer weather, nest predation and brown-headed cowbird parasitism (Green et al. 2003).

This species may attempt nesting as many as three times during a single breeding season in the Sierra Nevada (USDA 2004). Nest predation has been positively associated with edge-effects, distance of the nest to edges and isolated trees, and aspects of meadow size and wetness (Cain et al. 2003). Meadow restoration (i.e. restoring natural hydrologic regimes, mitigating erosion, and stemming forest encroachment) was suggested to reduce predation of willow flycatcher nests (Green et al. 2003). Nest parasitism by brown-headed cowbirds in the Sierra Nevada ranges from a low of 4 percent (Bombay et al. 2001) to a high of 66 percent (Whitfield and Sogge 1999). Conservation concerns begin at parasitism rates of approximately 30 percent (Mayfield 1977 and Laymon 1987 in Green et al 2003) and management actions to control cowbirds may be warranted above a 60 percent parasitism rate (USDA 2004).

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Willow flycatchers are insectivorous and known to hawk prey in flight and to aerially glean prey from foliage. Foraging occurs from perches within the territory. Average foraging flights are reported to be very short (mean=13 feet, range=up to 33 feet) (summarized in Sanders and Flett 1989).

Threats Urbanization and the draining, channelization and filling of wetlands, grazing, mining and pesticide-use are likely responsible for the decline in range and abundance of this species.

Livestock grazing, predation and human activity have all been considered threats to flycatcher nesting habitat. Poorly managed grazing can alter the hydrologic and vegetative characteristics of meadows and contribute to poor quality habitat for nest selection and increased visibility (vulnerability) of nests to predation (Brookshire et al. 2002, Auble et al. 1994, Stanley and Knopf 2002, Scott et al. 2003). Nest predation is the leading cause of nest failure in willow flycatcher nests in the Lake Tahoe Basin Management Unit (Mathewson et al. 2011). Human activity (presence of people, dogs and vehicles) has also been found to be a significant impact to land birds, surpassing that of habitat loss from development (Schlesinger et al. 2008).

Direct Effects There are no direct effects as a Forest Plan does not constitute an authorization but creates a management framework on how to proceed. Chapter 3 in the draft EIS contain specific analysis for willow flycatcher.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the biological evaluation for that plan.

Alternative B, C and D: Willow flycatchers were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Implementation of restoration of meadows and in riparian areas will provide for an opportunity for willow flycatcher populations to expand back into historically occupied areas.

Cumulative Effects The Forest Plan is a management framework on how to proceed. It does not authorize a project. Any specific project or action will have appropriate level NEPA completed including project minimization measures to provide protection to wildlife and fish species. All past actions in the plan area have contributed to the existing baseline and are described in the draft EIS. Therefore only future actions would be considered. Future actions will be addressed in the specific project level NEPA.

Determination Statement Inyo, Sequoia and Sierra NFs: Implementation of the Forest Plan may impact willow flycatcher but will not lead towards Federal listing or a loss of viability. Impacts to willow flycatcher are beneficial impacts such as restoration of meadow and riparian areas and removal of invasive species.

Literature Cited – Willow Flycatcher Auble, G. T., J. M. Friedman and M. L. Scott. 1994. Relating riparian vegetation to present and future streamflows. Ecological Applications 4: 544-554.

Bombay, H.L. 1999. Scale perspectives in habitat selection and reproductive success for willow flycatchers (Empidonax traillii) in the central Sierra Nevada, California. Masters Thesis, California State University, Sacramento. 225 pp.

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Bombay, H.L. Benson, T.M. Valentine, B.E. and R.A. Stefani. 2003. A willow flycatcher survey protocol for California. June 6, 2000. USDA Forest Service, Pacific Southwest Region, Vallejo, CA. 50 pp.

Bombay, H.L. Morrison, M.L. Taylor, D.E. and J.W. Cain. 2001. Annual report and preliminary demographic analysis for willow flycatcher monitoring in the central Sierra Nevada. White Mountain Research Station, University of California, San Diego and USDA Forest Service, Tahoe National Forest. 46 pp.

Brookshire, E. N.J., J. B. Kauffman, D. Lyten and N. Otting. 2002. Cumulative effects of wild ungulate and livestock herbivory on riparian willows, Oecologia 132: 559-566.

Cain III, J.W. Morrison, M.L. and H.L. Bombay. 2003. Predator activity and nest success of willow flycatchers and yellow warblers. Journal of Wildlife Management 67(3): 600-610.

Green, G.A. Bombay, H.L. and M.L. Morrison. 2003. Conservation assessment of the willow flycatcher in the Sierra Nevada. Foster Wheeler Environmental Corporation and the University of California. 67 pp.

Grinnell, J. and A.H. Miller. 1944. The distribution of the birds of California. Pacific Coast Avifauna No. 27. Cooper Ornithological Society, Berkeley, CA. 608pp.

Koronkiewicz, T.J. Sogge, M.K. Van Riper III, C. and E.H. Paxton. 2006. Territoriality, site fidelity and survivorship of willow flycatchers wintering in Costa Rica. Condor 108:558-570.

Laymon, S.A. 1987. Brown-headed cowbirds in California: historical perspectives and management opportunities in riparian habitats. Western Birds 18:63-70.

Mathewson, H. A., H. L. Loffland, M. L. Morrison. 2011. Demographic Analysis for Willow Flycatcher Monitoring in the Central Sierra Nevada, 1997–2010: Final Report for USDA Forest Service (May 15).

Mathewson, H. A., M. L. Morrison, H. L. Loffland, P. E. Bussard. 2012. Ecology of willow flycatchers (Empidonax trailli) in the Sierra Nevada, California: Effects of meadow characteristics and weather on demographics. Ornithological Monographs 75L1-32.

Mathewson, H.A. Loffland, H.L. Morrison, M.L. Vormwald, L. and C. Cocimano. 2007. 2007 annual report and preliminary demographic analysis for willow flycatcher monitoring in the central Sierra Nevada, in partial fulfillment of cost share agreement 06-CR-11052007 between Texas A&M University and U.S.D.A. Forest Service, Region 5. Tahoe National Forest. December 31, 2007. 59 pp.

Mayfield, H.F. 1977. Brown-headed cowbird: agent of extinction? American Birds 31:107-113.

Morrison, M.L. Bombay, H.L. Cain, J.W. and D.E. Taylor. 2000. 2000 Annual report and preliminary demographic analysis for willow flycatcher monitoring in the central Sierra Nevada in partial fulfillment of contracts RFQ-17-00-30 and RFQ-17-00-31 between California State University, Sacramento and USDA Forest Service, Tahoe National Forest. November 15, 2000.

Sanders, S.D. and M.A. Flett. 1987. Ecology of a Sierra Nevada population of willow flycatchers. Western Birds 18:37-42.

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Sanders, S.D. and M.A. Flett. 1987. Ecology of a Sierra Nevada population of willow flycatchers (Empidonax traillii), 1986-1987. California Department of Fish and Game, Wildlife Management Division, Nongame Bird and Mammal Section. 27 pp.

Sanders, S.D. and M.A. Flett. 1989. Montane riparian habitat and willow flycatchers: threats to a sensitive environment and species. General Technical Report PSW-110. U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station, Albany, California. Pages 262-266.

Schlesinger, M. D., P. N. Manley, M. Holyoak. 2008. Distinguishing stressors acting on land bird communities in an urbanizing environment. Ecology 89(8): 2302 -2314.

Scott, M., S. K. Skagen and M. F. Merigliano. 2003. Relating geomorphic change and grazing to avian communities in riparian forests. Conservation Biology 17: 284-296.

Siegel, R.B. Wilkerson, R.L. and D. DeSante. 2008. Extirpation of the willow flycatcher from Yosemite National Park. Western Birds 39:8-21.

Stanley, T. R. and F. L. Knopf. 2002. Avian responses to late-season grazing in a shrub-willow floodplain. Conservation Biology 16:225-231.

Stefani, R.A. Bombay, H.L. and T.M. Benson. 2001. Willow flycatcher. Pp. 143-195 in USDA Forest Service, Sierra Nevada Forest Plan Amendment Final Environmental Impact Statement, Volume 3, Chapter 3, Part 4.4. USDA, Forest Service, Pacific Southwest Region, Vallejo, CA.

Unitt, P. 1999. A multivariate approach to identification of the willow flycatcher and its subspecies. Unpublished report. Bureau of Reclamation, Phoenix, Arizona. In U.S. Department of Agriculture (2001) Sierra Nevada Forest Plan Amendment, Final Environmental Impact Statement. Chapter 3, Part 4. USDA Forest Service, Pacific Southwest Region, Vallejo, CA.

USDA 2001. Pacific Southwest Region. Sierra Nevada Forest Plan Amendment, Final Environmental Impact Statement. Vallejo, CA.

USDA 2004. Sierra Nevada Forest Plan Amendment Record of Decision and Final Supplemental Environmental Impact Statement. USDA Forest Service, Pacific Southwest Region. Vallejo, CA 492pp + 72 pp (ROD). January 7.

Whitfield, M.J. and M.K. Sogge. 1999. Rangewide impact of brown-headed cowbird parasitism on the southwestern willow flycatcher (Empidonax traillii extimus). Pp. 182-190 in M.L. Morrison, L.S. Hall, S.K. Robinson, S.I. Rothstein, D.C. Hahn and T.D. Rich (eds.). Research and management of the brown-headed cowbird in western landscapes. Studies in Avian Biology No. 18. Cooper Ornithological Society, Camarillo, CA.

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Sage-grouse (Centrocercus urophasianus) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species Account Sage-grouse in the Bi-State Distinct Population Segment occur throughout the eastern California and western Nevada border region. The BLM Bishop Field Office and Carson Field Office, Inyo National Forest and Humboldt-Toiyabe National Forest are the primary land managers of occupied sage-grouse habitat. Population Management Units (PMUs) are areas delineated around sub-population of Bi-State sage-grouse. Part or all of the following PMUs are contained within the Inyo National Forest: Bodie, South Mono and White Mountains.

The 2012 Bi-State Action Plan outlined the current populations for each of the PMUs within the Bi-State area. Although these population estimates are not specific for areas on the Inyo NF, the Forest does manage portions of these PMUs. There are two core sage-grouse populations, Bodie Hills (Bodie PMU) and Long Valley (South Mono PMU). These core areas account for approximately 94 percent of all strutting males counted during annual lek surveys in California (Bi-State Action Plan 2012).

The Bodie PMU is marked by four distinct population cycles:

 1989 – 1992: The trend in strutting males remained high  1993 – 2003: The trend was reversed and the average number of males decreased  2004 – 2009: The trend in strutting males remained relatively stable.  2010 – 2011: The trend increased substantially with the highest record of strutting males recorded since 1953. Population estimates have been made for the White Mountain PMU, but population trend is unknown. Due to the inaccessibility of this area to conduct ground surveys during the breeding season, aerial surveys are relied upon to cover this large area. Because of weather conditions and availability of aircraft, surveys may not be conducted every year, or in the same area year after year. This variability has led to different bird counts throughout the years. For example, in 2006 a total of 206 sage-grouse (males and females) were observed during a helicopter flight (Bi-State Action Plan 2012). In 2012 the survey effort included an intensive aerial search of the White Mountains by Nevada Department of Wildlife and California Department of Fish and Wildlife. Although two new leks were documented, only 90 birds were observed during the survey (Mortimore 2012).

In the South Mono PMU there are three breeding complexes including Long Valley, Granite Mountain and Parker. The Long Valley breeding complex includes eight trend leks and associated satellite leks along the upper Owens River drainage and the Crowley Lake basin. The Granite Mountain breeding complex includes two inactive trend leks and the Parker breeding complex includes one trend lek located in Parker Meadows (Bi-State Action Plan 2012). Lek count data collected from 1987 to 2011 indicates that the Long Valley sage-grouse population is stable to moderately increasing (Bi-State Action Plan 2012).

 1989 – 2003: The trend remained stable at around 250 birds  2004 – 2007: Peak male attendance increased  2008 – 2009: Peak male attendance declined slightly  2011: Male attendance increased

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The population of sage-grouse within the Bi-State area is considered stable and rising in some portions of the area. Due to increased survey efforts over the last two years, sage-grouse have been observed in several new locations (specifically breeding grounds), including some locations not near the core populations in the Bodie and Long Valley areas. These new occurrences are helping determine the extent of sage-grouse use within the Bi-State area. No new breeding grounds were observed on the Forest, but survey helicopter flights in the White Mountains did result in the ability to increase the area surveyed and resulted in more sage-grouse being observed than in previous years. The forest is currently working with California Department of Fish and Wildlife and Nevada Department of Wildlife in a sage-grouse collaring effort it the White Mountains that has the potential to lead to more information on breeding locations.

Most recent information (Draft 2015 Bi-State Sage-Grouse Accomplishment Report, May 2015): Average sage-grouse lek attendance within the Bi-State decreased by 15.5 percent from the previous year’s average of 25.1 males per lek in 2014 to 21.2 males per lek in 2015 (Table 1). The only PMU where an increase was detected occurred in the Desert Creek and Fales PMU, and that increase was largely attributable to lek attendance increases within the California portion of the PMU (Wheeler Flat 3 and Jackass 1 Leks). No changes were detected within either the Nevada portions of the Pine Nut or the White Mountains PMUs; however, these populations are considered so small that marginal changes in population size are difficult to detect. The stronghold populations within the Bi-State Distinct Population Segment include the Bodie and South Mono PMU’s, which were down 15.1 percent and 46.2 percent respectively from 2014.

The 15-year average male attendance rate for leks within Nevada was calculated at 21.9 males per lek and the 2015 attendance rate was 16.9 percent below that figure at 18.2 males per lek.

Within the Long Valley portion of the South Mono PMU, 15 leks were surveyed and a peak total of 195 males were observed. This represented the third year in a row in which there was declining male attendance observed within Long Valley since reaching an all-time high in 2012 of 418 males in attendance. The 2015 peak count of 195 males represented an 18.6 percent decrease from the number of grouse counted in 2014 and a 53 percent decrease from 2012. This population has exhibited a steadily increasing trend since 1965.

A total of 479 males were observed during lek counts conducted in the Bodie PMU in the spring of 2015. Eighteen leks were counted of which 14 had males in attendance. The overall attendance in 2015 represented a slight dip from the 524 males observed in 2014, which was also an all-time high for this particular PMU. This population has exhibited a sharp increasing trend since 1995.

Habitat Status The Local Area Working Group designated a Technical Advisory Committee (TAC) which includes representatives from the land management agencies, state agencies (California Department of Fish and Wildlife and Nevada Department of Wildlife) and US Geological Survey. The Technical Advisory Committee developed a priority habitat map for the Bi-State area, identifying the suitable habitat (sagebrush) sage-grouse have the potential to use as well as areas surrounding leks that should be managed to reduce disturbances to breeding habitat (TAC 2012). The Forest manages approximately 213,670 acres (20 percent) of a total of 1,075,730 acres of priority habitat. The following table outlines the number of acres within each of the PMUs located on the Forest.

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Table 2. Acres of both sage-grouse priority habitat and population management unit (PMU) on the Inyo National Forest Population Total Acres in Acres of Total Acres on Acres of Priority Management Unit PMU Priority Habitat Inyo NF Habitat Bodie 349,620 197,850 63,425 9,740 South Mono 440,460 179,510 327,860 106,400 White Mountains 1,753,875 132,080 450,960 94,620 Total 2,543,955 509,440 842,245 210,760

Threats On the Forest, stressors specific to sage-grouse habitat include changes in sagebrush communities due to increases in noxious weeds or invasive species, and reduction in habitat due to wildfires, pinyon-juniper and conifer expansion, predation by ravens and development on lands adjacent to the national forest.

Jeffery pine and pinyon-juniper expansion could be the result of a combination of factors, including fire suppression, livestock grazing and changing climate. The forest has completed several habitat improvement projects with the purpose of decreasing the amount of Jeffery pine within sage-grouse habitat.

Pinyon-juniper expansion has been observed throughout the Bi-State area (Bi-State Action Plan 2012). The White Mountains sage-grouse population may be the most affected by this change, but there is little information on the areas sage-grouse use in this mountain range to allow managers to begin treatments of pinyon expansion. The Forest is working with partners to begin a collaring effort in the White Mountains which will allow for information to be gathered on sage-grouse movements and habitat use in this area. Once that information is gathered the Forest will have a better understanding on where pinyon-juniper encroachment is hindering movement of birds or decreasing suitable habitat and where habitat improvement projects should be implemented.

Noxious and invasive weeds have been inventoried on the Forest. Cheatgrass presents the greatest threat to sagebrush ecosystems due to type conversions after wildfires or other disturbances such as development and creation of roads that remove sagebrush. The forest has experienced recent wildfires within sagebrush ecosystems that have led to some cheatgrass expansion. However, these wildfires have not led to complete type conversions or reduced the suitability of these areas for sage-grouse.

Nest predation by ravens has been documented throughout the western sage-grouse population, including the Bi-State DPS. Canover et al (2010) and Coates and Delehanty (2010) both found that there is a correlation with an increase in raven predation at nest sites where visual cover does not obscure nests. Coates and Delehanty also demonstrated that an increase in raven populations can lead to an increase in nest failure among sage-grouse (2010). Both papers report that high shrub canopy cover is needed at nest sites in order to protect nests from raven predation. There are some areas documented on the Forest that have low understory and shrub cover. This could be attributed to many different drivers and stressors such as historic livestock grazing pressures, changes in weather patterns, wildfire, or lack of disturbances leading to more decadent sagebrush stands.

Development leading to reduction of sage-grouse habitat is limited on the Forest, with most development occurring on private lands located adjacent to the Forest. Areas where development may have impacted sage-grouse use or movements include the Chiatovich Creek area on the eastside of the White Mountains in Nevada. Lower elevation sagebrush areas have been fragmented by road construction and some housing structures. The extent of this impact is unclear, but development in this area may have led to changes in winter range use by sage-grouse.

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While not identified as a top threat in the Bi-State Action Plan, livestock grazing is another stressor to sage-grouse. Livestock grazing can have negative or positive impacts on sage-grouse habitat depending on the timing and intensity of grazing (Crawford et al. 2004). For example, early season light to moderate grazing can promote forb abundance and availability in both upland and riparian habitats (Crawford et al. 2004). Heavier levels of utilization decrease herbaceous cover, and may promote invasion by undesirable species (Crawford et al. 2004). Direct impacts from livestock grazing can include disturbance during the breeding, nesting, early- and late-brood rearing seasons. Disturbances may lead to dispersal from the area or abandonment of nest sites. Direct impacts also include inadvertent trampling of sage-grouse and sage- grouse nest sites by livestock. Although nest destruction by livestock trampling is rare, the presence of livestock can cause sage-grouse to abandon their nests (Crawford et al. 2004; Call and Maser 1985). Direct impacts also include those related to vegetation structure. Grazing can remove grass or forb cover that helps conceal sage-grouse nests from predators (Hockett 2002; Beck and Mitchell 2000).

Indirect impacts to sage-grouse habitat can include changes in composition, density, and structure of vegetation and removal of brood forage and cover in meadows (Call and Maser 1985; Crawford et al. 2004). Trampling of vegetation by livestock can kill sagebrush, particularly the smaller plants (Beck and Mitchell 2000). Grazing can also move sagebrush-grass communities into lower successional stable states dominated by sagebrush with little herbaceous understory (Beck and Mitchell 2000). The reduction in herbaceous understory can reduce the understory cover and decreasing the suitability of these areas for nesting. The reduction in forbs during the spring and summer may also limit their availability for sage- grouse broods (Hockett 2002). Localized and concentrated use by livestock can reduce understory grass cover, which may impact the quality of nesting habitat the following year and may affect nesting if grazed during the late spring (Beck and Mitchell 2002).

There are 49 cattle and horse and sheep and goat allotments identified on the Inyo NF (Assessment Chapter 8 – Range). Of these, 30 occur within priority sage-grouse habitat; with 20 of those being active (open to livestock grazing). Table 3 outlines the allotment information for each of these active grazing allotments.

Since 2009, 17 of these allotments have been analyzed for re-issuance of livestock grazing permits (Crowley Lake Watershed Decision Notice 2009, White Mountains Livestock Grazing Decision Notice 2010, and Mono Basin Livestock Grazing Decision Notice 2011). Re-issuance of these permits included design features specific to reducing impacts to sage-grouse including:

 Adjustments in utilization standards in meadow systems to minimize potential effects to meadow habitats (Crowley Lake Watershed allotments and White Mountain allotments).  Delaying the on-date for livestock grazing until after nesting season in the respective nesting areas (after June 1st for the Crowley Lake Watershed allotments, after July 1st in the White Mountains allotments, and after June 15th and July 1st, respectively, for the Mono Basin allotments).  Restricting livestock grazing within two miles of lek locations.  Changes in allowable use standards and management techniques to keep livestock distributed as evenly as possible to maintain sage-grouse habitat (Mono Basin livestock grazing allotments). Amendment 6 to the 1988 Land and Resource Management Plan establishes livestock allowable use based on current vegetation and watershed conditions within key areas of grazing allotments. This information was used in the grazing decisions for the 17 allotments mentioned above. Livestock grazing allowable use and changes in livestock grazing techniques are allowing vegetation and watershed conditions to improve or be maintained in their current condition, while still allowing for sage-grouse habitat.

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Table 3. Active livestock grazing allotments within sage-grouse priority habitat on the Inyo NF Allotment Name Livestock Type Permitted Season of Use (respectively, grazing can occur at any time between these dates) Alpers Canyon Cattle and Horse 6/1 - 10/30 Antelope Cattle and Horse 6/20 – 8/15 Casa Diablo Cattle and Horse 6/15 – 9/20 Clark Canyon Cattle and Horse 5/15 – 10/15 Clover Patch Cattle and Horse 6/15 – 9/1 Crooked Creek Cattle and Horse 7/1 – 9/15 Dexter Creek Sheep and Goat 6/15 – 9/15 Davis Creek Cattle and Horse 7/1 – 9/15 Glass Mountain Cattle and Horse 6/15 – 9/1 Hot Creek Cattle and Horse 6/15 – 9/25 Indian Creek Cattle and Horse 6/15 – 9/30 (currently not being grazed) June Lake Sheep and Goat 7/1 – 8/31 Long Valley Cattle and Horse 6/1 – 9/15 Mono Mills Sheep and Goat 7/1 – 9/30 Mono Sand Flat Cattle and Horse 12/1 – 5/31 Perry Aiken Cattle and Horse 6/15 – 9/30 (currently not being grazed) Tobacco Flat Cattle and Horse 6/15 – 7/15 Turner Cattle and Horse 6/15 – 8/15 and 8/16 – 9/5 Sherwin/Deadman Sheep and Goat 7/5 - 9/30 Watterson Canyon Cattle and Horse 6/1 – 9/15 Wilfred Creek Cattle and Horse 7/1 – 9/10

Within the Bi-State Action Plan, livestock grazing is not mentioned as a top threat to sage-grouse within the Bi-State DPS due to several factors: 1) changes in management across many agencies and on private lands since the Conservation Plan in 2004 has led to improvement, restoration, or maintenance of sage- grouse habitat and 2) there has been an increase in the understanding of how livestock grazing operations and management function on both private and public lands, and 3) increased research in both the Bi-State and western-wide sage-grouse populations have led to better understandings of other drivers and stressors on the landscape, such as pinyon encroachment, predation and development, all of which have been shown to have a higher effect on sage-grouse populations and habitat use in the Bi-State area.

Wild horse use also occurs within sage-grouse priority habitat. The Montgomery Pass herd occurs in the South Mono PMU and White Mountains PMU and the White Mountain wild horse herd occurs in the White Mountains PMU. Wild horse use on the Forest is discussed in Chapter 8 – Range in the Inyo National Forest Assessment. Wild horse use is a driver and stressor to sage-grouse habitat, specifically brood-rearing meadow systems. Wild horse use can lead to changes in meadow or spring conditions, altering the suitability of these areas for sage-grouse use. Wild horse use has been documented as a risk in the White Mountains PMU (Bi-State Action Plan 2012). The extent of wild horse use and resulting impacts to meadow systems has not been fully reviewed or analyzed. Personal observations made by Nevada Department of Wildlife biologists and others have led to the understanding that meadow systems are receiving high use by wild horses and that there could be a potential loss of some meadow habitats due to this use.

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Changing climate has the potential to impact sage-grouse occupying higher elevations, such as in the White Mountains, where impacts from drier weather patterns may lead to reductions in suitable brood- rearing meadows and the continued expansion of pinyon-juniper.

The Inyo NF has implemented a sage-grouse interim policy that allows for consistent management across the forest when conducting or approving activities within sage-grouse habitat on Forest lands. This document uses design criteria that have been implemented over the past several years and have been shown to lead toward maintaining, improving or restoring sage-grouse habitat. This interim policy addresses livestock grazing, wildfire, vegetation management, and mineral and energy development. The Forest also used the best available science when developing this interim policy.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and the Inyo National Forest Plan, as amended for sagebrush and sage-grouse protection, and it allows for the continued protection of sagebrush areas.

Alternative B, C and D: Sage-grouse was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of sagebrush will occur, as a management framework for the Inyo National Forest as shown in Chapter 2 and Chapter 3. Removal of invasive species will reduce the impacts and or competition from other species.

Determination Statement Inyo NF: Implementation of the Forest Plan may impact sage-grouse but will not lead towards Federal listing or a loss of viability. Impacts to sage-grouse are beneficial impacts such as restoration of sagebrush habitat and removal of invasive species.

Sequoia and Sierra NF: Implementation of the Forest plan will have no impact to sage-grouse since they are not known to occur on these forests.

Literature Cited – Sage-grouse Beck, Jeffrey L. and Dean L. Mitchell. 2000. Influences of livestock grazing on sage grouse habitat. Published in Wildlife Society Bulletin 2000, 28: 993-1002.

Bi-State Action Plan. 2012. Bi-State Action Plan Past, Present, and Future Actions for Conservation of the Greater Sage-Grouse Bi-State Distinct Population Segment. March 15, 2012.

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Bi-State Sage-Grouse Conservation Action Plan 2015 Progress Report. Developed by the Bi-State Technical Advisory Committee. Draft May, 2016.

Bi-State Technical Advisory Committee. 2012. Bi-State Sage-Grouse Preliminary Habitat Map Briefing Document.

Call W. Mayo and Chris Maser. 1985. Wildlife Habitats in Managed Rangelands-The Great Basin of Southeastern Oregon, Sage Grouse. Pacific Northwest Forest and Range Experiment Station, U.S. forest Service. General Technical Report PNW-187. 1985. p. 17-21.

Coates, Peter S. and David J. Delehanty. 2010. Nest Predation of Greater Sage-Grouse in Relation to Microhabitat Factors and Predators. Journal of Wildlife Management 74(2):240-248.

Conover, Michael R., Jennifer S. Borgo, Rebekah E. Dritz, Jonathan B. Dinkins, and David K. Dalgren. 2010. Greater Sage-Grouse Select Nest Sites to Avoid Visual Predators but not Olfactory Predators. The Condor 112(2): 331-336.

Crawford, John A., Rick A. Olson, Neil E. West, Jeffrey C. Mosley, Michael A. Schroeder, Tom D. Whiteson, Richard F. Miller, Michael A. Gregg and Chad S. Boyd. 2004. Ecology and management of sage-grouse and sage-grouse habitat. Rangeland Ecology and Management. Volume 57, Issue 1 (January 2004) pp. 2-19.

Hockett, Glenn A. 2002. Livestock Impacts on the Herbaceous Components of Sage Grouse Habitat: A Review. Intermountain Journal of Sciences. Vol 8, No. 2, 2002. pp. 105-114.

Mortimore, Craig. 2012. Bi-State Greater Sage-Grouse Spring Lek Surveillance Program Part III: Survey Findings. Completed May 14, 2012.

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Bald Eagle (Haliaeetus leucocephalus) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account The bald eagle was listed as endangered by USFWS on March 11, 1967 (32 FR 4001) and downlisted to threatened on July 12, 1995 (60 FR 35999). The bald eagle was federally de-listed on August 8, 2007 (Federal Register Vol. 72, No. 130, pp. 37346-37372). The bald eagle is also protected under the Bald and Golden Eagle Protection Act of 1940 (16 USC 668-668d).

Bald eagles occur throughout most of North America and have undergone large population fluctuations over the past two centuries (Buehler 2000, Murphy and Knopp 2000, USDA 2001). This species occurs and winters throughout California, except in desert areas. Migratory individuals from north and northeast of the State arrive between mid-October and December and remain until March or early April.

In the NRIS database, the Inyo NF has 9 records, Sequoia NF has 88 records, and the Sierra NF has 802 records. Figure 2 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Bald eagles are considered long-lived, with the oldest living bald eagle was reported near Haines, Alaska as 28 years old (Schempf 1997). In captivity, bald eagles may live 40 or more years (USDI - Fish and Wildlife Service 1999). Bald eagles are active diurnally and yearlong. Bald eagles are primarily fish eaters; however, they are opportunistic and will utilize avian and mammalian prey and carrion if readily available, especially in the nonbreeding season (Evans 1982; Zeiner et al. 1990a). Swoops from hunting perches, or soaring flight, to pluck fish from water (Evans 1982; Zeiner et al. 1990a).

Requires large bodies of water, or free-flowing rivers with abundant fish, and adjacent snags or other perches (Zeiner et al. 1990a). Perches high in large, stoutly limbed trees, on snags or broken-topped trees, or on rocks near water (Zeiner et al. 1990a). Bald eagles engage in courtship flights consisting of the pair souring together for long periods of time at great heights (Evans 1982). Occasionally they will lock talons and somersault downward several hundred feet (Evans 1982). Breeds February through July; but may start as early as November (Zeiner et al. 1990a). Pair initiation begins in January and egg-laying occurs in early May. Clutch size is 1 to 3 eggs (Evans 1982; Zeiner et al. 1990a). Incubation is usually 34 to 36 days (Evans 1982; Zeiner et al. 1990a) and fledging occurs at 10 to 12 weeks (Evans 1982). Semi-altricial young hatch asynchronously (Zeiner et al. 1990a). Bald eagles are monogamous, and breed first at 4 to 5 years (Zeiner et al. 1990a).

Dispersal distances can be substantial; this species often disperses several hundred miles from the natal site. Females tend to disperse farther than males. Breeding home ranges vary substantially by location from 58 acres in Alaska to 5 acres in Arizona. Migration distances of up to 1,712 miles have been recorded. Fidelity to wintering grounds is strong (summarized in USDA 2001).

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Figure 2. Map of bald eagle locations in the NRIS Wildlife Database, 2016

Bald eagles require open water with juxtaposed mature trees or steep cliffs for nesting, perching, foraging, and roosting (Bent 1961 in Murphy and Knopp 2000). This species typically perches in “large, robustly limbed trees, on snags, on broken topped trees, or on rocks near water” (Peterson 1986, Laves and Romsos 2000). Bald eagles wintering in the Lake Tahoe basin have been documented to use “only

22 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests dominant trees (mostly snags) within the shorezone to perch” (Laves and Romsos 2000). Ninety six percent of the perch sites (n=23) identified by Laves and Romsos (2000) were located within 0.25 miles of a large, open body of water. Late successional Jeffrey pine vegetation was used most frequently for perching and montane chaparral the least (Ibid.). Habitat (unpubl.USFS data 1994) and perch sites (Laves and Romsos 2000) identified in the Lake Tahoe basin indicate that local bald eagles prefer late successional stands (particularly Jeffrey pine) and trees that are larger in diameter and taller than the dominant tree canopy (particularly trees greater than 40 inches diameter at breast height, greater than 98 feet tall, and dead topped trees with robust, open branch structures). Perches function as resting, preening, foraging and feeding sites for bald eagles.

Bald eagles roosts communally in winter in dense, sheltered, remote conifer stands (Zeiner et al. 1990a). Roost trees are perches where one or more bald eagles rest at night and may occur long distances from open water bodies. Roost trees are similar in structure compared to perch trees; “dominant trees that have open and robust branches, are sometimes defoliated (i.e., snags), are protected from prevailing winds, and are typically far from human development” (Anthony et al. 1982 in Murphy and Knopp 2000).

Eagle nests are characteristically large, ranging from a minimum of 3 feet in width and depth to 16 feet deep and 10 feet across; size and shape are determined partly by the supporting branches (Evans 1982). Where suitable nest trees are scarce, nests are placed on ridges, cliffs and sea stacks (Evans 1982). Nest is located 50 to 200 feet above ground, usually below tree crown (Zeiner et al. 1990a) and nests are usually located near a permanent water source (Zeiner et al. 1990a).

Nest trees are typically established in large, dominant live trees with open branches and are often located within 0.96 miles of open water (Murphy and Knopp 2000). Nest trees must be sturdy to support the large, heavy stick nests built by this species at or just below the tree canopy (Ibid.). Nests are located most frequently in stands with less than 40 percent canopy cover (Call 1978 in Murphy and Knopp 2000). Nest trees in the Lake Tahoe basin are located in close proximity to open water, often less than 656 feet, and far from developed shorelines, often greater than 1.5 miles (Murphy and Knopp 2000).

The following CWHR classes provide high capability nesting habitat for this species: Eastside Pine (5S, 5P and 5D), Sierran Mixed Conifer (5S, 5P, 5D and 6), and White Fir (5S, 5P, 5D, and 6). Moderate capability nesting habitats include Sierran Mixed Conifer (all strata in size classes 1 through 3) and White Fir (all strata in size classes 1 through 3). As bald eagles are known to use the Jeffrey Pine vegetation type for nesting in the Lake Tahoe basin, despite the CWHR model prediction that this vegetation type would normally provide low nesting capability for this species, the Jeffrey Pine vegetation type will be considered high capability (5S, 5P, and 6) and moderate capability (4S, 4P, and 4D) nesting habitat for the purposes of this analysis. Moderate and high capability nesting habitat is located within 1.0 mile of open water as described above. Within CWHR, size class 6 is only recognized for a subset of the forest vegetation types (Jeffrey Pine, Montane Riparian, Sierran Mixed Conifer, and White Fir).

The following CWHR classes provide high capability perching habitat for this species: Eastside Pine (5S, 5P, 5M, and 5D), Sierran Mixed Conifer (5S, 5P, and 5M), and White Fir (5S, 5P, and 5M). Moderate capability perching habitats include Eastside Pine (4S, 4P, and 4M), Juniper (5S, 5P, and 5M), Montane Hardwood (5S, 5P, and 5M), Montane Hardwood-Conifer (5S, 5P, and 5M), Sierran Mixed Conifer (all strata in size classes 1 through 3; and 5D and 6), and White Fir (all strata in size classes 1 through 3; and 5D and 6).

The following CWHR classes provide high capability foraging habitat for this species: Lacustrine (all strata except size class 3), Riverine (all strata except size class 3), Sierran Mixed Conifer (5S, 5P, and 5M), and White Fir (5S, 5P, and 5M). Moderate capability foraging habitats include Eastside Pine (all strata except 2D, 3D, 4D, and 5D), Fresh Emergent Wetland (all strata), Juniper (all strata except 2D, 3D,

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4D, and 5D), Montane Hardwood (all except 5D), Montane Hardwood-Conifer (all except 5D and 6), Montane Riparian (all strata except 2D, 3D, 4D, 5D, and 6), Sierran Mixed Conifer (all strata except 5S, 5P, and 5M), Wet Meadow (all strata), and White Fir (all strata except 5S, 5P, and 5M).

Threats The Recovery Plan for the Pacific Bald Eagle (USFWS 1986) states that the main threats to this species in Sierra Nevada Mountains (Zone 28) are disturbance at wintering grounds and loss of potential nesting habitat to logging or development. The Plan’s proposed management directions are maintenance of winter habitat and evaluation of potential reintroduction/expansion of ‘breeders’. The most urgent site-specific task (1.3211) identified for the Forest Service in the Sierra Nevada Mountains is to prohibit logging of known nest, perch, or winter roost trees (USFWS 1986).

Organochlorine (DDE) interferes with normal calcium metabolism, resulting in thin-shelled eggs, which cannot withstand normal incubation (Evans 982). Dierldrin, polychlorinated biphenyls (PCBs), and mercury have been linked to embryonic and early chick mortality (Evans 1982). High concentrations of dieldrin and DDT are known to result in outright deaths of bald eagles (Evans 1982).

Illegal shooting remains the greatest single known cause source of bald eagle mortality (Evans 1982). Roughly half of all recorded bald eagle deaths are a direct result of shooting (Evans 1982). Other causes of mortality include impact injuries (usually powerline or tower), electrocution, trapping injuries (eagles caught in "sight bait" sets for fur bearers), automobile or train accidents, and poisoning from contaminated coyotes or other carcasses (Evans 1982).

Territories have been abandoned after disturbance from logging, recreational developments, and other human activities near nests (Zeiner et al. 1990a). Recreational use of lakes and extensive shoreline development have reduced feeding habitat (Evans 1982). Usually does not begin nesting if human disturbance is evident (Zeiner et al. 1990a). Bald eagles are also sensitive to human/recreation disturbance. In Washington, bald eagles have been found to be adversely affected by recreation that involves both pedestrian traffic and boat use by adversely affecting feeding activity (Stalmaster and Kaiser 1998). Stalmaster and Newman (1978) found that wintering bald eagles were adversely affected by human disturbance and distribution patterns were significantly changed by human activity. Eagles were displaced in areas of high human activity and moved to areas of lower human activity. Flush distances were lower when the disturbance was on land than in the water and lower still if the eagle couldn’t see the cause of the disturbance. Knight and Knight (1984) found that bald eagles became habituated to canoes in areas where they were common.

Occasionally raccoons, bobcats, crows, and under unusual circumstances, gulls prey on eggs and small young, forcing the adults away from the nest (Evans 1982).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the Biological Assessment for that plan.

Alternative B, C, and D: Bald eagles were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. All alternatives provide for large tree retention that would provide for nesting and roosting structure. Restoration or maintenance of riparian

24 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests areas and improvement or maintenance of water quality will allow for adequate prey species. Restoration of conifer and shrub areas to reduce the risk of catastrophic wildfire will allow for continuation of large trees as well as keeping sediment and silt low within the waterways.

Determination Statements Inyo, Sequoia, and Sierra NFs: Implementation of the Forest Plan may impact bald eagle but will not lead towards Federal listing or a loss of viability. Impacts to bald eagle are beneficial impacts such as providing large trees, reducing wildfires, and protection of aquatic resources.

Literature Cited – Bald Eagle Anthony, R.G. Knight, R.L. Allen, G.T. McClelland, B.R. and J.I. Hodges. 1982. Habitat use by nesting and roosting bald eagles in the Pacific Northwest. Trans. N. American. Wildlife and Nat. Res. Conf. 47:332-342.

Bent, A.C. 1961. Life histories of North American birds of prey, part 1. Dover Publications, Inc. New York, New York. Pp. 321-349.

Buehler, D.A. 2000. Bald eagle (Haliaeetus leucocephalus), the birds of North America online (A. Poole, Ed.). Ithaca: Cornell Lab of Ornithology; retrieved from the birds of North America online: http://bna.birds.cornell.edu/bna/506

Call, M.W. 1978. Nesting habitats and survey techniques for common western raptors. US Dept. of the Inter. Bur. Of Land Manage. Sacramento. Tech. Rep. 24 pp.

Delaney, D. K., T. G. Grubb, P. Beier, L. L. Pater, and M. H. Reiser. 1999. Effects of helicopter noise on Mexican Spotted Owls. Journal of Wildlife Management 63:60–76.

Evans, David L. 1982. Status Report on Twelve Raptors. United States Department of Interior Fish and Wildlife Service. Special Scientific Report - Wildlife Number 238. Washington, D.C.

Knight, R. L. and S. K. Knight. 1984. Response of Wintering Bald Eagles to Boating Activity. Journal of Wildlife Management, 48 (3): 999-1004.

Lane, Don. 2012 (April 23). Conversation with Stephanie Coppeto, Wildlife Biologist with LTBMU, regarding visitor use of bald eagle habitat along portions of the south shore of Lake Tahoe.

Laves, K.S. and J.S. Romsos. 2000. Wintering bald eagle (Haliaeetus leucocephalus) and human recreational use on the south shore of Lake Tahoe. USDA Forest Service – Lake Tahoe Basin Management Unit, South Lake Tahoe, California. 30 pp.

Murphy, D.D. and C.M. Knopp. editors. 2000. Lake Tahoe watershed assessment: Volume I. Gen. Tech. Rep. PSW-GTR-175. Albany, CA: Pacific Southwest Research Station, Forest Service, U.S. Department of Agriculture; 736 pp.

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Peterson, A. 1986. Habitat suitability index models (breeding season). US Dept. Inter. Fish and Wildl. Serv. Biol. Rep. 82(10.126). 25 pp.

Postovit, H. R., and B. C. Postovit. 1987. Impacts and mitigation techniques. Pages 183–213 in G. Pendleton, B. A. Millsap, K. W. Cline, and D. M. Bird, editors. Raptor management techniques manual. National Wildlife Federation, Washington, D.C.

Schempf, PF. 1997. Bald Eagle Longevity Records from Southeastern Alaska. J Field Ornithology: 68 (1):150-151.

Stalmaster, M. V. and J. L. Kaiser. 1998. Effects of recreational activities on wintering bald eagles. Wildlife Monographs 137

Stalmaster, M. V. and J. R. Newman. 1978. Behavioral Responses of Wintering Bald Eagles to Human Activity. Journal of Wildlife Management, 42 (3): 506-513.

USDA 2001. Pacific Southwest Region. Sierra Nevada Forest Plan Amendment, Final Environmental Impact Statement. Vallejo, CA.

USDI Fish and Wildlife Service. 1999. Endangered and Threatened Wildlife and Plants; Proposed Rule to Remove Bald Eagle in the Lower 48 States from the List of Endangered and Threatened Wildlife. Federal Register: July 6, 1999 (Volume 64, Number 128).

Zeiner, David C., William F. Laudenslayer, Jr., Kenneth E. Mayer, and Marshall White, eds. 1990a. California's Wildlife. Volume II. Birds. California Statewide Wildlife Habitat Relationship System. Department of Fish and Game, The Resources Agency, Sacramento, California. 732 pages.

U.S. Fish and Wildlife Service (USFWS). 1986. Recovery Plan for the Pacific Bald Eagle. U.S. Fish and Wildlife Service, Portland, Oregon. 160 pp.

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California Spotted Owl (Strix occidentalis occidentalis) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account The range of the California spotted owl is divided into two major physiographic provinces, the Sierra Nevada Province and the Southern California Province, with Tehachapi Pass as the dividing line (Verner et al. 1992). The Sierra Nevada Province is comprised of the southern Cascade and Sierra Nevada ranges, while the Southern California Province is comprised of all the mountain ranges of Southern California and the Central Coast ranges at least as far north as Monterey County (Ibid.). The range of the California spotted owl was revised in 2005 based on mitochondrial deoxyribonucleic acid (mtDNA) haplotypes as follows: west slope (locally on east slope) of Sierra Nevada in California from Shasta (Pit River) and Lassen Counties south to Kern County, and mountains of central, coastal, southern, and transverse ranges of California from Monterey (south side of Carmel Valley) and Kern Counties south through San Diego County to Cuyamaca Mountains in California, and Sierra San Pedro Martir in Baja California Norte, Mexico (Gutierrez and Barrowclough 2005).

Four demographic studies of California spotted owl have been ongoing for a number of years within the Sierra Nevada: (1) Eldorado National Forest (since 1983); (2) Lassen National Forest (since 1990); (3) Sierra National Forest (since 1990); and (4) Sequoia-Kings Canyon National Park (since 1990). One of the primary objectives of the demographic studies is to monitor rate of change (lambda (λ)) in owl populations (i.e., the number of owls present in a given year divided by the number of owls present the year before). For these demographic models, a lambda of one indicates a stable population; less than one indicates the population is decreasing and greater than one indicates an increasing population. Lambda is estimated from models and is typically presented as an estimate of the rate of population change, along with the standard error or a 95 percent confidence interval. The 95 percent confidence interval represents the reliability of the estimate of lambda. Managers typically view a population as stable if the 95 percent confidence interval overlaps a lambda of one.

A meta-analysis of the data from 1990 to 2005 for the four spotted owl populations in the study areas concluded that, with the exception of the Lassen study area, owl populations were stable, with adult survival rate highest at the Sequoia-Kings Canyon study site (Blakesley et al. 2010). The 95 percent confidence limit for lambda in the Lassen study area ranged from 0.946 to 1.001 (estimated value 0.973), which barely included 1, and the analysis estimated a steady annual decline of 2 to 3 percent in the Lassen study population between 1990 and 2005 (Blakesley et al. 2010).

Recent analyses from the same four demography study areas suggest that there may be a concern for decline in spotted owls within the three National Forest demography study areas in the Sierra Nevada. A preliminary analysis conducted by the Sierra Nevada Adaptive Management Project in 2011 indicates that the owl population on the Eldorado National Forest may be declining but the 95 percent confidence interval for lambda overlaps one (1) (Gutierrez et al. 2012). Tempel and Gutiérrez (in press) conclude that data from the Eldorado Density Study Area (60 percent USFS managed land in Eldorado National Forest and 40 percent private land managed timber companies) suggest a 31 percent decline in the spotted owl population size from 1993 to 2010 but again, the 95 percent confidence interval slightly overlapped one (1) for all parameters. Using data for an 18-year study period, Conner et al. (in press) found that the different estimators for ‘realized population change’ (expressed as ‘delta’ - ratio of population size at end time to initial population size) indicated population declines of 21 to 22 percent for the Lassen study area and 11 to 16 percent for Sierra study area, and an increase of 16 to 27 percent for Sequoia-Kings Canyon study area. The annual rate of population change (lambda) also showed a declining trend. However,

27 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests similar to the analyses conducted by Tempel and Gutiérrez (in press) the confidence intervals overlapped one (1.0) for all estimators and all study areas. As stated in Conner et al. (in press) “If a population is growing (lambda greater than 1), managers cannot tell whether the growth is from internal recruitment or immigration. Likewise, if a population is declining, managers cannot determine whether the declines are due to deaths within the population or emigration. Thus, additional information on specific vital rates is necessary to understand what is driving lambda and ultimately, the mechanisms driving population dynamics.” Causation for any potential decline in occupancy is unknown.

In the NRIS database, the Inyo NF has no records, Sequoia NF has 3,223 records, and the Sierra NF has 5,385 records. Figure 3 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Across the range of this species, a broad array of habitat types such as western hemlock, mixed evergreen, mixed conifer, Douglas fir, pine-oak, ponderosa pine, western incense cedar, redwood, Douglas fir/hardwood, and conifer/hardwood are used (Gutierrez et al. 1995). In the Sierra Nevada Province, spotted owls occur in conifer, mixed conifer and hardwood, and hardwood forests (Verner et al. 1992). More specifically, spotted owls use the following five vegetation types in the Sierra Nevada: foothill riparian hardwood, ponderosa pine hardwood, mixed-conifer forest, red fir forest, and east side pine forest (USDA 2001). Mixed-conifer forest is used most frequently by this species in the Sierra Nevada: approximately 80 percent of known sites are found in mixed-conifer forest, 10 percent in red fir forest, seven percent in ponderosa pine/hardwood forest, and the remaining three percent in foothill riparian/hardwood forest and eastside pine (Ibid.).

Spotted owl home ranges, and nesting and roosting locations are strongly associated with mature coniferous forests with high tree canopy cover (at least 70 percent), multilayered canopies, and an abundance of large trees and snags (Forsman et al. 1984, Bias and Gutierrez 1992, Call et al. 1992, Verner et al. 1992, Bond et al. 2004, Chatfield 2005). Spotted owl foraging habitat consists of a broader range of vegetation types that may include younger, more open habitat (Williams et al. 2011, Roberts and North 2012, Keane 2013). Large coarse woody debris is a key habitat feature of spotted owl prey. It has been suggested that some level of landscape (forest) heterogeneity may be an important consideration for spotted owl management and can improve spotted owl conservation (Williams et al. 2011, Roberts and North 2012).

Bond et al. (2004) described spotted owl nesting habitat as typically comprised of “forested stands with large trees, moderate-to-high tree densities, high canopy cover, and structural complexity”. Structural complexity may be both horizontal and vertical. Habitats used for nesting typically have “greater than 70 percent total canopy cover (all canopy above 7 feet), except at very high elevations where canopy cover as low as 30 to 40 percent may occur (as in some red fir stands of the Sierra Nevada)” (Verner et al. 1992). Large snags and an accumulation of downed woody debris are typically present (Ibid.).

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Figure 3. Map of California spotted owl locations from NRIS Wildlife Database, 2016

Spotted owl habitat use and life history requirements may be discussed at spatial scales varying from the nest area (smallest) to the non-breeding home range (largest). The nest stand (approximately 100 acres) includes one or more forest stands, the nest tree, and possibly several roost sites. Nest stands may be occupied by breeding spotted owls from February until October, and are the focus of all movements and

29 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests activities associated with nesting. Spotted owls may have more than one nest stand within their home range, and nest stands may be used intermittently for many years. Nesting behavior is initiated in February or early March when pairs begin roosting together and calling to each other more frequently at dusk before foraging or when returning to roost before dawn (Forsman 1976, Forsman et al. 1984). Egg- laying occurs in March or April (Ibid.). The average incubation period is 30 days with a measurement error of 2 days, hatching peaks May 7 to 21 (Sierra Nevada), and fledging (i.e., defined as young leaving the nest) occurs generally when the nestlings are 34 to 36 days old (Forsman et al. 1984). The post- fledging dependency period extends through late summer; dispersal from the natal site occurs in September or October (Gutierrez et al. 1985, Miller 1989). A spotted owl ecology study on the Lassen National Forest (study area 1200 to 2100 meters) found that approximately 90 percent of juveniles fledged by July 8 (Blakesley et al. 2005b).

Investigations into the thermal ecology and ecological energetics of spotted owls (Weathers et al. 2001, Blakesley et al. 2005b) found that this species’ metabolic rate increases faster than predicted allometrically in response to thermal stress and that spotted owls have exceptionally low energy requirements compared to similar-sized non-passerine birds. There is considerable debate (Verner et al. 1992) regarding whether, or to what extent, spotted owls prefer or require the micro-habitats presumed to occur within old growth or late seral forested habitats for nesting or roosting based on species-specific thermal ecology and energetics. Several previous studies of roosting habitat use indicate that northern spotted owls move vertically and horizontally within the canopy to exploit more favorable micro-climates (Barrows and Barrows 1978, Forsman 1980, Barrows 1981, Solis 1983, and Forsman et al. 1984). Yet, Verner et al. (1992) presented evidence that California spotted owls occupy and breed in habitats with high ambient summer temperatures and at least occasionally nest or roost in full sunlight when ambient temperatures exceed 100 degrees Fahrenheit and are well above the thermoneutral zone of between 64.8 to 95.4 degrees Fahrenheit (Weathers et al. 2001).

The diet of spotted owls varies geographically (Gutierrez et al. 1995). Spotted owls in the Sierra Nevada Province prey mainly on northern flying squirrels (Glaucomys sabrinus) whereas owls in the Southern California Province prey almost exclusively on dusky-footed woodrats (Neotoma fuscipes) (Verner et al. 1992). On the Lassen National Forest, flying squirrels constituted 61 percent of the diet by mass (Blakesley et al. 2005a, 2005b). On the Eldorado National Forest the primary dietary component varies by elevation: flying squirrels in upper elevation (red fir) stands, ground squirrels and gophers in mid- elevation (Sierran Mixed Conifer) stands, and woodrats in lower elevation (conifer/oak forest) stands (Eldorado National Forest spotted owl demography crew unpublished data). Other prey species in the Sierra Nevada include “deer mice (Peromyscus maniculatus), voles (Microtus spp.), bats, amphibians, insects (which are consumed with the highest frequency but represent a much lower percentage of the diet by mass), ground and tree squirrels, chipmunks (Tamias spp.), and some species of bird” (summarized by Verner et al. 1992 and Gutierrez et al. 1995).

Across their range, spotted owls exhibit population-specific demographic relationships with local weather and regional climates (Glenn et al. 2010, Glenn et al. 2011, Peery et al. 2012). Based solely on projections of climate change (i.e., not incorporating other factors such as habitat, etc.), this population-specific variation is anticipated to result in population-specific responses to future climate scenarios, which could range from little effect to potentially significant effects. These population-specific responses could result in high vulnerability. For California spotted owls, Seamans and Gutiérrez (2007b) reported that temperature and precipitation during incubation most affected reproductive output, and conditions in winter associated with the Southern Oscillation Index most affected adult survival on the Eldorado National Forest. Weather variables explained a greater proportion of the variation in reproductive output than they did for survival. Further, these two weather variables were also included in the best models predicting annual population growth rate (Seamans and Gutiérrez 2007b). MacKenzie et al. (2012) found

30 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests that Southern Oscillation Index or other weather variables explained little variation in annual reproduction for this same population of owls. Future responses to climate change are likely to be governed by complex interactions of factors that directly affect spotted owls and their habitat, as well indirect factors that can affect habitat (e.g., insect pests, disease, increased fire risk, etc.). Carroll (2010) recommended that dynamic models that incorporate vegetation dynamics and effects of competitor species in addition to climate variables are needed for rigorous assessment of future climate change on spotted owls.

The following CWHR classes provide high capability nesting habitat for this species: Montane Hardwood and Red Fir (5D); and Montane Hardwood-Conifer, Montane Riparian, Sierran Mixed Conifer, and White Fir (5D and 6). Within CWHR, size class 6 is only recognized for a subset of the forest vegetation types (Montane Hardwood Riparian, Montane Riparian, Sierran Mixed Conifer, and White Fir). The following CWHR classes provide moderate capability nesting habitat for this species: Eastside Pine and Lodgepole Pine (5D). However, approximately 80.4 percent of the forested acres within known spotted owl nest stands (n=10, 100 acres each) in the Lake Tahoe basin are Sierran Mixed Conifer (SMC) 4M, 4D, and 5M, vegetation strata not described as high or moderate capability nesting habitat by CWHR. Why spotted owls consistently select these strata in SMC stands for nesting within the Lake Tahoe basin is not clear. SMC 4M, 4D, and 5M stands may provide the most suitable nesting habitat, relative to the habitats currently available, for this species on this edge of its range. SMC 4M, 4D, and 5M stands may be sub- optimal for nesting as spotted owls do not appear to prefer these types of stands elsewhere. Regardless, as evidenced by the recurrently successful reproduction occurring in these stands locally, it is clear that SMC 4M, 4D, and 5M stands provide at least moderate capability nesting habitat within the Lake Tahoe basin. Therefore, for the purposes of this analysis, SMC 4M, 4D, and 5M stands are considered moderate capability spotted owl nesting habitat (in addition to Eastside Pine 5D and Lodgepole Pine 5D stands as identified by CWHR).

The following CWHR classes provide high capability roosting habitat for this species: Montane Hardwood and Red Fir (5M and 5D); Montane Hardwood-Conifer, Sierran Mixed Conifer, and White Fir (5M, 5D, and 6); and Montane Riparian (5D and 6). The following CWHR types and strata provide moderate capability roosting habitat for this species: Eastside Pine and Lodgepole Pine (5M and 5D); Montane Riparian and Red Fir (4M, 4D, 5S, and 5P); and Sierran Mixed Conifer and White Fir (4M and 4D).

The following CWHR classes provide high capability foraging habitat for this species: Montane Hardwood and Red Fir (5M and 5D); Montane Hardwood-Conifer, Sierran Mixed Conifer, and White Fir (5M, 5D, and 6); and Montane Riparian (5D and 6). The following CWHR classes provide moderate capability foraging habitat for this species: Eastside Pine and Lodgepole Pine (5M and 5D); Montane Hardwood (4M and 4D); Montane Hardwood-Conifer, Red Fir, Sierran Mixed Conifer, and White Fir (4M, 4D, 5S, and 5P); and Montane Riparian (3M, 3D, 4M, 4D, 5S, 5P, and 5M).

Spotted owls are territorial, generally nonmigratory, and strongly philopatric (Blakesley et al 2005b, 2006). Zimmerman et al. (2003) investigated whether this territorial species follows an ideal despotic distribution and found a positive correlation between territory occupancy and “potential fitness” as estimated from survival and reproduction; generally supporting an ideal despotic distribution (though some noise in the data was observed). Perceptual limitations, prey dynamics, and large territory sizes were identified as potential factors affecting the ability of individuals to assess habitat quality accurately. Dispersal processes, high survival rates, and long life spans were suggested as other key factors that may prevent some individuals from selecting the highest quality sites as predicted by an ideal despotic distribution (Ibid.).

Throughout the Sierra Nevada, California spotted owl nesting habitat is protected in protected activity centers (also referred to as PACs). The protected activity center includes 300 acres of the highest quality

31 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests nesting habitat available, and the most recent nest site or activity center within a spotted owl breeding territory as described in management direction for the forest (USDA 2004b). A protected activity center size of 300 acres corresponds with the following two criteria reported by Verner et al. (1992) in the California Spotted Owl report: 1) the size of the nest stand and adjacent suitable nesting stands; and 2) the area encompassing approximately 50 percent of radio-telemetry locations within spotted owl territories on the Sierra National Forest (USDA 2001). The amount of high and moderate capability nesting, roosting, and foraging habitat within each spotted owl protected activity center varies according to what is available, given existing conditions, on the forest.

A home range core area includes its associated protected activity center is 600 acres for the Sequoia and Sierra National Forests; 1,000 acres in size on the Modoc, Inyo, Humboldt-Toiyabe, Plumas, Tahoe, Eldorado, Lake Tahoe Basin Management Unit, Stanislaus National Forests and the Almanor Ranger District on the Lassen National Forest; and 2,400 acres on the Hat Creek and Eagle lake Ranger Districts of the Lassen National Forests; and is composed of the best available contiguous habitat. Like protected activity centers, home range core area s are protected in the Sierra Nevada. The core area corresponds with 20 percent of a breeding pair home range plus one standard error. Home ranges vary substantially across the range of this subspecies. Home range sizes of California spotted owls tend to be smallest in lower elevation hardwood forests, intermediate in size in conifer forests of the central Sierra Nevada, and largest in true fir forests in the northern Sierra Nevada (Verner et al. 1992). Neal et al. (1990) reported that California spotted owl home ranges in Sierra Nevada mixed conifer forests averaged 3,400 acres, including about 460 acres in stands with 70 percent or greater canopy cover, and about 1,990 acres in stands with 40 to 69 percent canopy cover. Verner et al. (1992) generally concur with these data, indicating that Sierra National Forest owls were found to have a median home range for pairs of approximately 3,000 to 5,000 acres. However, Verner et al. (1992) cite an overall mean home range size of owl pairs during the breeding period in Sierran conifer forests of about 4,200 acres. The amount of high and moderate capability nesting, roosting, and foraging habitat within each spotted owl home range core area varies according to what is available, given existing conditions, on the forest.

Threats Potential threats and stressors to this species include high severity stand-replacing fires, expansion of barred owls (Strix varia), loss of large trees and dense canopy cover, habitat fragmentation, climate change, and disease.

Little information exists on disease prevalence in California spotted owl populations, and no information exists regarding the effects of disease on individual fitness or population viability. Blood parasite prevalence sampling for California spotted owls in the northern Sierra Nevada documented that 79 percent of individuals were positive for at least one infection, whereas 44 percent of individuals tested positive for multiple infections (West Nile virus), a mosquito-borne flavivirus that was first detected in eastern North America in 1999 and spread rapidly across the continent. West Nile virus has been demonstrated to have high acute species-specific mortality rates in many raptor species (owls, hawks, and their relatives) (Gancz et al. 2004, Marra et al. 2004). None of the 141 individual California spotted owl blood samples collected from the southern (Sierra National Forest, Sequoia-Kings Canyon National Park) or northern (Plumas and Lassen National Forests) Sierra Nevada from 2004 to 2008 have tested positive for West Nile virus antibodies, which would indicate exposure and survival (Hull et al. 2010). Adult, territorial California spotted owls have high annual survival (80 to 85 percent) that has been stable across years, and no evidence has been published from the four long-term demographic studies indicating changes in adult owl survival. Nevertheless, although no effects have been documented to date, future outbreaks of West Nile virus may pose a risk to California spotted owls.

32 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Barred owls are an increasing risk factor for California spotted owls in the Sierra Nevada. Barred owls can hybridize and also out-compete spotted owls. Barred owls were first recorded within the range of the California spotted owl in 1989 on the Tahoe National Forest. Two sparred owls (hybrids of spotted and barred owls) were reported in the Eldorado National Forest during 2003 to 2004 (Seamans et al. 2004), and one of these sparred owls is still present on the study area. Barred owls were first recorded in the southern Sierra Nevada in 2004 (Steger et al. 2006). Ongoing research has documented 73 records of barred or sparred owls in the Sierra Nevada to date, with the majority of records from the northern Sierra Nevada (Tahoe, Plumas, and Lassen National Forests). Of note, five new records of barred owls were documented in the Stanislaus and Sierra national forests in 2012, indicating further range expansion of barred owls in the southern Sierra Nevada. Barred owl numbers are likely higher than documented in the Sierra Nevada, as there have been no systematic surveys for them to date.

Years of fire suppression have led to dense forested conditions with heavy fuel loading; these conditions can reduce the quality of foraging and nesting habitat (Roberts and North 2012). For example, extremely dense stand conditions characteristic of fire suppressed forests are not typically used for spotted owl foraging (Verner et al. 1992, Irwin et al. 2007). Occupancy of nesting spotted owls in fire suppressed forests may also be negatively influenced by an increasing proportion of smaller trees (less than 23 inches in diameter) around the nest (Blakesley et al. 2005).

Dense conditions characteristic of fire suppressed forests (especially ladder fuels) can also be correlated with an increased risk of fire. In a synthesis of recent available scientific research on California spotted owls, Keane (2013) concluded that spotted owls continue to occupy landscapes that have experienced low- to moderate-severity fire as well as some mixed severity fire. However, the effects of varying fire severities on spotted owl demographics (e.g., survival, reproduction) across multiple spatial and temporal (short term versus long term) scales are not well understood and the current research presents mixed results.

High severity (catastrophic) fire is considered to be a major potential threat to the California spotted owl (USFWS 2006). Large-scale stand-replacing fires can be detrimental to spotted owls, at least in the short term, possibly because these large areas do not contain habitat features important to spotted owls (Anthony & Clark 2008). High severity fires that kill most or all of the living trees effectively reduces the availability of preferred nesting and roosting habitat (mature coniferous forests with high tree canopy cover (at least 70 percent), multilayered canopies, and an abundance of large trees and snags) that can take centuries to regrow. In southwest Oregon, Clark (2007) and Clark et al. (2011) found that annual survival rates were lower in northern spotted owls inhabiting burned areas or displaced by the wildfire as compared to owls that inhabited areas outside the burn perimeter. Clark (2007) observed that although 23 northern spotted owls used all types of fire severity, within burned areas owls strongly selected low severity or unburned areas with minimal overstory canopy mortality. In this burned landscape, owl high- use areas were characterized by lower fire severity and greater structural diversity. Clark (2007) and Clark et al. (2011) also found that post-fire salvage logging reduced owl habitat quality. Bond et al. (2009) reported that foraging may occur preferentially in high severity burned areas; the study followed seven owls in four year old burned areas and found higher than expected owl foraging in high severity burned areas. The study is limited by small sample size (7 owls), short duration (12 weeks), nonrandom selection of owls, and delay (4 years) following a wildfire. Bond et al. (2002) hypothesized that wildfires may have little short term impacts on spotted owls; the authors reported that northern, California, and Mexican spotted owl survival, site fidelity, mate fidelity, and reproductive success at 11 territories one year after fires seemed uninfluenced by the fires. Four of the territories were mapped as having experienced low-to moderate-severity fire and four experienced high severity fire that burned greater than 30 percent of the territories. Roberts et al. (2011) estimated that California spotted owls studied in Yosemite National Park had similar detection, density, and occupancy rates between randomly selected unburned sites (16) and

33 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests recently burned (less than 15 years since burn) sites (16) that had predominantly burned at low- to moderate- severity. Jenness et al. (2004) found no statistical relationship between fire with mixed severity effects and Mexican spotted owl occupancy and reproduction in Arizona and New Mexico but the authors caution that higher occupancy and reproduction in unburned sites may not have been detected as statistically significant because of small sample size, lack of information on temporal and spatial variability in owl occupancy rates, and high variability in burn extent and severity. In a comparison of owl occupancy dynamics in burned versus unburned sites in the Sierra Nevada, Lee et al. (2012) found that the probability (model mean-averaged) of colonization and local extinction did not differ substantially between burned and unburned sites and the authors concluded that fire has no significant effect on occupancy dynamics. The authors also found that owls continued to occupy sites (a distinct area in which a single or territorial owl or pair had been detected) where almost one third (32 percent) of suitable habitat had been burned at high severity. They hypothesize that there may be a critical spatial threshold (proportion of a site) above which a burn at high severity could adversely affect spotted owl occupancy. Collectively, a large number of studies of fire effects on owls suggest the presence of large trees and high overstory canopy closure are the most important pre- and postfire conditions associated with spotted owl occupancy (Roberts and North 2012). However, it is clear that additional information is needed to better understand the effects of fire intensity on spotted owls.

In the Sierra Nevada, between 1999 and 2002, 18 spotted owl protected activity centers were severely affected by wildfire and could be considered “lost” (USDA Forest Service 2004, SEIS pp. 145). From 2003 to 2008, a GIS exercise by the USFS found that 33 protected activity centers had more than 75 percent of their area burned at either high or moderate severity, and rendered unusable by spotted owl, due to 8 major wildfires on NFS lands (see Table 1 and footnotes in Yasuda Declaration on October 21, 2008 for Sierra Forest Legacy et al. vs Mark Rey, Tuolumne County Alliances for Resources and Environment et al., California Ski Industry Ass’n, and Quincy Library Group). The Moonlight fire on the Plumas National Forest burned approximately 65,000 acres (46,000 on National Forest System lands) in September 2007. Based on fire severity assessment methods and severity maps (Safford et al. 2007, Miller 2007, Miller and Thode 2007), a total of approximately 43,938 acres (National Forest and private) burned at high and moderate-high severity (Basal Area Mortality greater than 50 percent); approximately 31,682 acres of forest vegetation was burned at high and moderate-high severity on National Forest system lands (Rotta 2011). This fire resulted in the immediate long-term loss of 17 California spotted owl protected activity centers and home range core areas, as well as the removal of 96 percent of the suitable nesting habitat and 86 percent of the suitable foraging habitat within the landscape.

Fuel reduction treatments attempt to remove ladder and surface fuels to reduce the potential for stand- replacing fire. Often times these treatments are conducted using mechanical equipment or a combination of hand and mechanical treatments are conducted. Overall, there is not a lot of available information about the effects of mechanical vegetation treatments on spotted owls and habitat condition (Keane 2013). The results of simulation modeling research summarized in Keane (2013) suggests that some fuels treatments can reduce fire risk and with minimal effects on owl reproduction, and may have long-term benefits of reducing wildfire risk that outweigh short-term effects of treatments. Ultimately, the risk of not doing anything can outweigh the potential short term impacts from reducing the risk of stand-replacing fire that would essentially kill all trees. The USFWS (2006) recognized that short term impacts on California spotted owl could occur from fuel reduction projects for the greater, long-term benefit of protecting nesting habitat from being lost to a stand-replacing fire. However, the effects of fuel reduction treatments to prevent stand-replacing fires is not well understood and more on-the-ground information would be useful in an adaptive management framework. For example, Seamans and Gutierrez (2007) found that alteration of greater than 49 acres of mature forest in spotted owl territories may decrease the probability of colonization. The results from a separate opportunistic case study of fuel reduction treatments (mechanical thinning of understory trees and/or prescribed fire) on protected activity center

34 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests occupancy and owl reproduction in the Stanislaus National Forest indicates that such treatments can be compatible with owl use and reproduction as owls continued to occupy the treated protected activity centers and produce young (Rich 2007). In the Plumas National Forest, where the Moonlight fire resulted in the loss of protected activity centers, fuel reduction treatments in the Meadow Valley Project are demonstrating the effects of fuel reduction treatments on spotted owls. The technique used in the Meadow Valley project, Defensible Fuel Profile Zone shows results from this study demonstrate that although owls may incur short term impacts from fuel reduction treatments, this risk outweighs the potential consequences of losing the habitat to a stand-replacing fire like the Moonlight fire. In the Meadow Valley project area, of the seven original confirmed pairs of spotted owls, there were 3 confirmed pairs, one unconfirmed pair, and one barred owl in the project area in 2012 (Keane, pers. comm., 2013). In addition to the potential effects from fuel reduction treatments, more information is needed on the value of post- fire habitat and potential effects from alteration of this habitat. Northern spotted owls have avoided habitat treated during post fire salvage logging (Clark 2007, Clark et al. 2011).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the biological evaluation for that plan.

Alternative B, C, and D: The recommendations for the California spotted owl was utilized in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Late seral stage habitat is within each alternative and the combination of large trees as well as snags is provided for in each alternative. Risks due to uncharacteristic wildfires is ameliorated by the restoration that is proposed. Protection of protected activity centers that are inclusive of the nest and or common roost sites will allow for the unique characteristics of the site to be provided for.

Determination Statement Inyo NF: Implementation of the Forest plan will have no effect to California spotted owl since they are not known to occur on these forest. Sequoia and Sierra NFs: Implementation of the Forest Plan may affect California spotted owl but will not lead towards Federal listing or a loss of viability. Impacts to California spotted owl are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags.

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Great Gray Owl (Strix nebulosa) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account Great gray owls have a Holarctic distribution, occurring “south of the tree line in northern Yukon, northwest and central Mackenzie River basin (Lockhart River and Great Slave Lake), north Saskatchewan, Manitoba, north Ontario south through southern Yukon and interior British Columbia, north and central Alberta, Manitoba, and central Ontario” (Godfrey 1986, as summarized in USDA 2004). “In the U.S. its range includes Alaska, Washington, northern Idaho, western Montana south through the Cascade and Sierra Nevada ranges to east-central California, west-central Nevada, and northwest Wyoming” (USDA 2004). Populations throughout the range are known to be irruptive. The population in the Sierra Nevada is the southernmost in the world and is disjunct from the remainder of the range. Great gray owls are thought to occur throughout the Sierra Nevada range though local distribution may be highly variable. Core areas identified include Yosemite National Park and adjacent areas on the Sierra and Stanislaus National Forests. Nesting in Sequoia National Park likely continues, but had not been confirmed since 2001 (Sears 2006). Within the core area, this species generally occurs between 2,500 to 8,900 feet elevation (summarized in Sears 2006) though this likely varies by latitude (USDA 2004). Breeding in California occurs from 2,500 to 8,000 feet elevation (Green 1995 in USDA 2004).

In the NRIS database, the Inyo NF has no records, Sequoia NF has 28 records, and the Sierra NF has 330 records. Figure 4 shows the distribution of this species but note that many of the records are at the same location, but at a different time.

Habitat Status Great gray owls in California utilize pine and fir forests adjacent to meadows between 2,500 and 7,500 feet (Winter 1986). The two factors considered most important in determining habitat use by breeding great gray owls are availability of nest sites and availability of suitable foraging habitat such as meadows (Duncan and Hayward 1994). Foraging habitat in the Sierra Nevada is generally open meadows and grasslands in forested areas, and trees along the forest edge are used for hunting perches. Openings caused by fires or timber harvest serve as foraging habitat when the vegetation is in early successional stages (Hayward 1994, Greene 1995). Greene (1995) found that sites occupied by great gray owls had greater plant cover, vegetation height, and soil moisture than sites not occupied by owls. Canopy closure was the only variable of three variables measured (canopy closure, number of snags greater than 24 inches in diameter at breast height, and number of snags less than 24 inches in diameter at breast height) that was significantly larger in occupied sites than in unoccupied sites.

Great gray owls typically forage in meadows and early seral-stage habitats that support sufficient prey, primarily Microtus and Thomomys spp. (Sears 2006). In the Sierra Nevada, pocket gophers and voles appear to be important prey species (Winter 1982, Reid 1989). Meadows appear to be the most important hunting habitat for great gray owls, where approximately 93 percent of their prey is taken (Winter 1981).

This species is strongly associated with relatively large meadows, and the number of large meadows has not changed significantly in the last decade; therefore, habitat status for this species has not significantly changed over the last decade. The only exception to this is when a stand replacing fire removes a potential nest stand adjacent to a large meadow or meadow complex habitat. In the previous 10 years, this has occurred in two instances on the Sierra NF when the Big Creek Fire removed the nest stand adjacent to

40 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Mushroom Rock and the Snake and Cargyle2 Fire in the wilderness removed a potential nest stand next to Cargyle meadow.

Figure 4. Map of great gray owl locations from the NRIS Wildlife Database, 2016

Nesting and roosting occur in adjacent conifer forests, generally in areas where canopy cover averages greater than 40 percent (USDA 2004). Nests surveyed by Sears (2006) were located within 656 to 984

41 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests feet of associated foraging meadows and roosts were located within 33 to 328 feet of the same. Persistently occupied nests were generally associated with meadows larger than 25 acres in the Yosemite area though smaller meadows (as small as 10 acres) “supported infrequent nesting” (Ibid.). This species nests in unused hawk and raven stick nests, natural depressions in broken top snags and stumps, dwarf- mistletoe platforms, and, rarely, on the ground, rock cliffs, or haystacks (summarized in Hayward and Verner et al. 1994). Great gray owls do not build or add materials to the nest; and conspecific nests may occur in close proximity (closest distance observed between nests was 1,410 feet) (Bull and Henjum 1990). Nest sites on the Stanislaus National Forest and in Yosemite National Park were in trees larger than 30 inches diameter at breast height and in stands that averaged greater than 70 percent canopy cover. Suitability of foraging meadows depended primarily upon prey availability, meadow vegetation height and cover, and meadow soil moisture (Sears 2006).

In California, courtship starts in late February or March, eggs are laid in late March or April, incubation lasts 30 to 36 days, and fledging occurs mid-May to mid-June; however, these dates vary by latitude, elevation, and spring climate conditions (Bull and Henjum 1990, USDA 2004). Both parents typically tend the young during the post-fledging dependency period. Adults defend nests and young aggressively. Most juveniles remain near the natal site, but dispersal distances of up to 468 miles have been recorded. Nesting density varies substantially by area: 0.29 pairs per square mile in Oregon and 0.66 pairs per square mile in Manitoba (Bull and Henjum 1990); 0.73 pairs per square mile in Minnesota (Duncan 1987) (summarized in USDA 2004).

High and moderate capability nesting and roosting habitat is defined as all forest vegetation types in CWHR classes 4M, 4D, 5M, 5D, and 6 (USDA 2004). CWHR describes high capability habitats in greater detail, but does not identify moderate capability habitats. For the purposes of this analysis, high capability nesting and roosting habitats include those identified as such by CWHR (described below), and moderate capability nesting and roosting habitats include all forest vegetation types in size and density classes 4M, 4D, 5M, and 5D not considered high capability by CWHR. High capability nesting habitat includes Lodgepole Pine (5D), Sierran Mixed Conifer (5D and 6), and White Fir (5D and 6); high capability roosting habitat includes Lodgepole Pine (4M, 4D, 5M, and 5D), Red Fir (4M, 4D, 5M, and 5D), Sierran Mixed Conifer (4M, 4D, 5M, 5D, and 6), and White Fir (4M, 4D, 5M, 5D, and 6).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the biological evaluation for that plan.

Alternative B, C, and D: Habitat protection for the great grey owl was utilized in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Late seral stage habitat is within each alternative and the combination of large trees as well as snags is provided for in each alternative. Restoration and or maintenance of meadows will provide for habitat for the owls prey. Risks due to uncharacteristic wildfires is ameliorated by the restoration that is proposed.

42 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests contributed to the existing baseline and are described in the draft EIS. Therefore only future actions would be considered. Future actions will be addressed in the specific project level NEPA.

Determination statement Inyo NF: Implementation of the Forest plan will have no effect to great grey owl since they are not known to occur on this forest.

Sequoia and Sierra NFs: Implementation of the Forest Plan may affect great grey owl but will not lead towards Federal listing or a loss of viability. Impacts to great grey owl are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags and restoration of meadows to allow for prey species.

Literature Cited – Great Gray Owl Bull, E.L. and M.G. Henjum. 1990. Ecology of the great gray owl. Gen. Tech. Rep. PNW-GTR-265. Portland, OR: U.S. Department of Agriculture, Forest Service, Pacific Northwest Research Station. 39pp.

Duncan, J.R. 1987. Movement strategies, mortality, and behavior of radio-marked great gray owls in southeastern Manitoba and northern Minnesota. In: Nero, R.W. R.J. Clark, R.J. Knapton, and R.H. Hamre, eds.: Biology and conservation of northern forest owls: Symposium proceedings; 1987, February 3-7; Winnipeg, MB. Gen. Tech. Rep. RM-142. Fort Collins, CO. USDA, Forest Service, Rocky Mountain Forest and Range Exp. Station.

Duncan, J.R., and P.H. Hayward. 1994. Review of technical knowledge: Great Gray Owls in Hawyard, G.D. and J. Verner, tech editors. 1994. Flammulated, Boreal, and Great Gray Owls in the United States: a technical conservation assessment. Gen. Tech. Rep. RM-253. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station. 214 pp., 3 maps.

Godfrey, W.E. 1986. The birds of Canada (revised edition). National Museum of Natural Science, Ottowa, Canada.

Green, C. 1995. Habitat requirements of great gray owls in the central Sierra Nevada. M.S. Thesis. University of Michigan, MI. 94pp.

Greene, C. 1995. Habitat requirements of great gray owls in the central Sierra Nevada. Master thesis. Univ. of Mich. 94 pp.

Hayward 1994

Hayward, G.D. and J. Verner, editors. 1994. Flammulated, boreal, and great gray owls in the United States: a technical conservation assessment. Gen. Tech. Rep. RM-253. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station. 214pp.

Reid, L.M. 1989. Channel incision by surface runoff in grassland catchments. PhD. Dissertation, University of Washington, Seattle. 202 pp.

Sears, C.L. 2006. Assessing distribution, habitat suitability, and site occupancy of great gray owls (Strix nebulosa) in California. M.S. Thesis, University of California, Davis. 88pp.

43 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Winter, Jon. 1981. Some aspects of the ecology of the great gray owl in the Central Sierra Nevada. Supported by the USFS, Stanislaus National Forest contract #43-2276. Final Report. January.

Winter, Jon. 1982. Further Investigations on the Ecology of the Great Gray Owl in the Central Sierra Nevada. Supported by the USFS, Stanislaus National Forest contract #43-2348. Final Report. February.

Winter, J. 1986. Status, distribution and ecology of the Great Gray Owl (Strix nebulosa) in California. M.S. thesis, San Francisco State University, San Francisco, CA U.S.A.

44 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Sierra Marten (Martes caurina) Regional Forester’s Sensitive Species

Proposed Species of Conservation Concern

Species Account The Sierra marten (Martes caurina) was previously classified as American marten (Martes americana) but recent genetic and morphological evidence have led to a re-classification as Pacific marten (Martes caurina) and of the subspecies sierrae called the Sierra marten (Dawson and Cook 2012).

This species is continuously distributed in Alaska and Canada, but discontinuously distributed in the western contiguous United States, where it occurs only in mountain ranges with preferred habitats. Marten occurrence appears to be associated with protected areas (e.g. National Parks and Wildernesses) and late seral forests. Timber harvest, development, and fur-trapping (which occurred until the mid- 1950s) have adversely impacted the distribution of this species (Zielinski et al. 2005).

In California, marten occur in the southern Cascades and northern Sierra Nevada south to Tulare County. Historically, martens were understood to be well distributed throughout the Cascades and northern Sierra Nevada but recent surveys suggest that the populations are now fragmented, distribution is reduced, and suitable habitat has also been reduced and isolated in parts of the range (Zielinksi et al. 2005, Kirk and Zielinksi 2009, Spencer and Rustigian-Romsos 2012). In a study of marten in northeastern California, north of the Lake Tahoe Basin Management Unit, Kirk and Zielinksi (2009) reported that marten populations are associated with areas that contain the largest amount of reproductive habitat consisting of mature, old forest. The highest density of detections was located in the largest protected area in the study region. Moriarty et al. (2011) reported approximately 60 percent fewer detections of marten at Sagehen Experimental Forest on the Tahoe National Forest than reported in the 1980s. These results, although on a smaller spatial scale, are similar to those reported by Kirk and Zielinksi (2009). Although the cause of the decreased detections is unclear, Moriarty et al. (2011) hypothesized that this was associated with loss and fragmentation of habitat; during the period of decline 39 percent of forested areas at Sagehen Experimental Forest experienced some form of timber harvest (11 percent clear-cut or shelterwood and 28 percent salvage). Habitat and occupancy models developed by Spencer and Rustigian-Romsos indicate that habitat connectivity for marten is fragmented north of the Plumas National Forest where martens appear to be restricted to isolated or semi-isolated high elevation areas (consistent with Kirk and Zielinksi (2009) whereas south of the Plumas, habitat connectivity does not appear to be greatly limiting for martens although the authors suggest that Interstate 80 may be a significant barrier to movement. An emerging issue across marten populations in California is a highly skewed sex ratio, near 2 males to one female (Slauson pers. comm. 2013). Reproductive habitat appears to be much more limited throughout the range. Therefore, there is a greater need for maintaining existing reproductive habitat and restoring it where it has been lost (Slauson pers. comm. 2013).

In the NRIS database, the Inyo NF has 12 records, Sequoia NF has 399 records, and the Sierra NF has 398 records. Figure 5 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

45 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Figure 5. Map of Sierra marten locations from the NRIS Wildlife Database, 2016

Habitat Status In the Sierra Nevada, this species is known to inhabit high elevation (4,500 to 10,500 feet) late- successional, mature red fir and lodgepole pine forests with large, decadent live trees and snags, and complex physical structure near the ground comprised of an abundance of large dead and downed wood (Buskirk and Powell 1994 in Buskirk and Ruggiero 1994, Zielinksi 2013). In the Lake Tahoe Basin

46 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Management Unit, marten are also associated with mixed conifer stands. Martens can inhabit younger forests if important elements of the mature forest are still present, especially structures for resting and denning (Purcell et al. 2012, Zielinksi 2013). Riparian areas, especially near mature forest, are important for foraging (Zielinksi 2013). The abundant large trees and dead-wood structures associated with marten presence provide prey resources, resting structures, and escape cover (Zielinksi 2013). Rest structures typically include snags, logs, and stumps; trees and snags used for resting are often the largest available (greater than 35 inches in diameter) (Purcell et al. 2012). Rest structures vary with season such that above-ground cavities are used in summer and subnivean logs, snags, and stumps are used during the winter (Zielinski 2013). Den structures typically include arboreal cavities in live trees, snags (Gilbert et al. 1997, Raphael and Jones 1997, Bull and Heater 2000) and logs, rock crevices and red squirrel middens (Ruggiero et al. 1998). Resting and denning structures may be the most limiting resource for marten on the landscape since this species uses multiple structures within their ranges (Purcell et al. 2012).

Marten occur in regions with greater snow pack (greater than 9.2 inches) compared to fisher (less than 5.2 inches) and overlap in distribution with their larger and heavier cousin in areas of intermediate snow pack. Lower foot-loading, with its associated advantage in mobility over snow, is presumed to benefit marten in interactions between the two species. In lieu of snow pack, dense shrub cover may also provide an advantage to marten during these interactions (Zielinski et al. 2006). Complex physical structure is valuable as thermal cover, especially during winter, for denning, and as foraging habitat. Marten gain access to subnivean spaces created by coarse woody debris and other structures to forage, rest, and den. Squirrel middens (e.g. Douglas squirrel) often provide natal and maternal denning and resting sites (Buskirk and Ruggiero 1994). Marten can also benefit from an interspersion of open areas (depending on size of opening) within forested habitats, which increase foraging opportunities for mice (Clethrionomys and Microtus), chipmunks (Eutamias), pikas (Ochotona), and other small mammals (Perrine 2005).

Females appear more habitat-selective than males, presumably due to the higher energetic requirements of reproduction (Buskirk and Ruggiero 1994). Marten do not alter their home ranges seasonally; though habitat use within the home range varies. Areas of greater cover, for example, are utilized during periods of inclement weather, and areas with large structural cover are used seasonally as den sites. Breeding occurs from late June to early August, peaking in July. Active gestation, which follows a 233 to 248 day embryonic diapause, lasts 27 days (260 to 275 days gestation total). Parturition occurs in March and April. Young may be moved from the natal den to a maternal den(s) and emerge at approximately 50 days. Juveniles become independent in late summer but disperse later (summarized in Buskirk and Ruggiero 1994).

Important forest types include red fir, lodgepole pine, subalpine conifer, mixed conifer-fir, Jeffrey pine, and eastside pine (Zeiner et al. 1990 in USDA 2001). The following CWHR habitat strata are moderately to highly important for marten: 4M, 4D, 5M, 5D, and 6 (USDA 2001). CWHR (2005) identifies the following classes present in the Lake Tahoe basin as high capability denning habitat: Lodgepole Pine (4M, 4D, and 5D), Montane Hardwood Conifer (4M, 4D, 5D, and 6), Montane Riparian (5D and 6), Red Fir (4M, 4D, and 5D), Sierran Mixed Conifer (6), and Subalpine Conifer (4M, 4D, and 5D). Moderate capability denning habitat includes Aspen (4M, 4D, 5D, and 6), Jeffrey Pine (4M, 4D, and 5D), Lodgepole Pine (4P and 5P), Montane Hardwood Conifer (4P and 5P), Montane Riparian (4M and 4D), Red Fir (4P and 5P), Sierran Mixed Conifer (4M, 4P, and 5D), Subalpine Conifer (4P and 5P), and White Fir (4M, 4D, 5D, and 6).

High capability resting habitat includes Lodgepole Pine (4M, 4D, 5M, and 5D), Montane Hardwood Conifer (4M, 4D, 5M, 5D, and 6), Montane Riparian (5D, 5M, and 6), Red Fir (4M, 4D, 5M, and 5D), Sierran Mixed Conifer (6), and Subalpine Conifer (4M, 4D, 5M, and 5D). Moderate capability resting habitat includes Aspen (4M, 4D, 5M, 5D, and 6), Barren (all strata), Eastside Pine (5M, 5P, and 5D),

47 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

Jeffrey Pine (4M, 4D, 5M, and 5D), Lodgepole Pine (4P and 5P), Montane Hardwood Conifer (4P and 5P), Montane Riparian (4M and 4D), Red Fir (4P and 5P), Sierran Mixed Conifer (4M, 4P, 5M, and 5D), Subalpine Conifer (4P and 5P), and White Fir (4M, 4D, 5M, 5D, and 6).

High capability foraging habitat includes Lodgepole Pine (4M, 4D, 5M, and 5D), Montane Hardwood Conifer (4M, 4D, 5M, 5D, and 6), Montane Riparian (5D, 5M, and 6), Red Fir (4M, 4D, 5M, and 5D), Subalpine Conifer (4M, 4D, 5M, and 5D), and Wet Meadow (all strata). Moderate capability foraging habitat includes Aspen (4M, 4D, 5M, 5D, and 6), Barren (all strata), Eastside Pine (5M and 5D), Jeffrey Pine (4M, 4D, 5M, and 5D), Lodgepole Pine (3S, 3P, 4S, 4P, and 5P), Montane Hardwood Conifer (4P and 5P), Montane Riparian (4M and 4D), Pasture (all strata), Perennial Grassland (all strata), Red Fir (3S, 3P, 4S, 4P, and 5P), Sierran Mixed Conifer (4M, 4P, 5M, 5D, and 6), Subalpine Conifer (3S, 3P, 4S, 4P, and 5P), and White Fir (4M, 4D, 5M, 5D, and 6).

Threats Some of the threats facing martens include habitat loss and fragmentation, especially clear-cutting, fuel reduction treatments, and wildfire (Zielinksi 2013). Marten occupancy and geographic range is also predicted to be influenced by climate change such that the species will be highly sensitive to climate change, and would probably experience the largest climate impacts at the southernmost latitudes (i.e. in the southern Sierra Nevada) (Lawlor et al. 2011).

Marten are very sensitive to habitat loss and fragmentation and rarely occupy landscapes after greater than 30 percent of the mature forest has been harvested (Zielinksi 2013). Martens tend to avoid clear cut openings or will cross only small openings (e.g., less than 500 feet). However, openings that have some structure retained (e.g., isolated trees, snags, logs), were more likely to be crossed by marten in the Rocky Mountains, even if the openings were relatively large (maximum distance = 600 feet), than if the opening had no structures and were small (summarized in Zielinksi 2013). Females tend to be more specialized than males in their habitat needs and tend to avoid managed areas of lesser habitat value and greater predation risk (summarized in Zielinski 2013).

The effect of thinning treatments (including fuel reduction treatments) on marten in the Sierra Nevada is currently being studied. The effects can be positive and negative for marten; positive if treatments set the trajectory toward historical conditions while retaining key habitat features (e.g., snags, large and complex trees, coarse woody debris), and if unsuitable stands are treated to accelerate the recruitment of mature forest characteristics and reduce the chance of catastrophic wildfire (Slauson and Zielinksi 2008). The effects can be negative if the treated habitat increases the risk of predation by reducing canopy cover significantly, removing resting and denning structures and escape cover (e.g., tree boles), and/or reducing the complexity of the understory (clear cutting from below). Treatments effects can also be negative if habitat patches require a lot of energy and risk to travel between (increased fragmentation), if treatment has adversely effected prey resources, and if den structures are reduced or altered in a way that reduces the survival of young (Slauson and Zielinski 2008).

According to Zielinksi (2013), there is a need to understand the tradeoff between treating stands to reduce fuel loadings and loss of the stand to catastrophic wildfire. Some simulation work for fisher suggests that the indirect and immediate negative effects of treatment are justified for the long term positive effects for the prevention of large wildfires in fisher habitat that could damage and fragment habitat over larger areas (Scheller et al. 2011). For northern spotted owls, a simulation model indicated that active management of sites with high fire hazard was more favorable to spotted owl conservation over the long term (75 years) than no management (Roloff et al. 2012). Purcell et al. (2012) suggests that research findings support the validity of recommendations made in North et al. (2009) to treat habitat for marten in areas where historically, fire would have burned less frequently, such as north-facing slopes, canyon bottoms, and

48 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests riparian areas. Regardless, the type and timing of treatments as well as home range and landscape-level effects from treatments should be carefully evaluated to understand the short and long term outcomes.

Habitat connectivity is naturally limited by the spatial distribution of suitable habitats on the landscape. Green (2007) found that marten frequently occurred in continuous stands of Sierran mixed conifer, red fir, and subalpine forest dominated by large trees (size classes 5 and 6) and dense canopy cover at sites in western and southwestern parts of Sequoia and Kings Canyon National Parks. However, in the northern and eastern parts of the Park, Green found that marten in higher elevations utilized areas with smaller diameter trees (mean tree size class 4) and lower canopy cover (range from less than 10 percent to dense) that were less continuous, “often occurring in linear patches along streams or around edges of lakes” (2007). Continuity of forested habitat in high elevation areas may be ameliorated by the presence of boulder fields, talus slopes, abundant surface rocks, and/or shrubs (Buskirk and Zielinski 2003, Slauson 2003, summarized in Green 2007). These alternate sources of thermal, predator, and/or foraging cover may be important as linkages between preferred habitats and, at the landscape level, for habitat connectivity.

In addition to vegetation management, marten are also sensitive to recreation activities, particularly snow activities (e.g., ski facilities). Much of the information presented on marten and ski resorts comes directly from Zielinski (2013) and Slauson (unpublished data and in prep publications). Ski resorts are considered likely to affect marten populations because they remove and fragment high-elevation fir forest habitat. The operation of ski resorts includes the continued compaction of snow, presence of high densities of skiers, and nocturnal grooming activities. All these factors can have negative effects on marten both directly (females may avoid these areas) or indirectly (snow compaction and forest fragmentation facilitate high predation by coyotes) (Slauson et al. 2008). To create ski runs, chair lifts, and associated facilities, trees are removed, creating open areas and fragmenting forest. Skiers and staff are active during the majority of the day, and grooming and some skiing activity occur during the night. Thus, martens that are sensitive to these activities may not find time for important foraging activities. Ski resort effects are not limited to winter, as habitat fragmentation is a year-round effect and many resorts are developing summer recreational activities (e.g., hiking, mountain biking).

There are approximately 25 ski resorts in the Sierra Nevada, and nearly all occur within the range of the marten (Zielinski 2013). The Lake Tahoe region includes approximately half of these resorts, constituting the highest density of resorts in the Sierra Nevada and one of the highest in North America (Zielinski 2013). Kucera (2004) conducted the only published intensive study of martens in a ski area in California. He captured 12 individuals at the Mammoth Mountain ski area, 10 of which were males, 1 was female, and 1 was of unknown sex, resulting in a highly skewed sex ratio. The single female raised two kits, but did not use developed areas and only used natural rest sites. Martens appeared to move away from the ski area and into unmanaged forest after winter. Kucera (2004) suggested that this fits a seasonal use pattern where martens occupy ski areas during winter when natural prey is least available and human-supplied food is most plentiful, then they move into unmanaged forests in spring. This migration would allow them to exploit artificial food sources during winter, but return to places where females maintain home ranges to breed in summer.

Realizing that this study required confirmation and a larger sample, Slauson and Zielinski (unpubl. data) began a 4-year study in 2008 to evaluate the effects of ski area development and use on home range and demography of marten populations. Preliminary data are described here, and a published report is forthcoming. Similar to results from Kucera (2004), female martens appear to exhibit a higher sensitivity to forest fragmentation from ski-run creation than males, avoiding areas highly fragmented by ski runs (K. Slauson unpubl. data). These results are similar to those for clear cuts in that males may show some plasticity in their use of managed habitat but females are more selective. For example, although Slauson

49 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests et al. (in prep) have detected martens using approximately 70 percent of the Heavenly ski area during the spring, females use less than 33 percent for reproductive habitat. Furthermore, while males occupy more highly fragmented portions of ski areas than females, male survivorship appears to be lower in sites with higher fragmentation (K. Slauson unpubl. data). And like Kucera (2004), Slauson has also found a skewed sex ratio (more males to females) on ski resorts in the Lake Tahoe Basin Management Unit. Martens give birth to their young in late March and early April, typically coinciding with the end of the ski operations period in most years. As the snowpack breaks up martens shift their activity to be more active during the daytime to focus their foraging activities on diurnally active species, such as chipmunks and golden-mantled ground-squirrels (Slauson et al. in prep). Therefore, there is a greater potential for human-marten interactions during the late spring and early summer when both humans and martens are active during the daytime (Slauson et al. in prep) and this potential for interaction may increase with new national policy allowing ski facilities to host more summer activities.

Other snow activities may affect marten but data from the Lake Tahoe Basin Management Unit indicate that off-highway vehicles and over-snow vehicles use did not affect marten occupancy or probability of detection and that overall off-highway vehicles and over-snow vehicles use in the study areas was low (1 off-highway vehicles and over-snow vehicles pass every 2 hours) and exposure occurred in less than 20 percent of a typical home range (Zielinski et al. 2007).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Marten and its habitat needs were considered in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Late seral stage habitat is within each alternative and the combination of large trees as well as snags is provided for in each alternative. Corridors and landscape linkages are key development to allow for the species to move from one area to another, hence reducing the impacts of climate change. Risks due to uncharacteristic wildfires is ameliorated by the restoration that is proposed.

Determination Statement Inyo, Sequoia and Sierra NFs: Implementation of the Forest Plan may affect marten but will not lead towards Federal listing or a loss of viability. Impacts to marten are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags.

Literature Cited - Marten Bull and Heater 2000

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Buskirk, S.W. and L.F. Ruggiero. 1994. Pacific marten. In Ruggiero, L.F. Aubry, K.B. Buskirk, S.W. Lyon, L.J. and W.J. Zielinski, editors. The scientific basis for conserving forest carnivores: Pacific marten, fisher, lynx, and wolverine in the western United States. General Technical Report RM- 254. U.S. Department of Agriculture, Forest Service, Rocky Mountain Forest and Range Experiment Station, Fort Collins, Colorado.

Buskirk, S.W. and R.A. Powell. 1994. Habitat ecology of fishers and Pacific martens. Pages 283–296 in S.W. Buskirk, A.S. Harestad, M.G. Raphael, and R.A. Powell, editors. Martens, sables and fishers: biology and conservation. Cornell University Press, Ithaca, New York, USA.

Buskirk, S.W. and W.J. Zielinski. 2003. Small and mid-sized carnivores. Pages 207-249 in Zabel, C.J. and R.G Anthony, editors. Mammal community dynamics: Management and conservation in coniferous forests of western North America. Cambridge University Press, Cambridge, United Kingdom.

Dawson, N. G. and J. A. Cook. 2012. Behind the genes: Diversification of North American Martens (Martes americana and M. cuarina) pp 23-38 in Biology and Conservation of Martens, Sables, and Fishers: A New Synthesis (Ed. Aubry, K. B., Zielinksi, W. J., Raphael, M. G., Proulx, G., Buskirk, S. W.), Comstock Publishing Associates, Cornell University Press, Ithaca, NY.

Gilbert et al. 1997,

Green, R.E. 2007. Distribution and habitat associations of forest carnivores and an evaluation of the California Wildlife Habitat Relationships model for Pacific marten in Sequoia and Kings Canyon National Parks. Masters Degree Thesis, Humboldt State University. Arcata, CA. 103 pp.

Hall, E.R. 1981. The mammals of North America. 2d ed. New York: John Wiley & Sons. 1,181 pp.

Kirk, T. A. and W. J. Zielinski. 2009. Developing and testing a landscape habitat suitability model for the American marten (Martes americana) in the Cascades Mountains of California. Landscape Ecology 24 (6): 759-773.

Kucera, T. 2004. Ecology of American martens on the Mammoth Mountain Ski area, Inyo National Forest, California. Inyo National Forest, Bishop, California.

Lawler, J. J., H.D. Safford, and E. H. Girvetz. 2011. Martens and fishers in a changing climate in K.B. Aubry (ed). Biology and Conservation of Martens, Sables, and Fishers: A New Synthesis. Cornell University Press, Ithaca, NY. In Press

Moriarty et al. (2011)

North, M., P. Stine, K. O’Hara, W. Zielinksi, S. Stephens. 2009. An ecosystem management strategy for Sierran mixed-conifer forests, 2nd printing with addendum. Gen Tech. Rep. PSW-GTR-220. Albany, CA: US Department of Agriculture, Forest Service, Pacific Southwest Research Station, 49 p.

Perrine 2005).

Purcell, K. L., C. M. Thompson, and W. J. Zielinski. 2012. Fishers and American Martens in PSW-GTR- 237 Managing Sierra Nevada Forests, USDA Forest Service, Pacific Southwest Research Station.

Raphael and Jones 1997,

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Roloff et al. 2012

Scheller, R. M., W. D. Spencer, H. Rustigian-Romsos, A. D. Syphard, B. R. Ward, J. R. Strittholt. 2011. Using stochastic simulation to evaluate competing risks of wildfires and fuels management on an isolated forest carnivore. Landscape Ecology 26: 1491-1504.

Schempf, P.F. and M. White. 1977. Status of six furbearer populations in the mountains of northern California. U.S. Department of Agriculture, Forest Service. 51pp.

Slauson, K. M. 2011 (September 9). Email with USDA Forest Service Wildlife Biologist Stephanie Coppeto regarding status of martens in the Lake Tahoe Basin Management Unit.

Slauson, K. M. 2013 (March 22). Email exchange with Stephanie Coppeto (wildlife biologist at Lake Tahoe Basin Management Unit) regarding comments on the Limited Operating Period for marten.

Slauson, K. M. and W. J. Zielinski. In prep. Effects of Developed Ski Areas on the Population Dynamics of the Pacific Marten in the Lake Tahoe Region of California. Final Report. Pacific Southwest Research Station, Arcata, CA 95521.

Slauson, K.M. 2003. Habitat selection by American martes (Martes americana) in coastal northwestern California. Thesis, Oregon State University, Corvallis, Oregon, USA.

Slauson, K.M. and W. J. Zielinksi. 2008. A Review of the effects of forest thinning and fuels reduction on American martens (Martes americana) pertinent to the southern Cascades region of California. Final Report, US Department of Agriculture, Forest Service, Pacific Southwest Research Station, Redwood Sciences Laboratory, Arcata, CA. p. 15

Slauson, K.M., Zielinski, W. J., and J. Baldwin. 2008. Pacific marten population monitoring in the Lake Tahoe Basin, monitoring plan development and protocol, final report. USDA Forest Service Pacific Southwest Research Station. 63 pp.

Spencer, W. and H. Rustigian-Romsos. 2012 (August). Decision-Support Maps and Recommendation for Conserving Rare Carnivores in the interior mountains of California. Conservation Biology Institute, http://consbio.org/products/reports/decision-support-maps-and-recommendations- conserving-rare-carnivores-interior-mountains-california

Turner, M.G. 1989. Landscape ecology: the effect of pattern on process. Annual review of Ecology and Systematics. 20:171-197.

Zeiner, David C., William F. Laudenslayer, Jr., Kenneth E. Mayer, and Marshall White, eds. 1990. California's Wildlife. Volume III. Mammals. California Statewide Wildlife Habitat Relationship System. Department of Fish and Game, The Resources Agency, Sacramento, California. 407 pages.

Zielinksi, W. J. January 2013. The Forest Carnivores: Fisher and Marten. Chapter 7.1 in Science Synthesis to Support Land and Resource Management Plan Revision in the Sierra Nevada and Southern Cascades. Pacific Southwest Research Station

Zielinski et al. 2007

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Zielinski, W.J. R.L. Truex, F.V. Schlexer, L.A. Campbell, and C. Carroll. 2005. Historical and contemporary distributions of carnivores in forests of the Sierra Nevada, California, USA. Journal of Biogeography 32:1385–1407.

Zielinski, W.J. Werren, J. Kirk, T. and K. Slauson. 2006. Interactions between fishers and martens in California and implications for their conservation. Oral presentation in Fisher and marten in California, moving science & management forward; symposium presented by the California North Coast Chapter of the Wildlife Society and the Western Section of the Wildlife Society, Sacramento, CA. February 7-8, 2006.

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Sierra Nevada Red Fox (Vulpes vulpes necator) Regional Forester’s Sensitive Species

Proposed Species of Conservation Concern

Species Account The red fox is widely distributed in North America, from the arctic coast of Alaska and northwest Canada, east to the Atlantic Ocean and then south to the US Gulf Coast (NatureServe 2011). The Sierra Nevada subspecies (V. v. necator), along with two other subspecies, form the mountain red fox group of western North America (Aubry 1994). Sierra Nevada red foxes were historically found in the high elevations of the Sierra Nevada mountain range and from Mount Shasta and Lassen Peak westward to the Trinity Mountains (Aubry 1994). There are historical museum collection records from Siskiyou, Tulare, Mariposa, Mono, Lassen, and Nevada counties (Museum of Vertebrate Zoology 2007).

The only population known to exist in recent times is found in Lassen National Park and the surrounding Lassen National Forest (Perrine 2005, Perrine et al. 2007). The last reliable sighting in the Sequoia National Park occurred in 1993 (Graber 2007, personal communication with G. Bolen, North States Resources). A red fox was photographed in the winter of 1990 to 1991 at the Tioga Pass Resort (9,646 feet) on the Inyo National Forest (Chow 2007, personal communication). Little is known about the habitat requirements of this subspecies. Apparently, they inhabit various habitats in alpine and subalpine zones; their preferred habitat is apparently red fir, lodgepole pine forests and alpine fell-fields (CDFG 1987, CDFG 2005, CDFG 2012).

California Department of Fish and Game uses location and elevation to distinguish the Sierra Nevada red fox from other red foxes, as there are no visible characteristics to reliably distinguish between the two (Lewis et al. 1993, Perrine 2007). Red foxes found above 3,500 feet in the Cascade or Sierra Nevada Mountains are considered a Sierra Nevada red fox (Lewis et al. 1993, Perrine 2007)

Red foxes have been sighted elsewhere in the historical range of the Sierra Nevada red fox (Perrine 2005). However, these sightings are not common and have mostly come from inexperienced observers. Thus, their accuracy is questionable (Perrine 2005). The low number of sighting reports suggests that it is unlikely that significant red fox populations exist in the heavily visited parts of the Sierra Nevada. Beyond the Region 5 boundary, the Sierra Nevada red fox may be found in the mountains of western Nevada (NatureServe 2011) and on the Humboldt-Toiyabe National Forest (CNDDB 2011). Museum collection records occur for Carson City, Clark, and Churchill counties in Nevada (Museum of Vertebrate Zoology 2007). Again, the only recent known population is around Lassen National Park, California (Perrine 2007). Figure 6 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

No population estimates for the Sierra Nevada red fox were found. A study of mesocarnivores in the Sierra Nevada and the southernmost parts of the Cascade range failed to detect the Sierra Nevada red fox in any of the 344 sample units (six track plates and one baited camera station), leading Zielinski to conclude that if the Sierra Nevada red fox was not extirpated from this region then it must occur in very low densities (Zielinski et al. 2005). The only verified sightings of Sierra Nevada red fox in recent times have come from Lassen National Park (Perrine 2005) and near the Stanislaus NF and Humboldt-Toiyabe NF boundary (CNDDB 2011) where surveys were conducted in the spring of 2011. Nine individuals were identified from the Lassen Park population through direct capture and genetic analysis of feces (Perrine et al. 2007). These samples were collected from 1998 to 2002 and are thought to represent a considerable portion of a small isolated population (Perrine et al. 2007). Historically, Sierra Nevada red foxes were never a common species, and occurred at low densities even in prime habitat (CDFG 1991).

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The US Fish and Wildlife Service received a petition to list the Sierra Nevada red fox was in May 2011 and the action is currently under review.

The Sierra Nevada red fox occurs at elevations from 4,500 feet to 11,500 feet, but is most commonly found above 7,000 feet (Aubry 1994).

Red foxes, including the Sierra Nevada red fox, mate between December and April, with most breeding occurring in January and February (Aubry 1997). Gestation lasts about 50 days where after one to ten pups are born (Aubry 1997). Average litter size is five to six pups; mountain red fox litter sizes are typically on the low end of the range for foxes (Perrine et al. 2010). Both sexes can breed their first year (Perrine et al. 2010).

Sierra Nevada red foxes consume a wide variety of animals and vegetation. Small mammals, in particular rodents form the bulk of the diet year round (Perrine et al. 2010). In the Lassen Park area, this included pocket gophers (Thomomys sp.), mice (Peromyscus sp.), voles (Microtus sp.), and ground squirrels (Spermophilus sp.) (Perrine et al. 2010). Earlier records (Grinnell 1937 as reported in Aubry 1997 and Perrine et al. 2010) indicated woodrats (Neotoma sp.), tree squirrels (Tamiasciurus sp.), chipmunks, (Tamias sp.), hares (Lepus sp.), other small mammals, and birds. In Washington, Cascade red foxes consumed pocket gophers, red-backed (Myodes [Clethrionomys] sp.) and heather (Phenacomys sp.) voles, fruit, insects, birds, and garbage in the summer (Aubry 1983, Perrine et al. 2010). During the winter Cascade red fox diets consisted mostly of snowshoe hares and other small rodents (Aubry 1983, Perrine et al. 2010). In the Lassen Park areas, pocket gophers are particularly important and hares are not abundant; combined, these two factors may explain the seasonal downhill migration of Sierra Nevada red foxes once heavy snowfall begins (Perrine 2005). Foxes across their range scavenge, and deer carcasses are commonly consumed in the Lassen area (Perrine 2005).

Habitat Status Little is known about the habitat requirements of this subspecies. Apparently, they occupy various habitats in alpine and subalpine zones; their preferred habitat is apparently red fir, lodgepole pine forests and alpine fell-fields (CDFG 1991). A study on the Lassen National Forest found that Sierra Nevada red fox detections were disproportionately abundant at camera stations in high-elevation conifer community types and were under-represented at cameras in mid-elevation conifer communities; detections also tended to be in barren areas such as talus slopes (Perrine 2005).

Home range estimates for the Sierra Nevada red fox in Lassen Volcanic National Park area vary seasonally and are at the upper end of values for red foxes (CDFG 2007, Perrine 2005, Perrine et al. 2010). Summer home ranges were about 650 to 17,250 acres (mean 5,740 acres) and winter home ranges were 800 to 15,750 acres (mean 7,740 acres) (Perrine 2005). Aubry (1983 and 1997) reported smaller home range sizes for Cascade red foxes (V. v. cascadensis), another mountain fox. Aubry found summer home ranges were 60 to 2,880 acres (mean 580 acres); during winter, 225 to 760 acres (mean 480 acres) Aubry 1983). Summer home ranges are typically smaller than are winter home ranges, presumably because deep snow cover makes foraging ground less accessible (CDFG 2007, Perrine et al. 2010).

The Sierra Nevada red fox moves seasonally between high elevation summer habitat above 6,000 feet and winter habitat at lower elevations between 4,500 to 6,000 feet (CDFG 2007). Fox presence is positively correlated with barren areas (CDFG 2007).

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Figure 6. Map of Sierra Nevada red fox locations from NRIS Wildlife Database, 2016

Specific CWHR habitat community types vary between winters (Table 4) and summer (Table 5). During the winter, home ranges include Sierran mixed conifer, red fir, aspen, montane chaparral, and white fir communities (Perrine et al. 2010). Forest stands with trees greater than 60 centimeters in diameter at breast height (CWHR class 5, medium/large tree) and greater than 40 percent canopy closure (CWHR class M moderate or D dense) were preferred (Perrine et al. 2010). Mature, closed canopy forest (5/6, S/P)

56 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests was a significant predictor of Sierra Nevada red foxes in the winter (CDFG 2007). During the summer, foxes selected barren habitats and avoided conifer, hardwoods, and herbaceous types, except for day rests, which included dense, young conifer stands, gaps in shrubs, and talus (Perrine et al. 2010). There is a negative association with pole and small tree size classes with moderate and dense canopy closure (3/4, M/D) in the summer (CDFG 2007).

Table 4. Sierra Nevada red fox habitat (CWHR) associations in the winter Habitat CWHR Habitat CWHR Size and Density Landscape-level SMC, SCN 5,6 SPMD Landscape-level ADS, ASP, BAR, LPN, MCP, RFR, WTM, MRI, All Resting SMC 4M, 4D, 5D Resting WFR 4D, 5M Resting MCP, MHW, AGS, ASP All

Table 5. Sierra Nevada red fox habitat (CWHR) associations in the summer Habitat CWHR Habitat CWHR Size and Density Landscape-level SMC, SCN 5,6 SPMD Landscape-level ADS, BAR, LPN, MCP, RFR, WTM, MRI, All Resting RFR 4P, 4M Resting SNC 3S, 4P Resting LPN 4M, 4D Resting SMC 4D Resting WFR 3M, 4D Resting BAR, MCP All Note: Information here is based on Perrine (2005), as reported in CDFG (2007) and Perrine et al. (2010). AGS = annual grassland; ADS = alpine dwarf shrub; ASP = aspen; BAR = barren; LPN = lodgepole pine; MCP = montane chaparral; MHW = montane hardwoods; MRI = montane riparian; RFR = red fir; SCN = subalpine conifer; SMC = Sierran mixed conifer; WFR = white fir; WTM = wet meadow.

Winter resting sites were associated with cavities under downed wood and tree hollows formed by low branches (Perrine et al. 2010). Summer rest areas were located in barren and high-elevation conifer (CDFG 2007). At the landscape scale, higher elevation forests are naturally more open.

Information on den structures is lacking for Sierra Nevada red foxes. They likely use rock and ground dens in talus and boulder fields, using whatever structures are available (Aubry 1997, Perrine 2005). They apparently do not use earthen dens like other red foxes (Aubry 1997, Perrine 2005).

Threats There is not a lot of information about Sierra Nevada red foxes in general, and more research is needed (Perrine et al. 2010). Nonetheless, Perrine et al. (2010) indicate the following threats: non-native red foxes; development and recreation, forest management and livestock grazing; climate change; and trapping. Because information is lacking, the authors indicate this list is speculative.

The status of non-native foxes in California is evolving. It appears that some populations of lowland (or valley) foxes are indeed non-native (imports from the Midwest United States), and their range is increasing (Perrine et al. 2010). However, it now appears that the Sacramento valley population of red foxes is a distinct and native sub-species, and is most closely related to the Sierra Nevada red fox (Perrine

57 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests et al. 2007, Sacks et al. 2010). The Sacramento Valley red fox population is geographically the closest to the Sierra Nevada red fox population in the northern part of the state. However, it is the non-native populations that are expanding their range in the state (Perrine et al. 2010), and it is unknown how this may affect the native populations of foxes. At this point, it appears the habitat of the Sierra Nevada red fox is not being invaded by lowland native and non-native foxes (Perrine et al. 2010). Concerns remain regarding interbreeding with non-natives and subsequent genetic swamping, adverse impacts of competition, and disease and parasite transmission.

Roads and urban development can increase the means by which competitors and predators access geographic areas where Sierra Nevada red foxes are found (Perrine et al. 2010). Public access areas like snow parks, campgrounds, picnic areas, and ski areas can draw foxes in to food sources (garbage), leading to begging behavior and associated conflicts, as well as adverse interactions with competitors and predators (Perrine et al. 2010). Fish stocking is another recreation-related threat; consumption of certain infected fish can lead to salmon poisoning disease, which is often fatal to canids (Perrine et al. 2010).

Past livestock grazing was particularly intense in the range of the Sierra Nevada red fox, and may have decreased rodent productivity in summer habitat (Perrine et al. 2010). Modern grazing practices have improved, and may actually improve habitat for fox prey (Perrine et al. 2010). Fire suppression likewise has altered prey habitat and subsequent prey composition; how this affects the Sierra Nevada red fox remains to be seen (Perrine et al. 2010).

Forest management that opens overly dense stands may be beneficial to some rodent populations, and as such beneficial to red foxes (Perrine et al. 2010). Perrine and others (2010) caution that changes such as this may also increase access by fox predators and competitors. There is some indication that large openings are avoided in the winter (Benson et al. 2005).

Related to both forest management and grazing is rodent control, which, if using poisons (strychnine), may kill foxes, and other carnivores and scavengers (Perrine et al. 2010).

Climate change and the expected temperature increases may seriously affect the Sierra Nevada and other mountain red foxes, as these subspecies are apparently specialized in their adaptations to cool, high elevation habitats (Perrine et al. 2010). It is likely that there would be increased predator and competitor access to former mountain red fox habitat (Perrine et al. 2010). Furthermore, it is expected there would be a reduction in alpine habitat for the Sierra Nevada red fox, including a 62 percent reduction in the Great Basin and at least a 50 percent reduction in the Sierra Nevada Mountains (Perrine et al. 2010).

Trapping red foxes in California is illegal, so any take would be incidental; nonetheless, due to low density and isolated populations, any losses could threaten the population (Perrine et al. 2010).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

58 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests linkages are key development to allow for the species to move from one area to another, hence reducing the impacts of climate change. Risks due to uncharacteristic wildfires is ameliorated by the restoration that is proposed.

Determination Statement Inyo, Sequoia and Sierra NFs: Implementation of the Forest Plan may affect Sierra Nevada red fox but will not lead towards Federal listing or a loss of viability. Impacts to marten are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags.

Literature Cited – Sierra Nevada Red Fox Aubry, K. B. 1983. The Cascade red fox: Distribution, morphology, zoogeography and ecology. Seattle: University of Washington; 151 p. Ph.D. dissertation.

Benson, J., J. Perrine, R. Golightly, R. Barrett. 2005. Use of cover and response to cover type edges by female Sierra Nevada red foxes. Western North American Naturalist 65: 127-130.

University of Washington; 151 p. Ph.D. dissertation. Aubry, K. 1994. The Sierra Nevada Red Fox (Vulpes vulpes necator). IN: Harris, J., and C. Ogan., Eds. 1997. Mesocarnivores of Northern California: Biology, Management, and Survey Techniques, Workshop Manual. August 12-15, 1997, Humboldt State Univ., Arcata, CA. The Wildlife Society, California North Coast Chapter, Arcata, CA. 127 p.

California Department of Fish and Game. 1991. Annual report on the status of California state listed threatened and endangered animals and plants.

California Department of Fish and Game. 2005. The Status of Rare, Threatened, and Endangered Plants and Animals of California 2000-2004.

California Department of Fish and Game. 2007. Sierra Nevada red fox. Available online at https://r1.dfg.ca.gov/Portal/SierraNevadaRedFox/tabid/618/Default.aspx. Accessed 9 January 2012.

CNDDB. California Department of Fish and Game, Biogeographic Data Branch. 2011. California Natural Diversity Database. Sacramento, CA. Data downloaded November 2011.

California Department of Fish and Game, California Interagency Wildlife Task Group. 2008. CWHR version 8.2 personal computer program. Sacramento, CA.

Graber, David. Chief Scientist Pacific West Region National Park Service. Email conversation with Ginger Bolen (North State Resources). Sierra Nevada red fox. October 11, 2007.

Lewis, J., K. Sallee, and R. Golightly. 1993. Introduced red fox in California. Sacramento: Report prepared for California Department of Fish and Game.

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Museum of Vertebrate Zoology. Collections database [web database]. University of California Berkeley 2007 [cited October 4, 2007]. Available from http://mvzarctos.berkeley.edu/SpecimenSearch.cfm.

NatureServe. 2011. NatureServe Explorer: An online encyclopedia of life [web application]. Version 7.1. NatureServe, Arlington, Virginia. Available http://www.natureserve.org/explorer. (Accessed: December 22, 2011).

Perrine, J. 2005. Ecology of red fox (Vulpes vulpes) in the Lassen Peak region of California, USA. Ph.D. dissertation, University of California, Berkeley, California, USA.

Perrine, J., J. Pollinger, B. Sacks, R. Barrett, and R. Wayne. 2007. Genetic evidence for the persistence of the critically endangered Sierra Nevada red fox in California. Conservation Genetics.

Perrine, J., L. Campbell, and G. Green. 2010. Sierra Nevada red fox (Vulpes vulpes necator): A conservation assessment. US Forest Service, Region 5. Vallejo, CA. Report R5-FR-010.

Sacks B., H. Wittmer, and M. Statham. 2010. The Native Sacramento Valley red fox. Report to the California Department of Fish and Game, May 30, 2010, 49pp.

U.S. Department of Agriculture Forest Service (USDA FS). 2012. Natural Resource Information System (NRIS) database. Accessed April 2012.

Zielinski, W., R. Truex, F. Schlexer, L. Campbell, and C. Carroll. 2005. Historical and contemporary distributions of carnivores in forests of the Sierra Nevada, California, USA. Journal of Biogeography 32 (8):1385-1407.

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Townsend’s Big-eared Bat (Corynorhinus townsendii townsendii) Regional Forester’s Sensitive Species

Proposed Species of Conservation Concern

Species Account For most of its taxonomic history, the recognized generic name for this North American species was Corynorhinus. Beginning, however, with a taxonomic revision by Handley (1959) it became known as Plecotus. Two recent phylogenetic studies have reviewed relationships among plecotine genera (Frost and Timm 1992, Tumlison and Douglas 1992), and have recommended resurrecting the generic name of Corynorhinus to distinguish the North American from the Palearctic forms. This change has been recognized by Wilson and Reeder (2005).

There are five currently recognized subspecies of C. townsendii in the United States (Handley 1959); two (C. t. townsendii and C. t. pallescens) in the western U.S., two (C. t. ingens and C. t. virginianus) in the eastern part of the country, and one (C. t. australis) with a primarily Mexican distribution, which overlaps with C. t. pallescens in western Texas. Only the two western subspecies are found in California (Piaggio et al 2009).

Townsend’s big-eared bat is considered a Mammal Species of Special Concern by California Department of Fish and Game and a Sensitive Species by Region Five of the U.S. Forest Service and by the Bureau of Land Management. The Western Bat Working Group granted it High Priority for its known range (www.wbwg.org/speciesinfo/species_matrix/spp_matrix.pdf). This species is covered as an evaluation species under the Lower Colorado River Multi-Species Conservation Program administered by the Bureau of Reclamation (LCR MSCP 2004).

Townsend’s big-eared bat occurs in California, Oregon, Washington, Nevada, Idaho, and possibly southwestern Montana and northwestern Utah. C. t. pallescens occurs in all the same states as C. t. townsendii, plus Arizona, Colorado, New Mexico, Texas, and Wyoming (Handley 1959). Throughout much of their range in California, Idaho, Nevada, Oregon and Washington there are extensive zones of intergradation for the two subspecies. Throughout the zone of intergradation it is frequently impossible to assign individuals to one subspecies or the other. Handley (1959) distinguishes the two subspecies based on size and color characteristics, but he also notes that the full spectrum of characteristics for both subspecies can be found within a single population. The results of preliminary mitochondrial DNA studies, using PCR techniques, failed to distinguish between these two subspecies, but this may reflect the relatively conservative region sequenced (cytochrome b) (W. Rainey). For the purposes of this document, we make no distinction between these subspecies.

In California, Townsend’s big-eared bat is found throughout much of the state, except for the Central Valley and very high elevations. The largest populations are concentrated in areas offering caves (commonly limestone or basaltic lava) or mines as roosting habitat. The species is found from sea level along the coast to 6,000 feet in the Sierra Nevada (Dalquest 1947, Pearson et al. 1952, Pierson and Rainey 1996). In the White Mountains, summer records for males extend up to 7,900 feet, and hibernating groups have been found in mines as high as 10,460 feet (Szewczak et al. 1998). Maternity colonies are more frequently found below 6,560 feet (Pierson and Fellers 1998, Szewczak et al. 1998).

Outside California it has been found to 7,900 feet (Jones 1965, Jones and Suttkus 1972) and 9,500 feet (Findley and Negus 1953).

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Surveys conducted by Pierson and Rainey (1996) show marked population declines for both subspecies in California. This species has been petitioned for listing as threatened or endangered status in the state. Over the past 40 years, there has been a 52 percent loss in the number of maternity colonies, a 45 percent decline in the number of available roosts, a 54 percent decline in the total number of animals, and a 33 percent decrease in the average size of remaining colonies for the species as a whole statewide. The status of particular populations is correlated with amount of disturbance to or loss of suitable roosting sites. The populations that have shown the most marked declines are along the coast, in the Mother Lode country of the western Sierra Nevada foothills, and along the Colorado River.

A comparison of former and current population estimates for 18 historically known maternity colonies shows that six colonies (33 percent) appear to be extirpated; six others (33 percent) have decreased in size; one (6 percent) has remained stable; and five (28 percent) (four of which are protected within national parks) have increased.

A comparison of colony size for historically and currently known colonies, indicates that mean colony size has decreased from 165 measured at 18 sites to 111 measured at 34 sites. The median colony size has decreased from 100 to 75. There are currently 38 known maternity colonies, occupying 55 known roost sites, with an estimated total population of about 4,300 individuals. Only three of these colonies have adequately protected roost sites.

Hibernating Townsend’s big-eared bat have been found historically or during a recent survey (Pierson and Rainey 1996) at 44 sites (24 in mines, 19 in caves, one in a building). Most of these sites contain fewer than 20 individuals. Only three hibernating colonies number more than 100. The most significant aggregations (all those with greater than 100) occur in the most northern part of the state, particularly Siskiyou County. In other areas, particularly the desert, smaller aggregations (5 to 20) are more typical, although mine shafts, found by Altenbach and Milford (1991) to house the largest aggregations, remain essentially unexplored in California. Four additional hibernating sites, not visited by Pierson and Rainey (1994) were located in 1979 (Marcot 1984), one of which contained 40-50 individuals.

In the NRIS database, the Inyo NF has 802 records, Sequoia NF has no records, and the Sierra NF has 11 records. Figure 7 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Townsend’s big-eared bat occurs from the inland deserts to the cool, moist coastal redwood forests, in oak woodlands of the inner Coast Ranges and Sierra Nevada foothills, and lower to mid-elevation mixed coniferous-deciduous forests. Distribution is patchy, and strongly correlated with the availability of caves and cave-like roosting habitat, with population centers occurring in areas dominated by exposed, cavity forming rock and/or historic mining districts (Genter 1986, Graham 1966, Humphrey and Kunz 1976, Kunz and Martin 1982, Perkins et al. 1994, Pierson and Rainey 1996). Its habit of roosting on open surfaces makes it readily detectable, and it is often the species most frequently observed (commonly in low numbers) in caves and abandoned mines throughout its range.

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Figure 7. Map of Townsend’s big-eared bat locations from NRIS Wildlife Database, 2016

Roosting Habitat Townsend’s big-eared bat prefers open surfaces of caves or cave-like structures, such as mines (vertical and horizontal) (Barbour and Davis 1969, Graham 1966, Humphrey and Kunz 1976). It has also has been reported in such structures as buildings, bridges, and water diversion tunnels that offer a cavernous environment (Barbour and Davis 1969, Dalquest 1947, Howell 1920, Kunz and Martin 1982, Pearson et

63 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests al. 1952, Perkins and Levesque 1987, Brown et al. 1994, Pierson and Rainey 1996). Roosting structures often contain multiple openings. It seems to prefer dome-like areas, possibly where heat or cold is trapped (warm pockets for maternal roosting, cold pockets for hibernation). It has also been reported in rock crevices and large hollow trees (Fellers and Pierson 2002). The discovery of a maternity roost in a hollow redwood tree (Mazurek 2004) suggests that coastal populations may have historically relied on these structures.

Specific roosts may be used only one time of year or may serve many different functions throughout the year (i.e. maternal, hibernation, dispersal, bachelor, breeding, etc.). Roosting surfaces often occur in twilight conditions, however, some have been located very deep inside caves or mines. There is evidence that maternity colonies may use multiple sites for different stages (pregnancy, birthing, rearing, etc.) (Pierson et al. 1991, Sherwin et al. 2000). Males remain solitary during the maternity season.

This species appears to have fairly restrictive roost requirements (Humphrey and Kunz 1976, Perkins et al. 1994, Pierson et al. 1991). Roost temperature appears to be critical (Lacki et al. 1994, Pearson et al. 1952, Pierson and Rainey 1996). Temperatures vary in maternity roosts throughout California from 66 degrees Fahrenheit in cooler regions to 86 degrees Fahrenheit in warmer southern regions (Pierson et al. 1991). Some colonies are known to change roosts during the maternity season, using cooler roosts earlier in the year (Pierson et al. 1991, P. Brown pers. comm., V. Dalton pers. comm.) and using warmer roosts after pups are born. Roost dimensions are also important. The majority of the roosts examined in California are fairly spacious, at least 100 feet in length, with the roosting area located at least 6 feet above the ground, and a roost opening at least 6 inches by 24 inches (Pierson et al. 1991). Maternity clusters are always situated on open surfaces, often in roof pockets or along the walls just inside the roost entrance, within the twilight zone.

Townsend’s big-eared bat is very sensitive to human disturbance, however, in some instances can become habituated to reoccurring and predictable human activity.

Foraging Habitat Foraging associations include edge habitats along streams and areas adjacent to and within a variety of wooded habitats (Brown et al. 1994, Fellers and Pierson 2001, Pierson et al. 2002). Recent radio-tracking and light-tagging studies have found Townsend’s big-eared bats foraging in a variety of habitats. Brown et al. (1994) showed that on Santa Cruz Island in California, they avoided the lush introduced vegetation near their day roost, and traveled up to 3 miles to feed in native oak and ironwood forest. P. Brown (pers. comm.) also documented this species foraging in desert canyons with water on the west slopes of the Panamint Mountains in Inyo County. Radio-tracking and light-tagging studies in northern California have found Townsend’s big-eared bat foraging within forested habitat (Rainey and Pierson 1996), within the canopy of oaks (E. Pierson and W. Rainey unpubl. data), and along heavily vegetated stream corridors, avoiding open, grazed pasture land (G. Fellers pers. comm.). In Oklahoma, C. t. ingens preferred edge habitats (along intermittent streams) and open areas (pastures, agricultural fields, native grass) over wooded habitat (Clark et al. 1993). Light-tagging studies in West Virginia (V. Dalton pers. comm.) showed a bimodal foraging pattern for C. t. virginianus, with animals foraging over hayfields during the first part of the night, and within the forest later in the night, traveling up to 8 miles from the day roost. Townsend’s big-eared bat has been known to travel up to 15 miles from their roost sites while foraging (Dobkin et al. 1995). They forage as long as weather permits in the fall, and are periodically active in winter (Pierson et al. 1991).

Although diet has not been examined in detail for any California populations, it is likely that Townsend’s big-eared bat here, as elsewhere, is a lepidopteran specialist, feeding primarily (greater than 90 percent of the diet) on medium sized 0.2 to 0.5 inch moths (Dalton et al. 1986, Ross 1967, Sample and Whitmore

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1993, Whitaker et al. 1977, 1981). Shoemaker and Lacki (1993) determined that C. t. virginianus differentially selected noctuid moths, with geometrids, notodontids and sphingids also making up a significant portion of the diet. Representatives of the family Arctiidae constituted 37.5 percent of the available moth prey items, but were not consumed. Sample and Whitmore (1993) identified moth species from wing fragments collected at maternity caves. Of the 28 moth taxa identified, 15 were noctuids. Twenty-one species were forest dwelling, and six were associated with open, field habitats. In addition to Lepidopterans, small quantities of other insects have been detected in the diet of Townsend’s big-eared bat, particularly Coleoptera and Diptera (Dalton et al. 1986, Ross 1967, Sample and Whitmore 1993). Hemiptera, Hymenoptera, Homoptera, Neuroptera, Trichoptera, and Plecoptera have also been found sporadically (Dalton et al. 1986, Whitaker et al. 1977).

Reproduction Townsend’s big-eared bat is a colonial species with maternity aggregations forming between March and June (based on local climate and latitude). Colony size ranges from a few dozen to several hundred. Mating generally takes place in both migratory sites and hibernacula between September or October and February. “Swarming” has been observed in the Mojave Desert in the latter half of September (P. Brown pers. comm.). Females are generally reproductive in their first year, whereas males do not reach sexual maturity until their second year. Gestation length varies with climatic conditions, but generally lasts from 56 to 100 days (Pearson et al. 1952). Some evidence shows that maternity colonies may have up to three different sites for given stages - one each for pregnancy, birthing, rearing, etc. A single pup is born between May and July (Easterla 1973, Pearson et al. 1952, Twente 1955). Pups average 0.1 ounce at birth, nearly 25 percent of the mother's postpartum mass (Kunz and Martin 1982). Young bats are capable of flight at 2.5 to 3 weeks of age and are fully weaned at 6 weeks (Pearson et al. 1952). Nursery colonies start to disperse in August about the time the young are weaned, and break up altogether in September and October (Pearson et al. 1952, Tipton 1983). Pearson et al. (1952) estimated annual survivorship at about 50 percent for young, and about 80 percent for adults. Band recoveries have yielded longevity records of 16 years, 5 months (Paradiso and Greenhall 1967) and 21 years, 2 months (Perkins 1995).

Migration and Hibernation Townsend’s big-eared bat is a relatively sedentary species, for which no long-distance migrations have been reported (Barbour and Davis 1969, Humphrey and Kunz 1976, Pearson et al. 1952). The longest movement known for this species in California is 20 miles (Pearson et al. 1952). There is some evidence of local migration, perhaps along an altitudinal gradient.

Hibernation sites are generally caves or mines (Pearson et al. 1952, Barbour and Davis 1969), although animals are occasionally found in buildings (Dalquest 1947, E. Pierson pers. obs., G. Tatarian pers. obs.). Deep mine shafts, known to provide significant hibernating sites in New Mexico (Altenbach and Milford 1991), may also be important in California (P. Brown pers. comm.). Winter roosting is typically composed of mixed-sexed groups from a single individual to several hundred or several thousand, however, behavior varies with latitude. In areas with prolonged periods of non-freezing temperatures, C. townsendii tends to form relatively small hibernating aggregations of single to several dozen individuals (Barbour and Davis 1969, Pierson et al. 1991, Pierson and Rainey 1996). Larger aggregations (75 to 460) are confined to areas which experience prolonged periods of freezing temperatures (Pierson and Rainey 1996). Studies in the western United States have shown that Townsend’s big-eared bat selects winter roosts with stable, cold temperatures, and moderate air flow (Humphrey and Kunz 1976, Kunz and Martin 1982). Individuals roost on walls or ceilings, often near entrances (Humphrey and Kunz 1976, Twente 1955). If undisturbed, individuals will frequently roost less than 10 feet off the ground (Perkins et al. 1994), and have been found in air pockets under boulders on cave floors (E. Pierson pers. obs.). Temperature appears to be a limiting factor in roost selection. Recorded temperatures in Townsend’s big-

65 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests eared bat hibernacula range from 28 degrees Fahrenheit to 55 degrees Fahrenheit (Humphrey and Kunz 1976, Genter 1986, Pearson et al. 1952, Pierson et al. 1991, Twente 1955), with temperatures below 50 degrees Fahrenheit being preferred (Perkins et al. 1994, Pierson and Rainey 1996). In the Mojave Desert ecoregion in the winter, torpid Townsend’s big-eared bat have been found at temperatures of 60 degrees Fahrenheit as these might be the coolest temperatures available (P. Brown, pers. obs). The period of hibernation is shorter at lower elevations and latitudes.

Threats Inappropriate behavior on the part of well-intentioned researchers and others (i.e., entry into maternity roosts, capture of animals in roosts) could also contribute to population declines. Mark recapture studies are not without risk, since at least one wing band design causes serious injuries to Townsend’s big-eared bat (Pierson and Fellers 1994). Scientific collecting likely resulted in the extirpation of a population at Prisoner’s Harbor on Santa Cruz Island (Brown et al. 1994).

The combination of restrictive roost requirements and sedentary behavior suggests that Townsend’s big- eared bat is roost limited, and that roost loss, through disturbance or destruction, has been primarily responsible for population declines in most areas. Although fire, winter storms, or general deterioration are sometimes responsible, in all but two of 39 documented cases, roost loss in California can be directly linked to human activity (e.g., demolition, renewed mining, entrance closure, human induced fire, renovation, or roost disturbance). Population declines are most highly correlated with roost destruction in the San Francisco Bay area, along the northern coast, and in San Diego County, and with roost disturbance in the Mother Lode country and along the Colorado River.

Anthropogenic Roosts Although Townsend’s big-eared bat is often found using man-made structures, such as barns, large houses, historic buildings, and bridges, they are very sensitive to disturbance, and will readily abandon a day roost, particularly a maternity roost, if disturbed. Bats are often not tolerated in historic structures, even those that are not open to the public, due to concerns over damage to the historic fabric of a building, so even a rare species such as Townsend’s big-eared bat, one that forms relatively small colonies, is subject to permanent loss of critical roost habitat. Because Townsend’s big-eared bat is a large cavity- roosting species, and not a crevice-roosting species, they will not utilize bat houses as replacement habitat, so loss of structure roosts is highly significant for this species.

The tendency for Townsend’s big-eared bat to roost in highly visible clusters on open surfaces, near roost entrances, makes them highly vulnerable to negative human interactions. Inadequate management policies on public lands can lead to roost destruction. Of the 20 largest currently known colonies in California, 13 are on public lands. While the National Park Service and California Department of Parks and Recreation have made substantial commitments to protecting known roosts in some parks, they have failed to provide adequate protection in others. Other agencies have been less willing to recognize the biological significance of cave and mine roosts, often against the advice of their own biologists.

Caves Maternity colonies are impacted by inappropriate cave closures or disturbance during human visitation. The increasing and intense recreational use of caves in California provides the most likely explanation for why most otherwise suitable, historically significant roosts are currently unoccupied. It is well documented that Townsend’s big-eared bat is so sensitive to human disturbance that simple entry into a maternity roost can cause a colony to abandon or move to an alternate roost (Pearson et al. 1952, Graham 1966, Stebbings 1966, Mohr 1972, Humphrey and Kunz 1976, Stihler and Hall 1993, P. Brown pers. comm.).

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Forest Management This issue is restricted to commercially harvested areas of the State, particularly eastern and northern California. Large hollow redwood and sequoia offer cave-like structures for maternal roosting. Other conifer and hardwood snags offer male roosting sites. Harvested areas can also affect riparian edge habitats for foraging. Harvesting may alter microclimates around caves and mines, possibly rendering them uninhabitable.

Forest management activities, particularly timber harvest and spraying that kills non-target Lepidopteran species, may alter the prey base for Townsend’s big-eared bat. Perkins and Schommer (1991) suggest that Bacillus thuringiensis sprays may suppress Tussock moth and spruce budworm reproduction enough to suppress reproduction in resident Townsend’s big-eared bat.

Mines Maternity colonies are impacted by renewed mining activities, inappropriate mine closures, and disturbance during human visitation.

Old mines are significant roosting habitat for a number of bat species, particularly Townsend’s big-eared bat (Altenbach and Pierson 1995, Pierson and Rainey 1991, P. Brown pers. comm.). The intense recreational use of mines in California provides the most likely explanation for why most otherwise suitable, historically significant roosts are currently unoccupied. It is well documented that Townsend’s big-eared bat is so sensitive to human disturbance that simple entry into a maternity roost can cause a colony to abandon or move to an alternate roost (Pearson et al. 1952, Graham 1966, Stebbings 1966, Mohr 1972, Humphrey and Kunz 1976, Stihler and Hall 1993, Brown and Berry 2003, P. Brown pers. comm.). Liability and safety concerns have led to extensive mine closure programs in western states, particularly on public lands, often without consideration for the biological values of old mines. If non-bat compatible closures (backfilling or blasting) are done without prior biological survey or if surveys are conducted at the wrong time of year (Altenbach 1995, Navo 1995, Rainey 1995), they can result in the entrapment, and thus elimination of entire colonies. Even if the bats are excluded prior to hard closure, they may not be able to find suitable replacement habitat.

The resurgence of gold mining in the West potentially threatens cave dwelling bat species (Brown and Berry 1991, Brown et al. 1993, Brown 1995). Since open pits, created by current mining practices, are often located in historic mining districts, old mine workings are frequently demolished as part of the ore extraction process. While effective mitigation is possible (Pierson 1989, Pierson et al. 1991), there is currently no legal mandate requiring that existing populations be protected. Renewed mining is known to account for the loss of one substantial colony in the Panamint Mountains of the California desert (P. Brown pers. comm.).

Additionally, process water containing cyanide has caused substantial wildlife mortality at a number of mine sites in the West. Although one study found that bats constitute 33.7 percent of documented wildlife fatalities (Clark and Hothem 1991), they frequently are not considered in assessment of cyanide risks (Nevada Mining Assoc. et al. 1990). A Corynorhinus maternity colony in a mine on the west slope of the Inyo Mountains disappeared after an open cyanide pond was constructed within 1.2 miles of the roost (P. Brown pers. comm.). Similarly, process residues in open oil sumps are another significant source of wildlife mortality (Flickinger and Bunck 1987, Esmoil and Anderson 1995).

Urban Development Urban expansion often leads to the removal of older buildings that provide access to internal roosting structures. Newer buildings often do not provide inner access. Foraging habitat is most often removed.

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Transportation Townsend’s big-eared bat colonies could also be impacted by bridge modifications. The mandate for earthquake retrofitting on bridges could either disturb active roosts or render roost sites unsuitable. A number of older bridges are being removed and replaced with those that have bat-unfriendly designs. There is a potential loss of riparian habitat for foraging where bridges are constructed.

Agricultural Practices There is a risk from coastal agriculture from Santa Barbara to San Francisco. Population declines along the Colorado River are also attributable to foraging habitat loss due to agricultural conversion (Brown and Berry 2003).

Rangeland Management The presence of livestock can severely reduce ground and shrub cover (when not managed properly) which can lead to a reduction in prey species abundance. Many species of bats do benefit from properly designed water impoundments as a drinking source.

Although the effects of grazing have not been specifically addressed for this species, a radio-tracking study at Point Reyes National Seashore indicated that telemetered bats avoided grazed pastureland (E. Pierson pers. obs.).

Solar Energy Some proposed solar farms in the Mojave Desert are near mines sheltering Townsend’s big-eared bat maternity colonies (P. Brown, pers. comm.) Radio-telemetry surveys would be useful to determine how much foraging habitat would be removed by these projects. Increased road construction and human access in these areas could lead to roost disturbance if the mines were not protected.

Water Management Water impoundments could inundate active roost sites, potentially trapping and killing entire colonies, as well as, removing the structure altogether. Foraging areas in the riparian zones are likely to be removed. Often roosts close to water.

Roosting areas adjacent to water sources may be essential for the larger maternity colonies of Townsend’s big-eared bat in the desert (P. Brown pers. comm.).

Wind Energy Townsend’s big-eared bat are known to fly very high when commuting to and from roost sites to reach foraging habitat. An increase in wind facilities could cause a high degree of mortality, especially if they are sited near a roost site.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

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Alternative B, C, and D: Bats were given specific recognition in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Snags were provided for in each alternative. Reduction to impacts of climate change due to uncharacteristic wildfires is addressed by the restoration that is proposed. Bat gates are proposed for protection of caves and mines to allow for decreased disturbance to the bats, as well as reduced risks of disease transmission.

Determination Statement Inyo, Sequoia, and Sierra NFs: Implementation of the Forest Plan may impact Townsend’s big-eared bat but will not lead towards Federal listing or a loss of viability. By reducing uncharacteristic wildfires, providing large trees and large snags, and providing protection to caves and mines, the management framework of the Forest Plans should be beneficial to Townsend’s big-eared bats.

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Altenbach, J.S. and Milford, H.E., 1991. A program to evaluate bat use and occupancy of abandoned mines in New Mexico. Abstract. Bat Research News, 32, p.63.

Altenbach, J.S. and Pierson, E.D., 1995. The importance of mines to bats: an overview. Inactive mines as bat habitat: guidelines for research, survey, monitoring and mine management in Nevada, Biological Resources Research Center, University of Nevada, Reno, pp.7-18.Barbour, R.W., and W.H. Davis. 1969. Bats of America. University of Kentucky Press, Lexington, KY, 286 pp.

Brown, P. E. 1995. Impacts of renewed mining in historic districts and mitigation for impacts on bat populations. Pages 138-140 in B. R. Riddle, editor. Inactive mines as bat habitat guidelines for research, survey, monitoring and mine management in Nevada. Biological Research Center, University of Nevada, Reno, Nevada, USA.

Brown, P.E., and R.D. Berry. 1991. Bats: habitat, impacts and mitigation. In: Proceedings V. Issues and technology in the management of impacted wildlife; April 8-10. Boulder, CO: Thorne Ecological Institute: pp. 26-30.

Brown, P.E. and R.D. Berry. 2003. Baseline surveys and the development of monitoring protocol for Lower Colorado River bat species; May 2001 through September 2002. Report prepared for National Fish and Wildlife Foundation, Lower Colorado River Multi-Species Conservation Program, Project # 2000-0304-002, Washington D.C. 76 pages.

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Brown, P.E., R. Berry, and C. Brown. 1994. Foraging behavior of Townsend’s big-eared bats (Plecotus townsendii) on Santa Cruz Island. Pp 367-369 in W.L. Halvorson and G.J. Maender, editors. Forth

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Clark Jr, D.R. and Hothem, R.L., 1991. Mammal mortality at Arizona, California, and Nevada gold mines using cyanide extraction. California Fish and Game 77(2), pp.61-69.

Clark, B.S., D.M. Leslie, Jr., and T.S. Carter. 1993. Foraging activity of adult female Ozark big-eared bats (Plecotus townsendii ingens) in summer. Journal of Mammalogy 74(2):422-427.

Dalquest, W.W. 1947. Notes on the natural history of the bat Corynorhinus rafinesquii in California. Journal of Mammalogy 28(1):17-30.

Dalton, V.M., V.W. Brack, and P.M. McTeer. 1986. Food habits of the big-eared bat, Plecotus townsendii virginianus, in Virginia. Virginia Journal of Science 37:248-254.

Dobkin, D.S., R.D. Gettinger, and M.G. Gerdes. 1995. Springtime movements, roost use, and foraging activity of Townsend's big-eared bat (Plecotus townsendii) in central Oregon. Great Basin Naturalist 55(4): 315-321.

Easterla, D. A. 1973. Ecology of the 18 species of Chiroptera at Big Bend National Park, Texas. Northwest Missouri State Univ. Studies 34(2-3):1-165.

Esmoil, B.J. and Anderson, S.H., 1995. Wildlife mortality associated with oil pits in Wyoming. Prairie Naturalist 27: 81-81.

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Fellers, G. M., and E. D. Pierson. 2002. Habitat use and foraging behavior of Townsend's big-eared bat (Corynorhinus townsendii) in coastal California. Journal of Mammalogy 83: 167-177.

Findley, J.S., and N.C. Negus. 1953. Notes on the mammals of the gothic region, Gunnison County, Colorado. Journal of Mammalogy 34(2):235-239.

Flickinger, E.L., and C.M. Bunck. 1987. Number of oil-killed birds and fate of bird carcasses at crude oil pits in Texas. Southwestern Naturalist 32(3):377-381.

Frost. D. R., and R. M. Timm. 1992. Phylogeny of plecotine bats (Chiroptera: "Vespertilionidae"): Summary of the evidence and proposal of a logically consistent taxonomy. American Museum Novitates 3034:1-16.

Genter, D.L. 1986. Wintering bats of the Upper Snake River plain: occurrence in lava-tube caves. Great Basin Naturalist, 46(2):241-244.

Graham, R.E. 1966. Observations on the roosting habits of the big-eared bat, Plecotus townsendii, in California limestone caves. Cave Notes, 8:17-22.

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Jones, C. 1965. Ecological distribution and activity periods of bats of the Mogollon Mountains area of New Mexico and adjacent Arizona. Tulane Studies in Zoology 12(4):93-100.

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Pearson, O. P., M. R. Koford, and A.K. Pearson. 1952. Reproduction of the lump-nosed bat (Corynorhinus rafinesquei) in California. Journal of Mammalogy 33(3): 273-320.

Perkins 1995

Perkins, J.M., and C.E. Levesque. 1987. Distribution, status and habitat affinities of Townsend’s big-eared bat (Plecotus townsendii) in Oregon. Oregon Department of Fish & Wildlife Technical Report 86- 6-01:1–50.

Perkins, J.M. and T. Schommer. 1991. Survey protocol and an interim species strategy for Plecotus townsendii in the Blue Mountains of Oregon and Washington. Unpublished Report, Wallawa- Whitman National Forest, Baker, Oregon, USA.

Perkins, J.M., J. R. Peterson, and A. J. Perkins. 1994. Roost selection in hibernating Plecotus townsendii. Bat Research News.35:110.

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Pierson, E.D. and G. M. Fellers. 1994. Injuries to Plecotus townsendii from lipped bands. Bat Research News 34:89-91

Pierson, E.D. and G.M. Fellers. 1998. Distribution and ecology of the big-eared bat, Corynorhinus townsendii in California. Biological Resources Division, U.S. Geological Survey, Species at Risk Report, 92 pp.

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Pierson, E.D. and W.E. Rainey. 1996a. The distribution, status and management of Townsend’s big-eared bat (Corynorhinus townsendii) in California. Calif. Dept. of Fish and Game, Bird and Mammal Conservation Program Rep. 96-7. 49 pp.

Pierson, E. D., P. W. Collins, W.E. Rainey, P.A. Heady, and C.J. Corben. 2002. Distribution, status and habitat associations of bat species on Vandenberg Air Force Base, Santa Barbara County, California, Santa Barbara Museum of Natural History, Santa Barbara CA. Technical Report No. 1:1-135.

Pierson, E. D., W. E. Rainey, and D.M. Koontz. 1991. Bats and mines: experimental mitigation for Townsend's big-eared bat at the McLaughlin Mine in California. Pp. 31-42, in Issues and technology in the management of impacted wildlife, Snowmass, CO. April 8-10, 1991, Proceedings, Thorne Ecological Institute.

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Rainey, W.E. and E.D. Pierson. 1996. Cantara spill effects on bat populations of the upper Sacramento River, 1991- 1995. Unpublished Report, California Department of Fish and Game, Redding, California, USA.

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Sample, B.E., and R.C. Whitmore. 1993. Food habits of the endangered Virginia big-eared bat in West Virginia. Journal of Mammalogy, 74(2):428-435.

Sherwin, R. E., W. L. Gannon, J.S. Altenbach, and D. Stricklan. 2000. Roost fidelity of Townsend’s big- eared bat in Utah and Nevada. Transactions of the Western Chapter of the Wildlife Society 36:15- 20.

Sherwin, R. E., D. Stricklan, and D.S. Rogers. 2000. Roosting affinities of Townsend's big-eared bat (Corynorhinus townsendii) in northern Utah. Journal of Mammalogy 81(4): 939-947.

Shoemaker, L. G., and M. J. Lacki. 1993. Selection of lepidopteran prey by Plecotus townsendii virginianus in the Daniel Boone National Forest of Kentucky. Bat Research News 34:128

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Stebbings, R. E. 1966. Bats under stress. Studies in Speleology 1:168-173.

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Pacific Fringed-tailed Myotis (Myotis thysanodes) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account Fringed-tailed myotis was originally described in 1897 (Miller 1897), and has never been known by any other name. Most fringed-tailed myotis in California are referable to M. t. thysanodes; populations in the northwestern part of the state (Humboldt, Siskiyou and Shasta counties) have recently been placed in the new subspecies, M. t. vespertinus (Manning and Jones 1988), although relatively few specimens have been examined and the boundary between subspecies has not been clearly delineated.

The type locality for M. thysanodes is Old Fort Tejon (at Tejon Pass) in the Tehachapi Mountains, Kern County, California (Miller 1897). Four subspecies are recognized (Hall 1981, Manning and Jones 1988): M. t. aztecus, M. t. thysanodes, M. t. pahasapensis, and M. t. vespertinus. M. t. pahasapensis in western South Dakota, western Nebraska and eastern Wyoming, M. t. aztecus in southern Mexico (Hall 1981), and M. t. vespertinus in southwestern Washington, western Oregon, and northwestern California (Manning and Jones 1988). M. t. thysanodes, the primary subspecies found in California, ranges from southern British Columbia (Rasheed et al. 1995) to southern Mexico (Hall 1981).

Fringed-tailed myotis is not considered a Mammal Species of Special Concern by California Department of Fish and Wildlife but is a Sensitive species for the Bureau of Land Management and the Forest Service. The Western Bat Working Group granted it High Priority for most of its range in California. (www.wbwg.org/speciesinfo/species_matrix/spp_matrix.pdf).

The limited data available suggest serious population declines. Maternity colonies identified between 1891 (Old Fort Tejon) and the early 1970s (Point Reyes National Seashore, Marin County) were likely considerably larger than any colonies known today. Forty-two animals were collected at the Fort Tejon site (five different collections between 1891 and 1945), 58 at Point Reyes National Seashore between 1973 and 1974, 40 in one year from a site in Napa County, 20 from a Tuolumne County site, and 14 from a Kern County site. Although, in the context of surveys not targeting this species, we have identified six new maternity sites in northern California, none of these contains more than 10-30 females. One site in Napa County was described by Dalquest (1947) as having about 50 animals in July 1945. Forty animals were collected at that time. In June 1987 the site contained 10-15 animals, and in August 1988, there were none. The grounds around this building had been considerably modified in 1988 for a new winery installation, and the building which housed the bats was experiencing more human activity and scheduled for renovation. P. Brown (pers. comm.) observed two somewhat larger colonies (40 to 50 animals) in southern California, although one was in a house from which it has since been excluded. This species appears to be extremely sensitive to disturbance at roost sites and to human handling. While some species of Myotis, like Myotis yumanensis, seem tolerant of human incursions into their roosting space, fringed- tailed myotis is not. A cave in Sequoia National Park was documented in 1951 as being a fringed-tailed myotis maternity site. Sixteen animals were collected at that time. Additionally, this cave has experienced very heavy recreational use for many years. Repeated attempts by the National Park Service to gate the cave have been thwarted by vandalism. Although fringed-tailed myotis has been mist-netted in the vicinity of this cave, it has not apparently been observed roosting there recently.

A comparison of historic and current records indicates limited re-colonization at sites from which it has been extirpated. What may have been the largest documented colony in California occupied a barn at Point Reyes National Seashore. Fifty-eight animals were collected from this site in 1973 and 1974. Monitoring of this site since 1979 showed annual reoccupation by a Myotis yumanensis maternity colony,

74 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests but fringed-tailed myotis was not detected until 1996. The site has been protected by the National Park Service for at least ten years, with no known human incursions into the roosting space.

Fringed-tailed myotis widely distributed across southern British Columbia, Washington, Oregon, Idaho, Montana, Wyoming, Colorado, Utah, Nevada, California (including Santa Cruz Island), Arizona, New Mexico, western Texas, western South Dakota, western Nebraska, and south to Chiapas, Mexico.

In California, the species is found the length of the state, from the coast (including Santa Cruz Island) to greater than 5,900 feet in the Sierra Nevada. Records exist for the high desert and east of the Sierra Nevada (e.g., lactating females were captured in 1997 by P. Brown near Coleville on the eastern slope of the Sierra Nevada). However, the majority of known localities are on the west side of the Sierra Nevada. Museum records suggest that while fringed-tailed myotis is widely distributed in California, it is everywhere rare. Our personal experience is that although this species occurs in netting and night roost surveys in a number of localities, it is always one of the rarest taxa (Pierson et al. 1996). Available museum records offer documentation for only six maternity sites: two in Kern County (including the type locality at Old Fort Tejon), and one each in Marin, Napa, Tuolumne, and Tulare counties. Investigation of four of these sites since 1990 has shown that while the roosts are still available this species is no longer present at any of these sites.

In the NRIS database, the Inyo NF has no records, Sequoia NF has no records, and the Sierra NF has 20 records. Figure 8 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Fringed-tailed myotis occurs in xeric woodland (oak and pinyon-juniper most common (Cockrum and Ordway 1959, Hoffmeister and Goodpaster 1954, Jones 1965, O’Farrell and Studier 1980, Roest 1951), hot desert-scrub, grassland, sage-grassland steppe, spruce-fir, mesic old growth forest, coniferous and mixed deciduous/coniferous forests (including multi-aged sub-alpine, Douglas fir, redwood, and giant sequoia) (O’Farrell and Studier 1980, Pierson and Heady 1996, Pierson et al. 2006, Weller and Zabel 2001). In a study in the Mogollon Mountains of New Mexico and Arizona, Jones (1965) found M. thysanodes occurred almost exclusively in evergreen forest (greater than 6,600 feet elevation), and was the fourth most common species in this habitat. Barbour and Davis (1969) found it to be one of the more common species in oak forest at 4,900 feet to 5,900 feet elevation in the Chiricahua Mountains. In a long- term study in western New Mexico (Jones and Suttkus 1972), fringed-tailed myotis was found predominantly at the highest elevation sampled (8,500 feet), and was the ninth most common bat species in this habitat.

In mist-netting surveys it is often found on secondary streams. Although nowhere common, the species occurs in netting records from sea level to at least 6,500 feet in the Sierra Nevada, California. It occurs primarily from sea level to approximately 3,900 feet to 6,900 feet (O’Farrell and Studier 1980) with an isolated record from 9,500 feet in New Mexico (Barbour and Davis 1969).

A paucity of records makes it difficult to assess habitat preferences for this species in California. The earliest records for the state (Grinnell 1933) are all between 1,200 feet and 3,000 feet elevation. Orr (1956) in reviewing specimens held at the California Academy of Sciences, notes two localities from the coastal region (Carmel in Monterey County and Woodside in San Mateo County). P. Brown (pers. comm.) reports finding a colony in 1991 at Big Bear in the San Bernardino Mountains. More recently, records have accumulated from the upper Sacramento River (Rainey and Pierson 1996), and the Sierra Nevada (Pierson and Rainey unpubl. data). Although nowhere common, the species occurs as one of the rarer taxa in netting records from the central coast to at least 6,400 feet in the Sierra Nevada. It has been found in

75 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests mixed deciduous and coniferous forest and in both redwood and giant sequoia habitat (Pierson and Rainey unpubl. data).

Figure 8. Map of Pacific fringe-tailed myotis locations from NRIS Wildlife Database, 2016

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Roosting Habitat Studies conducted in California, Oregon, and Arizona, have documented that M. thysanodes roosts in tree hollows, particularly in large conifer snags (Cross and Clayton 1995, Chung-MacCoubrey 1996, Rabe et al. 1998, Weller and Zabel 2001, Pierson et al. 2006). Most of the tree roosts were located within the tallest or second tallest snags in the stand, were surrounded by reduced canopy closure, and were under bark (Ibid.). In California, a small colony was located in a hollow redwood tree in the Carmel Valley (Pierson and Rainey unpublished observation). Tree roosting behavior is consistent with an observed association between M. thysanodes and heavily forested environments in the northern part of its range (M. Brigham pers. comm., Cross et al. 1976, E. Pierson and W. Rainey pers. obs.).

Fringed-tailed myotis is also known to use a variety of roost sites, including rock crevices (Cryan 1997), caves (Baker 1962, Burt 1934, Commissaris 1961, Easterla 1966, 1973), mines (Cahalane 1939, Cockrum and Musgrove 1964), buildings (Barbour and Davis 1969, Musser and Durrant 1960, O’Farrell and Studier 1980, Orr 1956, Studier 1968), and bridges. It is also one of the species thought to be most reliant on abandoned mines (Altenbach and Pierson 1995).

Fringed-tailed myotis is a colonial roosting species. Colonies can be up to 2,000 individuals (Barbour and Davis 1969, O’Farrell and Studier 1975, P. Brown pers. comm.), but in California in recent years smaller colonies of 25 to 50 are more typical (E. Pierson and W. Rainey unpubl. data). Within buildings, this species tends to roost in the open in tightly packed clusters, mostly utilizing the sides of ceiling joists (O’Farrell and Studier 1980). Any of these types of structures are used as both day and night roosts (Barbour and Davis 1969).

Work by Studier and O’Farrell (1972) on a colony in New Mexico suggested that fringed-tailed myotis could fly at lower ambient temperature than many species, and sought cooler roosting conditions than did M. lucifugus with which it shared an attic roost. The two mine roosts which were identified recently in California were both relatively cool and damp (one mine had standing water). In contrast, a mine used as a nursery roost in the southern Sierra Nevada is dry and moderately warm (P. Brown pers. comm.). Barbour and Davis (1969) noted that this species was readily captured at the entrances to night roosts in buildings, mines and caves. In a five year study on the upper Sacramento River, M. thysanodes, though one of the least commonly encountered bats, was more readily detected at bridge night roosts than in netting surveys conducted over water (Rainey and Pierson 1996).

This species shows high roost site fidelity (O’Farrell and Studier 1980), especially when roost structures are durable or in low availability (Brigham 1991, Kunz 1982, Kunz and Lumsden 2003, Lewis 1995). Weller and Zabel (2001) noted frequent roost switching in tree roosts, but high fidelity to a given area. Roost switching has also been reported for caves (Baker 1962) and buildings (O’Farrell and Studier 1973, Studier and O’Farrell 1972). This species is highly sensitive to roost site disturbance (O’Farrell and Studier 1973, 1980).

Foraging Habitat This species often forages along secondary streams, in fairly cluttered habitat. It also has been captured over meadows (Pierson et al. 2001). Only limited information is available on diet in fringed-tailed myotis. In a study conducted in New Mexico, Black (1974) concluded the species appeared to be a beetle strategist. In western Oregon (Whitaker et al. 1977), the dominant prey item in the diet of three out of four animals examined was Lepidopterans (moths). The diet also included phalangids (harvestmen), gryllids (crickets), tipulids (crane flies), and araneids (spiders). The feces of one individual captured on the upper Sacramento River in California contained predominantly coleopterans (beetles) and Hemipterans (bugs) (Rainey and Pierson 1996). Relatively heavy tooth wear on animals examined in a five year study on the Sacramento River would suggest that in this area the species feeds primarily on heavy bodied insects,

77 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests such as Coleopterans and Hemipterans. The presence of non-flying taxa in the diet of the Oregon animals suggests a foraging style that relies at least partially on gleaning.

Fringed-tailed myotis is known to fly during colder temperatures (Hirshfeld and O’Farrell 1976) and precipitation does not appear to affect emergence (O’Farrell and Studier 1975). Post-lactating females have been known to commute up to 8 miles with a 3,100 feet elevation gain between a roost and foraging area (Miner and Brown 1996).

Reproduction Maternity roosts have been found in sites that are generally cooler and wetter than is typical for most other Vespertilionids. Recent radio-tracking studies in the forested regions of northern California have shown that this species forms nursery colonies in predominantly early to mid- decay stage, large diameter snags 23 inches to 66 inches in diameter at breast height (Weller and Zabel 2001).

Clough Cave in Sequoia National Park is the only cave found in California housing a maternal colony, for which there are multiple records. Outside of California maternity colonies have been found in caves (e.g. Baker 1962, Easterla 1966, Judd 1967). The majority of maternal roost sites documented in California have been found in buildings (e.g., Orr 1956), including the type locality at Old Fort Tejon (Miller 1897). Mines are also used as roost sites (Cahalane 1939, Cockrum and Musgrove 1964, Barbour and Davis 1969). Since 1987, two small maternity roosts in mines were located (about 10 adult females each) in the coast range north of San Francisco. P. Brown (pers. comm.) in 1992 also located a maternity colony of about 50 in a mine in the southern Sierra foothills, and in 1991 captured lactating females entering a mine in the Castle Mountains, in the eastern Mojave Desert. Five roosts in the Laguna Mountains, San Diego County, located by radio-telemetry in the summer of 1996, were in rock crevices on cliff faces (Miner et al. 1996).

Mating occurs in fall following break-up of maternity colony. Ovulation, fertilization, and implantation occur from April to May and are followed by a gestation of 50 to 60 days. One young is born from May to July, capable of flight in 16 days, and volant within 20 days. Prenatal and postnatal growth has been described by O’Farrell and Studier (1973). Young are born unfurred, with their eyes open, at about 22 percent adult weight.

Available evidence suggests that births take place earlier in California. In Napa County, females in late stage pregnancy have been observed in early May, and young 10 to 14 days old by the third week in May (Pierson and Rainey unpubl. data). Farther north, in Shasta County, females in late pregnancy or with newly born young were observed in late May and early June for three consecutive years from 1992 to 1994 (Rainey and Pierson unpubl. data).

Migration/Hibernation Winter behavior is even more poorly understood than summer behavior. Fringed-tailed myotis is thought to migrate short distances to lower elevations or more southern areas (O’Farrell and Studier 1980). Scattered winter records suggest, however, that the species does not complete long distance migrations, and like many species in the more temperate parts of California, may be intermittently active throughout the winter (O’Farrell and Studier 1980). The species has been found hibernating in buildings and mine tunnels along the coast in the San Francisco Bay area and in the coast range north of San Francisco.

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Threats

Anthropogenic Roosts Although fringed-tailed myotis does not occur in urban areas, it has often been found in buildings in rural and semi-rural settings (e.g. wineries, Hearst Castle, Big Bear attic, Bale Grist Mill State Historic Park). These colonies are typically at high risk for negative human interactions.

A significant number of the few known maternity roosts in California are in historic buildings. Restoration of historic buildings may pose a threat to this species. One historic roost site (Old Fort Tejon) and two current roost sites are located in historic buildings owned by the California Department of Parks and Recreation. Another is located in a utility building on a State wildlife refuge. There are no known protective measures in place. The tendency for bats to occupy historic buildings creates potential conflicts between the goals of historic preservation, access for public education, and wildlife protection. Although these conflicts are generally resolvable, and bat populations can almost always be accommodated in buildings without damaging historic values, this is frequently not appreciated.

Urban expansion often leads to removal of older buildings that provide potential roosts. Newer buildings generally do not provide suitable roosting habitat.

Intervention by pest control operators and public health departments can result in the elimination of many roost sites.

Forest Management Fringed-tailed myotis appears to be highly dependent on tree roosts within forest and woodland habitats and potentially requires denser vegetation for foraging. In some forested settings, it appears to rely heavily on tree cavities and crevices as roost sites (Weller and Zabel 2001), and may be threatened by certain timber harvest practices. For example, Chung-MacCoubrey (1996) in Arizona found that this species prefers large diameter (18 to 26 inches diameter at breast height) conifer snags.

The removal of snags and hardwoods during timber harvesting and the loss of hardwoods through conifer and brush competition (from a lack of fire management) have caused reductions for both roosting structures and foraging habitat. These practices are likely to be more severe on private lands. An increased demand for firewood can also leads to a decrease in available snags as roosts.

Increasing tree densities in forest settings could limit foraging and flight access.

Mines The resurgence of gold mining in the western United States potentially threatens mine dwelling bat species (Brown and Berry 1991, Brown et al. 1993, Brown 1995). Since open pits, created by current mining practices, are often located in historic mining districts, old mine workings are frequently demolished as part of the ore extraction process. While effective mitigation is possible (Pierson 1989, Pierson et al. 1991), there is currently no legal mandate requiring that existing populations be protected. Renewed mining is known to account for the loss of one substantial colony in the California desert (P. Brown pers. comm.). The largest known fringed-tailed myotis maternity colony, in California, has been located in a mine.

Increasing recreational mine exploration has also played a significant role in roost disturbance and abandonment. Human interactions have often led to the death of bats through vandalism and malicious mischief or untimely arousals. Aggressive mine closure programs for hazard abatement have been

79 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests underway for ten or more years in a number of western states. Until very recently, most closures were undertaken without any prior biological assessment.

Closure of old mines for hazard abatement and renewed mining in historic districts both pose considerable risks to this and other cavern dwelling bat species (Belwood and Waugh 1991, Brown and Berry 1991, Altenbach and Pierson 1995, Riddle 1995). One of two fringed-tailed myotis mine nursery sites found since 1987 has been destroyed by renewed mining. The colony persists by default, now occupying the lower level of a mine gated as a mitigation site for Townsend’s big-eared bat (Pierson et al. 1991).

Urban Development Urbanization typically destroys roosting and foraging habitat for this species. Urban expansion often leads to the removal of older buildings that provide access to internal roosting structures. Newer buildings often do not provide suitable roosting. Foraging habitat is often greatly reduced or removed. Colonies within human dwellings become vulnerable to disturbance and vandalism of their roosts.

Agricultural Practices Conversion of forests for crops, such as vineyards, eliminates roosting and foraging opportunities for fringed-tailed myotis.

Pesticides have been shown to have detrimental effects on bat populations (Clark 1981, Clark et al. 1978, Clark et al. 1983). Persistent chlorinated hydrocarbons are now banned. While the shorter half-life organophosphates, now in wide use, are known to have negative impacts on raptor (Wilson et al. 1991), their effect on bats has not been investigated. Short-term neurotoxic insecticides could be lethal or impair maneuverability, leading to reduced foraging efficiency and increased vulnerability to predators. The loss of prey species to insecticides could also affect foraging success.

Caves In the southern ecoregions (San Diego) there is no evidence of cave use, however, in other parts of the state, fringed-tailed myotis can be located in caves.

The increasing and intense recreational use of caves in California could lead to roost abandonment, death of young or adults, or untimely arousals during hibernation.

Oak Woodlands Intact oak woodlands are often used as foraging habitat. The loss of hardwoods due to firewood cutting, urban expansion, conversion to agriculture, rangeland management, and disease (e.g. Sudden Oak Death Syndrome) has caused a serious reductions for both roosting and foraging habitat.

Rangeland Management Fringed-tailed myotis frequently forages along riparian corridors or over meadows. To the extent that intensive grazing, trampling, and befouling of meadows or riparian habitats by livestock alters the insect diversity and productivity, it may affect foraging success. The presence of livestock can severely reduce hardwood recruitment. Loss of oak woodland recruitment trees, due to browsing and acorn consumption by livestock, could affect foraging or roosting habitat into the future. They can also greatly reduce ground cover (when not managed properly) which can lead to a reduction in prey species abundance.

Many species of bats do benefit from properly designed water impoundments as a drinking source. Fringed-tailed myotis is known to utilize water troughs.

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Transportation Bridge retrofitting often renders bridges unsuitable (day and night roosts) and/or disturbs colonies that are present during construction. There would likely be a loss of riparian habitat for foraging where bridges are constructed. River drainages, because they frequently offer the easiest routes through mountain ranges, are favored corridors for highway construction. Such construction commonly entails blasting of cliff faces, either for initial highway construction or later improvements (i.e., widening and straightening). Cliff roosting species are at risk of both direct impacts from blasting, and long-term loss of roosting habitat from cliff modifications. In some settings, it is possible that soil removal and blasting may expose rock and create habitat, but this is not generally the case since fractured, potentially unstable rock is often removed.

Water Management Water impoundments could inundate active mine or cave roost sites, potentially trapping and killing entire colonies, and eliminating the structure altogether. Tree roosts can be similarly affected. Foraging areas in the riparian zones and on lower slopes would be removed.

Canyons which offer suitable rock crevices for fringed-tailed myotis also provide basins for storage reservoirs, and other water projects. Almost every river which drains the west side of the Sierra Nevada range in California has one or more such reservoirs. It is almost certain that both roosting and foraging habitat has been lost for this species as a result of these projects. For example, the filling of Hetch Hetchy Reservoir inundated riparian habitat and lower cliff faces of the valley.

Encroachment of non-native species (e.g. tamarisk) can replace native flora. These invasive species seem to be most prevalent in riparian habitats. Land owners and managers all differ in their aggressiveness (or the lack thereof) in treating these non-natives. It is unknown if fringed-tailed myotis would utilize non- native vegetation for foraging.

Small springs, ponds, and artificial water sources are often used for drinking.

Collecting Practices Although the species is protected from over-collection under current Department permitting practices, there is no doubt that scientific collection contributed to or accounted for the extirpation of the colony at Point Reyes National Seashore, and possibly the colony at Old Fort Tejon. While these museum records are invaluable in providing the only historic data we have, historic collecting practices appear to have harmed some populations.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and provided review in the EIS and supplemental documents.

81 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests are proposed for protection of caves and mines to allow for decreased disturbance to the bats, as well as reduced risks of disease transmission.

Determination Statement Inyo, Sequoia, and Sierra NFs: Implementation of the Forest Plan may impact fringe-tailed myotis but will not lead towards Federal listing or a loss of viability. By reducing uncharacteristic wildfires, providing large trees and large snags, and providing protection to caves and mines, the management framework of the Forest Plans should be beneficial to fringe-tailed myotis.

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Baker, J.K. 1962. Notes on the Myotis of the Carlsbad Caverns. Journal of Mammalogy, 43:427-428.

Barbour, R.W. and W.H. Davis. 1969. Bats of America. University of Kentucky Press, Lexington, KY, 286 pp.

Belwood, J.J., and R.J. Waugh. 1991. Bats and mines: abandoned does not always mean empty. Bats 9(3):13-16.

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Brigham, R.M. 1991. Flexibility in foraging and roosting behavior by the big brown bat (Eptesicus fuscus). Canadian Journal of Zoology, 69:117-121.

Burt, W.H. 1934. The mammals of southern Nevada. Trans. San Diego Soc. Nat. Hist., 7:375-428.

Cahalane, V.H. 1939. Mammals of the Chiricahua Mountains, Cochise County, Arizona. Journal of Mammalogy, 20:418-440.

Chung-MacCoubrey, A.L. 1996. Bat species composition and roost use in pinyon-juniper woodlands of New Mexico. Pp. 118-123, in R.M.R. Barclay and M.R. Brigham, Editors. Bats and Forests Symposium, October 19-21, 1995, Victoria, British Columbia, Canada, Research Branch, Ministry of Forests, Victoria, British Columbia, Working Paper 23/1996.

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Clark et al 1978 - Clark, D. R. Jr., et al. 1978. Dieldrin-induced mortality in an endangered species, the gray bat (Myotis grisecens). Science, 199: 1357-1359.

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Cockrum, E.L., and E. Ordway. 1959. Bats of the Chiricahua Mountains, Cochise County, Arizona. Amer. Mus. Novitates, 1938:1-35

Commissaris, L.R. 1961. The Mexican big-eared bat in Arizona. Journal of Mammalogy, 42:61-65.

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Cryan, P.M. 1997. Distribution and roosting habits of bats in the southern Black Hills, South Dakota. M.S. Thesis, Univ. of NM, Albuquerque, NM. 98 pp.

Dalquest, W.W. 1947. Notes on the natural history of the bat Corynorhinus rafinesquii in California. Journal of Mammalogy, 28(1):17-30.

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Easterla, D.A. 1973. Ecology of the 18 species of Chiroptera at Big Bend National Park, Texas. Northwest Missouri State Univ. Studies, 34:1-165.

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Hall, E.R. 1981. The mammals of North America. Vol 1. John Wiley and Sons, Inc., New York. 600 pp.

Hirshfeld, J.R., and M.J. O’Farrell. 1976. Comparisons of differential warming rates and tissue temperatures in some species of desert bats. Comp. Biochem. Physiol., 55A:83-87.

Hoffmeister, D.F., and W.W. Goodpaster. 1954. The mammals of the Huachuca Mountains, southeastern Arizona. Illinois Biol. Monogr., 24:1-152.

Jones, C. 1965. Ecological distribution and activity periods of bats of the Mogollon Mountains area of New Mexico and adjacent Arizona. Tulane Studies Zool., 12:93-100.

Jones, C. and R.D. Suttkus. 1972. Notes on netting bats for eleven years in western New Mexico. Southwestern Naturalist 16(3/4):261-266.

Judd, F.W. 1967. Notes on some mammals from Big Bend National Park. The Southwestern Naturalist, 12(2):192-194.

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Kunz, T.H. 1982. Roosting ecology of bats in T.H. Kunz, editor, Ecology of bats. Plenum Press, New York, USA. Pages 1-55.

Kunz, T.H., and L.F. Lumsden. 2003. Ecology of cavity and foliage roosting bats in T.H. Kunz and M.B. Fenton, editors, Bat Ecology. University of Chicago Press, Chicago, USA. Pages 3-89.

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Mannings, R.W. and J.K. Jones. 1988. A new subspecies of fringed myotis, Myotis thysanodes, from the northwestern coast of the United States. Occas. Pap. Mus. Texas Tech Univ. 123:1-6.

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Miner, K. and P. Brown. 1996. A report on the southern California forest bat survey and radio-telemetry study of 1996. Contract Report for USDA Forest Service, San Diego, CA, 13 pp.

Miner, K., P. Brown, R. Berry, C. Brown-Buescher, A. Kisner, S. Remington, D. Simons, D. Stokes, J. Stephenson, and L. Underwood. 1996. Habitat use by Myotis evotis and M. thysanodes in a southern California pine-oak woodland. Bat Research News, 37(4):141.

Musser, G.G., and S.D. Durrant. 1960. Notes on Myotis thysanodes in Utah. Journal of Mammalogy, 41(3):393-394.

O’Farrell, M.J., and E.H. Studier. 1973. Reproduction, growth, and development in Myotis thysanodes and M. lucifugus (Chiroptera: Vespertilionidae). Ecology, 54(1):18-30.

O’Farrell, M.J., and E.H. Studier. 1975. Population structure and emergence activity patterns in Myotis thysanodes and M. lucifugus (Chiroptera: Vespertilionidae) in northeastern New Mexico. Amer. Midland Nat., 93:368-376.

O’Farrell, M.J. and E.H. Studier. 1980. Myotis thysanodes. American Society of Mammalogists, Mammalian Species, 137:1-5.

Orr, R.T. 1956. The distribution of Myotis thysanodes in California. Journal of Mammalogy general notes, 37(4):545-546.

Pierson, E.D. 1989. Help for Townsend’s big-eared bats (Plecotus townsendii) in California. Bats 7(1):5- 9.

Pierson, E.D., W.E. Rainey, and D.M. Koontz. 1991. Bats and Mines: experimental mitigation for Townsend’s big-eared bat at the McLaughlin Mine in California. In Proc. V: Issues and technology in the management of impacted wildlife. Thorne Ecological Institute. Boulder, CO.

Pierson, E.D., and P.A. Heady. 1996. Bat surveys of Giant Forest Village and vicinity, Sequoia National Park. Report for National Park Service, Denver Service Center, Denver, CO, 27 pp.

Pierson, E.D. and W.E. Rainey. 1996. The distribution, status and management of Townsend’s big-eared bat (Corynorhinus townsendii) in California. California Department of Fish and Game, Bird and Mammal Conservation Program Report 96-7:1–49

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Pierson, E. D., W. E. Rainey, and L.S. Chow. 2006. Bat use of the giant sequoia groves in Yosemite National Park. Report to Yosemite Fund, San Francisco, CA and Yosemite National Park, El Portal, CA, 154 pp.

Pierson, E.D., W.E. Rainey, and C.J. Corben. 2001. Seasonal patterns of bat distribution along an altitudinal gradient in the Sierra Nevada. Report to California State University at Sacramento Foundation, Yosemite Association, and Yosemite Fund, 70 pp.

Rabe, M.J., T.E. Morrell, H. Green, J.C. DeVos, Jr., and C.R. Miller. 1998. Characteristics of ponderosa pine snag roosts used by reproductive bats in northern Arizona. Journal of Wildlife Management, 62:612-621.

Rainey, W.E. and E.D. Pierson. 1996. Cantara spill effects on bat populations of the upper Sacramento River, 1991-1995. Report to California Department of Fish and Game, Redding, CA, (Contract # FG2099R1). 98 pp.

Rasheed et al 1995 –Rasheed, S.A., P.F.J. Garcia, and S.L. Holroyd. 1995. Status of the fringed myotis in British Columbia. B.C. Min. Environ., Lands and Parks, Wildl. Br. Victoria, B.C. Wildl. Work. Rep. No. WR-73.

Rasheed, S.A. and S.L. Holroyd. 1995. Roosting habitat assessment and inventory of bats in the MICA Wildlife Compensation Area. Report prepared for B.C. Hydro, B.C. Min. Environment, Lands and Parks, and Parks, Canada. Pandion Ecological Research Ltd. 77 pp.

Riddle 1995 -Riddle, B. R., ed. 1995. Inactive Mines as Bat Habitat: Guidelines for Research, Survey, Monitoring and Mine Management in Nevada. Proceedings. Jan. 21-22, 1994. Biological Resources Research Center, University of Nevada, Reno. 148 pp.

Roest, A.I. 1951. Mammals of the Oregon Caves area, Josephine County. Journal of Mammalogy, 32:345- 351.

Studier, E.H. 1968. Fringe-tailed bat in northeast New Mexico. The Southwestern Naturalist, 13(3):362.

Studier, E.H., and M.J. O’Farrell. 1972. Biology of Myotis thysanodes and M. lucifugus (Chiroptera: Vespertilionidae) – I. Thermoregulation. Comp. Biochem. Physiol., 41A:567-595.

Weller, T.J., and C.J. Zabel. 2001. Characteristics of fringed Myotis day roosts in northern California. Journal of Wildlife Management, 65:489-497.

Whitaker, J.O., Jr., C. Maser, and L.E. Keller. 1977. Food habits of bats of western Oregon. Northwest Science, 51:46-55.

Wilson et al 1991

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Pacific Fisher (Martes pennanti) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account Pekania [Martes] pennanti is the only extant species of the fisher. A member of the family Mustelidae, the fisher is the largest member of the genus Martes, which includes the yellow-throated martens, true martens, and fishers. The fisher is a mammal with a long slender body and short legs. Sexual dimorphism is pronounced with males weighing between 7.7 and 12.1 pounds and females weighing between 4.4 and 5.5 pounds (Powell 1993). Males range in length from 35 to 47 inches and females range from 30 to 37 inches (Powell 1993). Based on an examination of several skins, Grinnell et al. (1937) noted that fishers from the Sierra Nevada tend to be paler in color than fishers from other parts of the United States.

The State of California designated fisher as a protected furbearer in 1946 and as a Species of Concern since at least 1986 (http://www.dfg.ca.gov/wildlife/species/publications/mammal_ssc.html#a64). In March 2009, the California Fish and Game Commission recommended that the fisher be assessed for listing as threatened or endangered under the California State Endangered Species Act. This recommendation initiated a 12-month status review by the California Department of Fish and Wildlife culminating in a determination by the Commission on June 23, 2010, that the listing was not warranted (CDFG, 2010). In 2013, the California Fish and Game Commission set aside its prior decision and made fisher a candidate species pending a new status review.

The fisher has been included on this Forest Service Sensitive Species list since 1984 (MacFarlane, 1994).

Distribution and Status in North America, the West Coast, and California Fishers range from Quebec, the Maritime Provinces, and New England west across boreal Canada to southeastern Alaska, south in the western mountains to Utah, Wyoming, Idaho, and California, and formerly south to Illinois, Indiana, Tennessee, and North Carolina. Most populations in the Rocky Mountains are the result of reintroductions. Reintroductions have led to fisher reoccupation of former habitats in Idaho, Wisconsin, Michigan, Montana, Nova Scotia, Vermont, West Virginia, Virginia, Maine, Manitoba, Minnesota, Ontario, Oregon, Tennessee, Connecticut, the Hudson Valley in New York, and New Jersey. In recent years, fisher have spread from Vermont into southern New Hampshire, Massachusetts, and Rhode Island including making inroads into suburban backyards, farmland, and even semi-urban areas in Michigan and Pennsylvania as well as Ontario and Quebec in Canada. Fishers were also reintroduced on the Olympic Peninsula in Washington in 2008. The species is relatively abundant in the eastern provinces of Canada, with low populations in British Columbia (USFWS, Federal Register, 1 March 1996, NatureServe 2011).

Fisher populations are presently at low numbers or absent throughout most of their historic range in Montana, Idaho, Washington, Oregon, and California (Heinemeyer et al. 1994). In recent decades, a scarcity of sightings in Washington, Oregon, and the northern Sierra Nevada may indicate fisher extirpation from much of this area (Aubry et al. 1999, Carroll et al. 1999, Zielinski et al. 1995). The Sierra Nevada and northwestern California and Southwest Oregon populations appear to be the only naturally-occurring, known breeding populations of fishers in the Pacific region from southern British Columbia to California (Zielinski et al. 1997).

Grinnell et al. (1937) described the distribution of fishers in California as a continuous arc from the northern Coast Range eastward to the southern Cascades, and then south through the western slope of the Sierra Nevada, but did not attempt to estimate population numbers. In the Sierra Nevada, the fisher

86 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests historically occurred in the Lassen, Plumas, Tahoe, Lake Tahoe Basin, Eldorado, Stanislaus, Sierra, Sequoia, and a small portion of the southern Inyo National Forests, but was not known to occur in the Modoc or Humboldt-Toiyabe National Forests. As of 1995, Zielinski et al. (1995) determined that fishers remain extant in just two areas comprising less than half of the historic distribution: northwestern California and the southern Sierra Nevada from Yosemite National Park southward, separated by a distance of approximately 260 miles.

Fishers are long-lived, have low reproductive rates, large home ranges (for carnivores of their size) and exist in low densities throughout their range (Powell 1993). This implies that fishers are highly prone to localized extirpation, colonizing ability is somewhat limited, and that populations are slow to recover from deleterious impacts. Isolated populations are therefore unlikely to persist.

Sierra Nevada Population Status and Trend Based on extensive track plate and camera surveys (1997 to present) in the Region 5 Status and Trend Monitoring Program and the systematic surveys coordinated by Bill Zielinski from 1996 to 2002, the following observations can be made about the population (Rick Truex, USFS, pers. comm. 2006). Fisher currently appear to be limited in distribution from approximately the southern extent of the Sierra Nevada in Kern County (Greenhorn Mountains and Kern Plateau) to Yosemite National Park. Fishers appear to be absent from the Stanislaus NF, and the northern extent of the population in Yosemite National Park is not well defined. It appears fishers do not occur north of State Highway 120 in Yosemite National Park. Within the southern Sierra population, fishers occur on the west slope of Sierra and Sequoia NFs as well as on the Kern Plateau portion of Sequoia NF (and southernmost Inyo NF). Patterns of detection within the southern Sierra Nevada fisher population suggest the following:

9. Fisher are well distributed on the west-slope Sequoia NF, from the Kings River south through the Greenhorn Mountains. Annual rates of occupancy (i.e., proportion of sites sampled that detected fisher) are generally consistent, and the spatial distribution of detections is more consistent from year to year than elsewhere in the southern Sierra. This area has been consistently occupied since surveys began in earnest during the early 1990s. 10. Recently the detection rate of fisher on the Sierra NF is roughly half what it is on the Sequoia NF. Fisher may have increased their spatial distribution on Sierra NF since the mid-1990s. The annual occupancy rate within Sierra NF seems to be consistent, though the spatial pattern of detections appears more variable among years than on the Sequoia NF. Mark-recapture data collected over the last several years estimate the density of fisher in the Kings River Project area at approximately 1 per 2,500 acres (Mark Jordan, University of California, pers. comm. 2006). Status and trend monitoring for fisher and American marten was initiated in 2002 as part of the Sierra Nevada Forest Plan Amendment Final EIS; the monitoring objective is to be able to detect a 20 percent decline in population occupancy (USDA-FS 2006a). This monitoring includes intensive sampling to detect population trends on the Sierra and Sequoia national forests, where fisher currently occur, and is supplemented by less intensive sampling in suitable habitat in the central and northern Sierra Nevada specifically designed to detect population expansion. From 2002 to 2014, 456 sites were surveyed throughout the Sierra Nevada on 1,861 sampling occasions, with the bulk of the sampling effort occurring within the Southern Sierra fisher population monitoring study area (USDA-FS 2014).

In the southern Sierra Nevada (south of the Merced River in Yosemite National Park) from 2002 thru 2014, over 1500 primary sample units were completed, consisting of 8,848 individual survey stations sampled for over 140,000 survey nights (USDA-FS 2006a) (J. Tucker pers. comm.). In the twelve southern Sierra Nevada monitoring seasons to date (2002 to 2009 and 2011 to 2014), fishers were detected at a total of 132 of 275 sample units within the southern Sierra Nevada, or 48 percent of sites

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(USDA-FS 2014). Preliminary proportions of number of sample sites with fisher detections divided by the number of sites surveyed are presented in Table 6. Using future data, the proportions will be adjusted based upon ability to detect fisher, potentially resulting in higher annual estimates than those reported here. Annual estimates will be used to monitor trend (USDA-FS 2006a, Zielinski et al. 2013).

Table 6. Proportion of sites occupied in the Sequoia and Sierra National Forests Sequoia NF Year Sequoia Kern Plateau* Sierra NF Entire Area West Slope 2002 0.54 0.11 0.19 0.27 2003 0.52 0.13 0.18 0.25 2004 0.41 0.23 0.16 0.22 2005 0.45 0.26 0.16 0.24 2006 0.64 0.19 0.21 0.31 2007 0.60 0.23 0.18 0.28 2008 0.43 0.14 0.21 0.25 2009+ 0.57 0.46 0.16 0.25 2011 0.50 0.29 0.33 0.36 2012 0.57 0.22 0.18 0.27 2013 0.55 0.15 0.19 0.27 2014 0.54 0.34 0.27 0.35 (Updated 3/11/2010) *USDA Forest Service 2009, Truex et al. 2009, Truex, pers. comm. 2010. Geographic areas are defined as Sequoia NF West Slope (including Hume Lake Ranger District), Sequoia Kern Plateau (the Kern Plateau portion of Sequoia National Forest), and Sierra (Sierra National Forest). Habitat availability and detection rates on the Kern Plateau may be affected by habitat loss due to large fires. In 2007 the SQF West Slope sampling included one unit in Sequoia National Park, and the Sierra NF included six units in Yosemite National Park. + Sampling effort during 2009 was reduced on the Kern Plateau due to safety and operational considerations. Sampling was limited to the northern portion of the plateau and the observed occupancy is likely higher than it would otherwise have been if sampling had occurred throughout the area as in previous years (Truex, pers. comm.). Sampling effort in 2014 included 13 new units not previously surveyed which may have increased occupancy estimates for this year.

Zielinski (2004a) noted that a comparison of current vs. historic distribution is an essential step to assess the status of any wildlife population. Fishers appear to be absent from a 260 mile long expanse of their historic range in the northern and central Sierra Nevada (Zielinski et al. 1995). This gap is likely a result of negatively synergistic interactions resulting from historic patterns of logging, trapping, and porcupine (Erethizon dorsatum) poisoning (Zielinski 2004a); deeper snow levels in the northern Sierra Nevada (Krohn et al. 1997); higher densities of roads and human developments in the northern Sierra Nevada (Duane 1996); and the current distributions of other generalist carnivores (Campbell 2004). The current disjunct distribution pattern may also be partially attributed to movement and dispersal constraints imposed by the elongated and peninsular distribution of montane forests in the Pacific states (Wisely et al. 2004). Several large gaps in this narrow band of suitable habitat have been created by large-scale, stand- replacing wildfire. Given the apparent reluctance of fishers to cross open areas (Earle 1978), and the more limited mobility of terrestrial mammals relative to birds, it is more difficult for fishers to locate and occupy distant, but suitable, habitat.

Preliminary results indicated that fishers are well-distributed in portions of the Sequoia and Sierra NFs, with annual occupancy rates consistently higher on the Sequoia (33.3 percent to 41.1 percent) than the Sierra (14.5 percent to 22.7 percent) (USDA-FS 2005). Fishers have not been detected in the northern, central or eastern Sierra Nevada Mountains. Comparisons to southern Sierra Nevada survey data from the 1990’s suggest that the area of occurrence for fisher may have expanded during the past 10 years (USDA-

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FS 2005). Additionally, seven years of monitoring results (Table 6) suggest that there has been no conspicuous difference in occupancy rates among years, and no seasonal effects on detection probabilities within the June to October sampling periods (Truex et al. 2009). Some variability in detection rates occur among the years for the three geographic areas (Table 6). Fishers are detected most often on the west slope of Sequoia National Forest (USDA-FS 2014).

A recent analysis of the SNFPA Long Term Monitoring data was completed which analyzed a core of 243 sample units from 2002 through 2009 (Zielinski et. al 2013). Findings suggest that over the 8-year period, there was no trend or statistically significant variation in fisher occupancy rates in the southern Sierra populations; however, given the variety of continuing risk factors, continued monitoring is highly favored.

Taxonomy The fisher is one of the larger members of the weasel family (Mustelidae), belonging to the subfamily Mustelinae, and genus Martes. The fisher is the only extant member of the subgenus Pekania and the largest member of the genus Martes (Anderson, 1994). Goldman (1935) found evidence of three subspecies: Martes pennanti pennanti (eastern and central North America), M. pennanti columbiana (Rocky Mountains), and M. pennanti pacifica (west coast of North America). However, Grinnell et al. (1937) found no evidence of subspecies differentiation after examining morphology and pelage characteristics of fisher from Maine, Quebec, Washington, and California. Hagmeier, in Douglas and Strickland (1987), also concluded the subspecies could not be separated on the basis of pelage or skull characteristics. Hall (1981) retained all three subspecies in his compilation of North American mammals, as did Anderson (1994), but neither addressed Hagmeier’s conclusion that the subspecies should not be recognized (Powell 1993). Several authors address genetic variation in fisher populations in their northern and eastern ranges (Williams et al. 1999, Kyle et al. 2001) and in the west (Drew et al. 2003, Aubry et al. 2003, Wisely et al. 2004). These analyses found patterns of population subdivision similar to the earlier described subspecies (Drew et al. 2003). Drew et al. (2003) stated that, “although it is not clear whether Goldman’s (1935) subspecific designations are taxonomically valid; it is clear (based on genetic results) that population subdivision is occurring within the species, especially among populations in the western USA and Canada.”

For the purposes of the 2010 fisher status review, which evaluated the need for protection of the species under the State Endangered Species Act, the California Department of Fish and Game identified fisher, Martes pennanti, as the taxonomic designation for native fishers found in California historically, and at this time until new scientific information is provided (CDFG, 2010). However, The new Conservation of Fishers (Martes pennanti) in South-Central British Columbia, Western Washington, Western Oregon, and California, Volume I: Conservation Assessment (Lofroth, et al. 2010) recognizes the subspecies delineations.

In the NRIS database, the Inyo NF has 1 records, Sequoia NF has 665 records, and the Sierra NF has 336 records. Figure 9 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

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Figure 9. Map of Pacific fisher locations from the NRIS Wildlife Database, 2016

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Habitat Status

General Fisher Habitat and Biogeography for the Sierra Nevada In the Sierra Nevada, fisher habitat occurs in mid-elevation forests (Grinnell et al. 1937, Zielinski et al. 1997) largely on National Forest System lands. The Sierra Nevada status and trend monitoring project (USDA-FS 2006a) has detected fishers as low as 3,110 feet and as high as 9,000 feet in the southern Sierra Nevada, which are considered to be extremes of the elevation range. Mapped female home ranges from the Tule River area were between 3,600 and 7,500 feet in elevation. Males appear to have a much wider range in elevation, 4,000 to 9,300 feet, but also appear to be much less selective in use of habitat in general (Zielinski et al. 2004b). It is expected that this elevation range will vary by latitude and corresponds generally to the lower end of the mixed conifer hardwood cover type at the lower end and the red fir cover type at the upper elevation.

The following California Wildlife Habitat Relationships (CWHR) types are thought to be important to fishers: generally structure classes 4M, 4D, 5M, 5D and 6 (stands with trees 11 inches in diameter at breast height or greater and greater than 40 percent cover) in ponderosa pine, montane hardwood-conifer, Klamath mixed-conifer, Douglas-fir, mixed conifer, montane riparian, aspen, redwood, red fir, Jeffrey pine, lodgepole pine, subalpine conifer, and eastside pine (Timossi 1990). CWHR assigns habitat values according to expert panel ratings. CWHR2 is a derivative of the CWHR fisher habitat relationship model constructed by Davis et al. (Davis et al. 2007). They used best available science to devise a model for predicting fisher occupancy and eliminated some forest types that appeared to contribute little to predicting fisher occupancy although they may be used by fisher. Aspen, eastside pine, lodgepole pine, montane riparian, red fir, and subalpine conifer were eliminated from the CWHR2 fisher model. We have further refined CWHR2 to reflect only those forest types present in the southern Sierra Nevada: Jeffrey pine, montane hardwood-conifer, ponderosa pine, Sierran mixed-conifer and white fir. This southern Sierra Nevada version of the CWHR model is termed CWHR2.1 (Table 7).

Fishers are among the most habitat-specific animals in North America, and changes in quality, quantity and distribution of available habitat can affect fisher distribution in California (Buskirk et al. 1994). The southern Sierra Nevada mountain range provides habitat for the southernmost population of fishers in the world. Despite what appears to be historical isolation from populations to the north, the small southern Sierra fisher population has persisted for many decades (Spencer et al. 2008). Systematic monitoring since the 1990s has provided some evidence of recent population expansion (Truex et al. 2008).

Table 7. CWHR2.1 high and moderate capability habitat for fisher (CWHR 2008 as modified by Davis et al. 2007 [CWHR2] and applied to southern Sierra Nevada forest types [CWHR2.1]) CWHR2.1 High and Moderate Capability Size, Canopy CWHR2.1 Habitats Cover, and Substrate Classes Jeffrey pine 4P, 4M, 4D, 5P, 5M, 5D Montane hardwood-conifer 4P, 4M, 4D, 5S, 5P, 5M, 5D, 6 Ponderosa pine 4P, 4M, 4D, 5P, 5M, 5D Sierran mixed conifer 4P, 4M, 4D, 5S, 5P, 5M, 5D, 6 White fir 4P, 4M, 4D, 5S, 5P, 5M, 5D, 6

Southern Sierra Fisher Studies Substantial information regarding fisher biology in the southern Sierra Nevada was derived from the Tule River Fisher and Marten Study that tracked both fisher and marten on the Sequoia National Forest from 1994 to 1999 (Zielinski et al. 2004a and 2004b, Zielinski and Duncan 2004c, Truex et al. 1998). This area is also referenced as the most southern Sierra Fisher study site.

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The well-studied Kings River Project area is centrally located within the southern Sierra on the Sierra National Forest, which is adjacent to the Sequoia National Forest. Fishers have been studied and monitored within the Kings River Project area since the mid-1990’s (Boroski et al. 2002, Mazzoni 2002, Zielinski et al. 1997, Zielinski et al. 2005, Truex et al. 2008, Jordan 2007, Underwood et al. 2010). More recently, Kathryn Purcell and Craig Thompson from the USFS Pacific Southwest Research Station have initiated a research project on fishers in and around the Kings River Project area. Utilizing a combination of radio collared individuals, fisher scat detecting dogs and remote cameras, they have captured 78 fishers (36 males and 42 females) and collared a total of 72 fishers (33 males and 39 females), have adequate data (at least 25 locations) to delineate home ranges for 70 individuals (31 males and 39 females), and located a total of 93 structures used as dens by reproductive female fishers (42 structures), including 30 natal dens and 63 maternal dens as of August 2010 (Thompson et al. 2011).

The Sierra Nevada Adaptive Management Project (http://snamp.cnr.berkeley.edu/) conducted an intensive investigation into fisher use of habitat and response to management disturbance, largely on the Bass Lake Ranger District of Sierra National Forest. They assessed fisher occupancy in relation to fire history, elevation and canopy cover and evaluated the response of fishers to fuel reduction activities. Fishers used areas with higher canopy cover and occupancy was lower in areas with active recent fire histories (both natural and prescribed). Persistence was lower in areas with more fuels reduction activities. They speculated that fishers would resume the use of treated areas within a few years (Sweitzer et al. 2016).

A recent study by Hanson (2013) examined fisher habitat use throughout a large mixed severity burned landscape located on the Kern Plateau in the Sequoia National Forest. The investigation was conducted for more than 10 years post-fire which allowed for some level of vegetative recovery. In this study, scat detector dogs were used to determine presence of fisher across the burned and unburned landscape. Hanson (2013) found that fisher selected pre-fire mature and old forest that experienced moderate/high- severity fire more than expected based upon availability, just as fishers are selecting dense, mature/old forest in its unburned state. It was further noted that when fishers were near fire perimeters, they strongly selected the burned side of the fire edge (Hanson 2013).

While this study reports valuable evidence of fisher using low severity burned landscape for more than 10 years post fire, further conclusions are limited given the methodology and analysis used to interpret the results. For example, Hanson cites Miller et al. (2009) to define low, moderate, and high fire severity categories. However, the ranges of values used for each fire severity category identified in Miller et al. (2009) were adjusted by Hanson (2013) for his analysis of data. Due to the adjustment of the definitions and the subsequent combining of moderate and higher-severity fire in Hanson’s (2013) analysis, it is difficult to assess the use of moderate and high severity burned landscapes by fisher as defined by Miller et al. (2009) and the Forest Service. It is also problematic to conclude that fisher used pre-fire mature/old forest that experienced moderate/high severity fire more than expected based upon availability when a statistically non-significant result was reported by Hanson (2013) in Table 2a. Spencer et al. (2015) also reviewed the study by Hanson (2013, 2015) and determined it was unclear to what degree fishers preferentially use recently burned areas—especially large, severely burned patches—and it is unlikely that fishers can establish home ranges and obtain all their life requisites (e.g., den sites) within severely burned areas due to diminished canopy cover.

Forest Service policy recognizes the ecological importance of low/moderate mixed severity fire regimes in Sierran mixed conifer forests in that it provides regeneration and habitat for numerous species. Hanson’s (2013 and 2015) studies confirm fisher use of low (and perhaps low-moderate) severity post burn fire areas further supports this policy. But, large scale uncharacteristically severe wildfire poses a risk to fisher denning and resting habitat, as well as habitat connectivity (Lofroth et al. 2010). While Hanson (2013 and 2015) provides a starting point to begin to understand how fishers use post fire

92 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests landscapes as they recover, exactly how the McNally fire and other fires across the southern Sierra Nevada have affected fisher behavior, occupancy, and abundance is unknown and further research is necessary. However, the regional occupancy results indicate that fishers have persisted in a landscape that has experienced a mosaic of low to high severity fires, albeit with the lowest recorded occupancy rates in the assessment area (Zielinski et al. 2013).

An additional research project, utilizing GPS collars to document the immediate response of fishers to fuel treatment actions throughout the southern Sierras was initiated in fall 2009. A portion of this work is being conducted on the Western Divide Ranger District, Sequoia National Forest.

A number of southern Sierra Nevada population estimates and simulations have been conducted, with results converging as presented in Table 8 below.

Table 8. Estimates of the southern Sierra Nevada fisher population occurring across the Sequoia and Sierra National Forests, Mountain Home State Park, tribal lands, Yosemite and Sequoia and Kings Canyon National Parks Estimate of So. Sierra Number of Reproductive Study or Researcher Adult Population Adult Females (Ne) Lamberson et al. (2000) 100 - 500 Not Estimated Spencer et al. (2008) = CBI 160 - 360 50 - 120 Spencer et al. (2008) based on Jordan (2007). 276 - 359 55 - 83 From Spencer et al. (2008) based on Truex (2007) from Southern Sierra Nevada 160 - 250 Not Estimated Monitoring (sampling theory extrapolation) Self et al. (2008) 548 - 598 Not Estimated Spencer et al. (2010) less than 300 Not Estimated

The population estimate in Self et al. (2008) is believed to be higher than the others presented largely due to two factors: the inclusion of juvenile individuals in the estimate (which have high mortality rates and may not survive to become reproductively contributing members of the population), and the habitat basis used for calculating populations that assumed 100 percent occupancy of suitable habitat.

The Conservation Biology Institute (CBI) conducted an assessment of the status of fisher habitat and populations in the southern Sierra Nevada as of 2008. The report evaluated fisher habitat at the landscape, home range and the finer microhabitat scales. Results were provided as maps and digital GIS layers. The CBI report (Spencer et al., 2008) analyzed occupancy data as well as modeled habitat suitability to conclude that the fisher population in the southern Sierra Nevada (ignoring juveniles) is between 160 and 360 total individuals (probably fewer than 300). Spencer et al. (2008) used the PATCH population model (Schumaker 1998) to determine that the southern Sierra Nevada supports a population of between 50 to 120 adult female fishers (i.e. Ne, or the number of females effectively contributing to reproduction).

Other authors (Lamberson et al. 2000, Jordan et al. 2005) have also developed population estimates for the fisher population in the Southern Sierras, including the Kings River Project area. Most recently, Purcell (K. Purcell, pers. comm.) estimated the population for the Kings River Project area alone on the southern Sierra National Forest based on Jordan’s (2007) population density estimates. Purcell estimated that 28 to 36 adult fishers occur in the Kings River Project area, and the ongoing research has collected home range and habitat use data on 70 fishers (39 females) in and around the Kings River Project area.

Spencer et al. (2008) estimated habitat carrying capacity for fishers on the combined suitable areas of the Stanislaus, Sierra, and Sequoia National Forests as mapped by the Landis II model to be between 230 and

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392 adults, of which 73 to 147 individuals are reproductive females (Ne). This is 23 to 27 more effective females than estimated to be present in the population at this time. Most potential habitat on the Sierra and Sequoia National Forests is occupied; this is in stark contrast to the Stanislaus National Forest where no habitat is occupied (Spencer et al. 2008).

Spencer, et al. (2008) modeled habitat suitability at elevations between 3,500 to 8,000 feet based on current and historic occupancy records at the fisher home range and micro-site scales. This should be considered the core of fisher range, despite occasional detections above or below these elevations. Outputs reflected habitat suitability based on the probability of occurrence. The CBI predicted probability of occupancy map (Spencer et al. 2008) serves as a “best available science” refinement of the crude elevation-based Southern Sierra Fisher Conservation Area designated in the 2001 and 2004 Sierra Nevada Framework analyses and decisions (USDA-FS 2004a). Data utilized in this suitability analysis were restricted to locations where fishers were detected in more than one year to provide a more accurate estimate of occupancy, and avoid overemphasis on what might be transient individuals.

The CBI Phase II report: Baseline Evaluation of Fisher Habitat, Fires and Vegetation Dynamics in the Southern Sierra Nevada (Spencer et al. 2008) linked and fine-tuned the Phase I habitat and population models with a vegetation dynamics model, LANDIS-II, simulating predicted changes in forest vegetation in response to wildfires, management actions, climate change, and ecological succession. Results of this modeling exercise are discussed in the Effects Section of this document. They generally concluded that the threat of uncharacteristically severe wildfire outweighs the threat of short term declines in habitat suitability, with specific caveats disclosed in the report.

Local status Surveys for fisher within the Tobias project area have been conducted through a variety of efforts. There are sample units for the Southern Sierra Nevada Fisher and Marten Status and Trend Monitoring Project in the vicinity (USDA-FS 2014). This project conducts systematic surveys across the National Forests of the Sierra Nevada to track the status and trend of carnivore populations, specifically fisher and marten (Martes americana). There have been numerous fisher detections both within the Tobias Project area and in adjacent areas (Figure 9). NRIS Wildlife documents fisher detections from 1991 to 2009 in this area.

Reproduction, Recruitment, and Survival Fishers have relatively low reproductive rates, with birth occurring in late March to early April. Conservation of reproductive sites (generally greater than 41 inches diameter at breast height live or dead white fir or black oak in greater than 90 percent canopy closure - see description of Sequoia National Forest den sites selected in Table 9) is essential to prevent degradation of habitats used by females during the reproductive period.

Caution must be exercised in interpreting measurements of canopy cover. Measurements from spherical densiometers, which were used in most pre-2002 studies, differ significantly (are much higher) than measures used in project level analysis by the Forest Service, which are generally based on aerial photo interpretation or satellite imagery. Based on moosehorn readings taken at the 14 dens found on the Kings River Project in 2008, the average canopy cover was 73 percent (range 56 to 93 percent), which converts to 66 percent (range 52 to 76) based on Forest Inventory and Analysis non-overlapping points (K. Purcell, pers. comm.). Therefore the 90 percent canopy closure figure in the previous paragraph may be easily misapplied.

Selection of natal (birthing) and maternal (kit raising) dens is highly specific. Reproductive, thermal, and escape cover must exist in the proper juxtaposition within specific habitats in order to provide a secure environment for birth and rearing of fisher kits. All known natal and maternal dens in the western United

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States have been in large diameter logs, snags, or cavities of large diameter live conifers or oaks (Powell et al. 1994, Zielinski et al. 1995, Truex et al. 1998). Natal dens in the Southern Sierra Nevada have not occurred in downed woody material of any diameter (C. Thompson pers. comm., R. Sweitzer, unpublished data).

The current Kings River Project fisher study reported a 70 percent rate of reproduction for adult females (7 of 10) in 2007, 91 percent (10 of 11 adult females) in 2008, 75 percent in 2009 (12 of 16 females), and 81 percent in 2010 (13 of 16 females), with an average litter size of 1.7 kits (n = 45); (Thompson et al. 2011). Thirty natal and 63 maternal den trees have been located to date. Forty-nine percent of dens (n = 46) were found in California black oak, with all but two living. Conifer species used included white fir (12 live trees and 12 snags), incense cedar (8 live trees, 7 snags and 1 log), ponderosa pine (3 live trees), and sugar pine (4 snags) (Thompson et al. 2011). This ongoing study as well as the concurrent Sierra Nevada Adaptive Management Project is anticipated to provide the only information ever recorded for fisher population recruitment in California.

The Sierra Nevada Adaptive Management Project monitored fisher populations on a portion of Sierra National Forest from 2007 to 2013. Denning rates were estimated at 84 percent, weaning rates at 70 percent and average litter size was 1.6 kits Sweitzer, et al. 2015). Survival was lowest in the spring to mid-summer time period; the overall survival rate for females was 72 percent and 62 percent for males (Sweitzer, et al. in press).

The upper limit of life expectancy for fishers in the wild was generally thought to be about 10 years (Powell 1993) until a 12-year-old individual was wild-trapped in British Columbia (Weir 2003). On the Kings River Project fisher study, 8 females previously marked by M. Jordan for his thesis were captured (Purcell et al. 2009). One died at a minimum of 11 years (exact age pending) and had reproduced in her final year. A second is still alive at a minimum of 10 years and reproduced in both 2006 and 2008. The others currently range from 6 to 8 years old and most are currently reproducing. We have used 10 years to be generally representative of a fisher generation in this analysis.

Diseases and Parasites All wild populations are infected to some extent by diseases and parasites. The potential negative impact of disease increases as populations become small and isolated (Brown et al. 2008). Blood samples show evidence of past exposure but the animal is not necessarily infected and cannot transmit the virus; scat samples show active infections where the disease can be transmitted (Purcell et al. 2009). Viruses and bacteria associated with fishers or related mustelids are summarized in Brown et al. (Brown et al. 2008) and include: canine distemper, parvoviruses, influenza, corona viruses, and canine adenovirus (the cause of canine infectious hepatitis), West Nile virus, Brucella spp. (the cause of brucellosis), Yersinia pestis (the cause of the plague), and Leptospira interogans (the cause of leptospirosis). Thompson et al. (2011) found 8 (5 males and 3 females) of 51 (25 males and 26 females) fishers captured and tested in the Kings River Project area showed evidence of past exposure to parvovirus, 3 (1 male and 2 females) showed evidence of past exposure to canine adenovirus, 5 (2 males and 3 females) showed evidence of past exposure to canine distemper, and one (male) animal was actively infected with parvovirus (M. Gabriel, pers. comm.). The Sierra Nevada Adaptive Management Project study documented six disease mortalities from 2007 to 2011, with canine distemper, canine parvovirus, bacteria, and an unknown nematode responsible for the deaths (derived from http://snamp.cnr.berkeley.edu/static/documents/2011/07/23/Sweitzer_FisherIT_July19_2011_PostPart2.pd f as accessed July 25, 2011).

Domestic dogs may act as infectious agents to transmit some of the above pathogens. The parvovirus strain found in southern Sierra fishers is identical to that typically carried by domestic dogs (M. Gabriel,

95 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests pers. comm.). Some infectious agents are shed in the saliva, nasal discharge, or feces of infected animals, and persist in the environment, creating a contamination risk for equipment used for the trapping and handling of fishers leading to infection of subsequent captures (Brown et al. 2008). Negatively synergistic effects may occur when fishers are infected by one or more pathogens, lowering general fitness, increasing risk of predation, or decreasing reproductive capability.

Mortality Truex et al. (1998) reported that mortality rates of adult female fishers in the southern Sierra population appeared to be high. The Kings River Project Fisher Project on the Sierra National Forest has confirmed 27 mortalities (14 males and 13 females) since the inception of the project (Thompson et al. 2009). Twenty-two of the mortalities (81 percent) can be attributed to predation, with bobcats and mountain lion as the main predators (Thompson et al. 2011). Mortality occurs year round with a peak for males occurring in late winter or early spring, and an increase for females in summer and fall (Thompson et al. 2011).

Thru June 2011, Sierra Nevada Adaptive Management Project has captured and radio collared 82 individual fishers (34 males and 48 females), and documented causes of mortality for 48 fishers (39 collared and 9 noncollared). The major causes of mortality include predation, disease and roadkill distributed as follows: 19 (+5 pending) from predation (bobcats, mountain lions, coyotes), 6 disease (canine distemper virus, bacterial, CPV, unknown nematode), 3 (+5 noncollared animals in Yosemite National Park) roadkills, 3 indeterminate, 2 from starvation and one each from rodenticide and drowning (derived from http://snamp.cnr.berkeley.edu/static/documents/2011/07/23/Sweitzer_FisherIT_July19_2011_PostPart2.pd f as accessed July 25, 2011. Sierra Nevada Adaptive Management Project has also documented rodenticide as an emerging issue with 21 of 24 livers or 88 percent testing positive for exposure to the poison.

The potential of direct and indirect exposures and illicit use of anticoagulant rodenticides on public forest lands have recently raised concern for fishers. Gabriel et al. (2012) found 46 of 58 (79 percent) fishers exposed to an anticoagulant rodenticide with 96 percent of those individuals having been exposed to one or more second-generation anticoagulant rodenticide compounds. Additionally, Gabriel et al. (2012) diagnosed four fisher deaths, including a lactating female, that were directly attributed to anticoagulant rodenticide toxicosis and documented the first neonatal or milk transfer of an anticoagulant rodenticide to an altricial fisher kit. The spatial distribution of exposure suggests that anticoagulant rodenticide contamination is widespread within the fisher’s range in California and points to illegal marijuana cultivation as a likely point source.

On the Hoopa Reservation in northwest California, fisher predation mortality is high, and primarily due to bobcats (Felis rufus), although coyotes (Canis latrans) and mountain lions (Felis concolor) also take fishers (Higley et al. 1998). One might speculate that habitat alterations favoring bobcats, mountain lions or coyotes could increase fisher mortalities (D. Macfarlane, pers. comm.). Bobcats find optimal habitat in brushy stages of low and mid-elevation conifer, oak, riparian, and pinyon-juniper forests, and all stages of chaparral (http://www.dfg.ca.gov/biogeodata/cwhr/cawildlife.aspx). Coyotes frequent open brush, scrub, shrub, and herbaceous habitats, and are also found in younger stands of deciduous and conifer forest and woodland with low to intermediate canopy, and shrub and grass understory (http://www.dfg.ca.gov/biogeodata/cwhr/cawildlife.aspx). Mountain lions are most abundant in riparian areas, and brushy stages of most habitats (http://www.dfg.ca.gov/biogeodata/cwhr/cawildlife.aspx).

In the Kings River Project study area, (Thompson et al. 2011) reported four mortalities due to confirmed mountain lion predation, four confirmed from bobcat predation, two confirmed coyote predations, one

96 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests entombment, one vehicle collision, one drowning (on the Sierra Nevada Adaptive Management Project), one disease, one died during capture (female was severely emaciated and tested positive for canine distemper virus) and twelve unknown predation incidents (genetic confirmation of predator species is pending). Fishers are innately curious and tend to explore small openings. Multiple deaths have been documented from fisher entering cisterns, pipes, tubes, open water tanks and other dead ends from which they cannot escape. These structures have been associated with recreation residences or summer homes, forest in-holdings and at a timber operator’s staging area.

Spencer et al. (2010) concluded that high mortality levels are likely interfering with the natural re- establishment of a breeding fisher population north of Yosemite Valley. In this area habitat generally declines in suitability and becomes more fragmented (Spencer et al. 2008).

Food Habits Food habit studies by Grenfell et al. (1979) in northwestern California and Zielinski et al. (1999) in the southern Sierra Nevada show a wide diversity of prey. Common prey in both studies were California ground squirrel (Spermophilus beecheyi); western gray squirrel (Sciurus griseus); and Douglas squirrel (Tamiasciurus douglasii); deer mouse (Peromyscus spp.); harvest mouse (Reithrodontomys megalotis); voles (Microtus spp.); deer (Odocoileus hemionus); carrion; beetles; social wasps; and false truffles (Rhizopogan spp.). Southern Sierra fishers also fed on alligator lizards (Elgaria spp.), yellow jackets (Vespula vulgaris) and berries (Ribes spp., Arctostaphylos spp.) (Zielinski et al. 1999, Thompson et al. 2011), indicating that this most southern of fisher populations was exploiting a variety of food as well as relatively small prey species. Golightly et al. (2006) found that reptiles were seasonally important in their northern California study area. Some prey species (e.g., Peromyscus, Scapanus, Thomomys) frequent early-seral stages of coniferous forests, while chaparral provides optimal habitat for some reptiles, lagomorphs, deer mice, and deer (Verner et al. 1980) as well as shrub species that produce fruits consumed by fishers (Zielinski et al. 1999).

California fishers respond to changing prey densities (Kuehn 1989) by shifting their seasonal diets. Increased use of mice and carrion in the winter, and berries in the fall probably compensate for the unavailability of hibernating squirrels and reptiles (Zielinski et al. 1999), and reflects a generally opportunistic strategy. Further, the diversity of prey in the diet of southern Sierra fishers may also reflect the diversity of habitats (and niches) in this region. Many of the prey species found in the diet of fishers occur primarily in large tree and dense canopy coniferous forests and oak woodland habitats, while others prefer chaparral and deciduous riparian areas (Zielinski et al. 1999). Oak, especially, may be important as sources of mast that may stimulate higher prey densities (Powell et al. 1994). In general, as compared to more northern fisher ranges, the drier southern Sierra Nevada provides a diversity of habitats and prey species, following a recognized pattern of increased species richness with decreased latitude (Pianka 1966). This pattern may also manifest in the diverse diet of the southern Sierra population of fisher, especially with the emergence of reptiles, insects, false truffles, and berries as important food items.

Zielinski et al. (1999) note that prey habitat provides limited inferences about the habitat of fishers because fishers spend much of their time in non-foraging habitat like resting sites. Zielinski and Duncan (Zielinski et al. 2004c) compared the diets of fishers and martens (Martes americana) where they co- occur in the southern Sierra Nevada. Both species are at the southernmost extremes of their geographic range in this area. Populations occurring at the margin of their range usually exist in environmental conditions differing from those in the heart of the range, including prey availability and foraging conditions (Hoffman et al. 1994). Zielinski and Duncan (Zielinski et al. 2004c) confirm this to be the case, and document diverse and similar diets for both species in this area. They offer two explanations for this phenomenon: the relative absence of larger prey species like leporids (rabbit family) plus an

97 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests alternative theory they prefer, that fishers conduct opportunistic exploitation of the high variety of foods available for consumption.

Territoriality and Home Range Fishers were thought to exhibit intrasexual territoriality, where individuals defend a home range against members of the same sex, but there is considerable overlap between sexes (Johnson et al. 2000). These territories are maintained year-round except during the breeding season when males trespass on each other’s territories while they search for receptive females (Powell 1993). However, intriguingly, initial results indicate high intrasexual territory overlap in the Kings River area (Mazzoni 2002, Purcell et al. 2009). This constitutes a departure from traditional thinking on fisher intrasexual home ranges.

Table 9. Average fisher home range sizes in the Sequoia National Forest MEAN MALE MEAN FEMALE Source Home Range Home Range (acres) (acres) Zielinski et al. (1997) 9,855a 1,644a Zielinski et al. (2004b) 7,409d 1,304b Arithmetic Mean 8,632 1,474 a Mean of two home range estimating techniques: 95 percent minimum convex polygon, and adaptive kernel. b 100 percent Minimum Convex Polygon method

Table 10. Average fisher home range sizes in the Sierra National Forest MEAN MALE MEAN FEMALE Source Home Range Home Range (acres) (acres) Thompson et al. (2011)a 6,511 2,708 Mazzoni (2002)b 5,421 2,945 Sweitzer (2011)c 23,524 5,659 Arithmetic Mean 11,819 3,771 a 95 percent fixed kernel estimates based on 14 male and 46 female territories. b 95 percent Minimum convex polygon estimate c 95 percent fixed kernel estimates based on 17 male and 30 female territories.

The home range acreages in Table 9 and Table 10 are limited by the differences in home range size calculation techniques, which vary from very restrictive (likely underestimation of size) to very liberal (probable overestimation of size). Sierra Nevada Adaptive Management Project data (from http://snamp.cnr.berkeley.edu/static/documents/2011/07/23/Sweitzer_FisherIT_July19_2011_PostPart2.pd f as accessed July 25, 2011) from the northern Sierra National Forest have found that the large home range sizes on the Sierra Nevada Adaptive Management Project study site compared to the other sites in the southern Sierra were due to the different methods used to collect the data. R. Sweitzer (pers. comm.) hypothesizes that ground based telemetry may underestimate home ranges due to the difficulty of regularly locating animals in remote rugged habitat.

Movements – Daily, Breeding, and Dispersal Fishers may be active day or night, but peaks occur at sunrise and sunset (Powell 1993). Assuming two activity periods per day, Powell (Powell 1993) estimated fishers may move up to 3.1 miles per day. In southern Oregon, breeding movements by males began one to two months prior to the breeding season of

98 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests late February to the end of April (Douglas et al. 1987) and terminated at their non-breeding home range by the end of April (Aubry et al. 2006).

Dispersal has profound implications to mammalian population structure, affecting the ability to colonize vacant habitat (Bowman et al. 2001), dispersion (Shaw 1995), spacing patterns (Krebs et al. 1969), allelic frequencies (Landry et al. 1999), demographics (Krohne et al. 1999), and extinction thresholds (Fahrig, 2001). Bowman et al. (2002) demonstrated that dispersal distance of mammals after translocation was more closely related to the size of their home range than body size. Weir (2003) summarized information about fisher dispersal movements and concluded that they were capable of long distance movements, and that neither major rivers nor other topographic features appear to pose impenetrable barriers. Recent anecdotal evidence shows that rivers may serve as filters, however (K. Purcell, pers. comm.) other considerations such as availability of suitable habitat and prey, avoidance of mortality factors like predators, and the presence of other fisher (which seems to serve as an attractant), may interact to influence dispersal (Weir 2003). In southern Oregon, juvenile female fishers dispersed from 0 to 10.6 miles, with a mean of 3.7 miles, while males in the same area ranged from 4.4 to 34.2 miles, with a mean of 18.0 miles (Aubry et al. 2006). Male fishers on the Sierra Nevada Adaptive Management Project study area had a documented mean dispersal distance of 12.3 kilometers give or take 2.4 kilometers and mean female dispersal was 7.5 kilometers give or take 1.4 kilometers.

Population Genetics Because the southern Sierra fisher sub population appears to have adapted to use of more open and drier habitats than its relatives in the more northern temperate rain forests, it has been theorized that the maintenance of the southern Sierra Nevada fisher population may be critical to conserving fisher populations in the western United States (Zielinski 2004a). Several studies have revealed genetic patterns that appear to arise from the disjunct nature of fisher population distributions in the Pacific States, and point to reduced genetic diversity in the southern Sierra Nevada population (Drew et al. 2003, Wisely et al. 2004, Knaus et al. 2011, Tucker et al. 2012). Wisely et al. (2004) analyzed 33 fisher genetic samples from three different locations to investigate the role of landscape features in genetics in the southern Sierra Nevada. The study concluded that fisher expansion southward into the west coast mountain chains occurred less than 5,000 years ago, leading to reduced genetic diversity and increased population structure at the southern periphery of its range. This study suggested that dispersal was limited and thus indicated that aggressive conservation strategies may be needed to preserve the existing southern Sierra Nevada fisher sub population and reconnect extant populations to the north. Consistent with this genetic analysis, the Kings River was postulated to constitute a major barrier to gene flow, perhaps permeable to just one migrant every 50 generations (Wisely et al. 2004). The number of migrants needed per generation to maintain genetic viability is highly dependent on the specific demographic and genetics characteristics of the population (Mills et al. 1996, Vucetich et al. 2000). The results reported by Wisely et al. (2004) were cause for concern.

More recently, fisher DNA samples from 127 individuals from a broad distribution across the entire southern Sierra fisher sub population have been analyzed as part of an on-going doctoral dissertation. A progress report on this work (Tucker et al. 2009, Tucker 2009 in CDFG 2010, Truex et al. 2009) indicated much higher levels of population connectivity in the southern Sierra Nevada. A cluster analysis using the program GENELAND (Guillot et al. 2005) signaled the presence of three intermixing population groupings: one in the of the Sierra National Forest north of the Kings River, another encompassing the Hume Lake District of the Sequoia National Forest and Sequoia/Kings Canyon National Park, and a southern third on the Sequoia National Forest (Tucker et al. 2009, Tucker 2009 in CDFG 2010). Preliminary data indicate that at least one individual per generation moves from the northwest Sierra to the central population group, and up to 3.5 individuals per generation are interchanged between the central and southern genetic group (Tucker et al. 2009, Tucker 2009 in CDFG 2010, Truex et al. 2009).

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Based on this preliminary information, the Kings River does not appear to constitute a significant barrier to fisher movement, as hypothesized in Wisely et al. (Wisely et al. 2004). It should be emphasized that Tucker’s work is ongoing and the results and interpretations may change in the continuing process. However, the results are based on a much larger and better distributed dataset than the previously published information and appear to be the best available and most current data.

Another study reported to the California Department of Fish and Game for their status review of fisher in California (letter from M. Schwartz 2009 in CDFG 2010), suggests the two fisher populations in California (northern California and southern Sierra Nevada) may have been separated for thousands of years (Knaus et al. 2011) and (Tucker et al. 2012). These preliminary reports, if validated, would have implications to the understanding of historical fisher distribution in the Sierra Nevada because such genetic differences would indicate a discontinuous range between the population (an apparent “gap” in occupied range) may have naturally occurred somewhere in the Sierra Nevada. This indicates that the isolated southern Sierra population may have persisted for a much longer period than previously hypothesized.

Habitat Relationships Fishers use large areas of primarily coniferous forests with fairly dense canopies and large trees, snags, and down logs. A vegetated understory and large woody debris appear important for their prey species. It is assumed that fishers will use patches of quality habitat that are interconnected by other forest types, whereas they will not likely use patches of habitat that are separated by large open areas lacking canopy cover (Buskirk et al. 1994). Buck et al. (1994) described 1970s research in managed Douglas-fir and white fir forests in northwestern California. They detected a selection pattern favoring residual stands of mature forest in areas heavily harvested for timber.

Riparian corridors (Heinemeyer et al. 1994) and forested saddles between major drainages (Buck 1983) may provide important dispersal habitat or landscape linkages for the species. Riparian areas are important to fishers because they provide concentrations of large rest site elements, such as broken top trees, snags, and coarse woody debris (Seglund 1995), perhaps because they persisted in the mesic riparian microtopography through historic fires.

Habitat suitable for resting and denning sites is thought to be most limiting to the population; therefore, these habitats should be given more weight than foraging habitats when planning or assessing habitat management (Powell et al. 1994, Zielinski et al. 2004b). Fishers generally use at least one rest site per day, and rarely reuse rest site structures (Kilpatrick et al. 1994, Seglund 1995, Zielinski et al. 2004b). Zielinski et al. (2004b) argue that retaining and recruiting trees, snags and logs of at least 39 inches in diameter at breast height, encouraging dense canopies and structural diversity, and retaining and recruiting large hardwoods are important for producing high quality fisher habitat and resting/denning sites. Freel (1991) also recommended 2 snags per acre over 44 inches in diameter at breast height and 4 to 5 snags per acre over 20 inches in diameter at breast height for suitable fisher habitat.

Average fisher home range sizes in the southern Sierra Nevada of California were previously displayed in Table 9 and Table 10 above. Zielinski et al. (2004c) speculated that the relatively small home range sizes of fisher in the southern Sierra Study Site located on Sequoia National Forest reflect higher habitat quality due to greater abundance of black oak that provides cavities and prey food resources.

West Coast Fisher Distinct Population Segment Habitats Examining numerous habitat association studies from British Columbia southward to California (Weir et al. 2007, Weir et al. 2010, Aubry et al. 2004, Klug 1997, Carroll 2005, Yaeger 2005, Self et al. 2001, Truex et al. 1998, Seglund 1995, Dark 1997, Buck et al. 1994, Slauson et al. 2007, Jordan 2007, Zielinski

100 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests et al. 2004b, Zielinski et al. 2004c, Zielinski et al. 2006, Zielinski et al. 2010b) a number of unifying themes emerge, as identified by the interagency, International West Coast Fisher Biology Team (Lofroth et al. 2010) and presented below.

1. Fishers occur in a variety of low and mid-elevation forested plant communities. Fisher populations occupy a diverse range of conifer, mixed conifer, and mixed conifer-hardwood forests. Southern fisher populations occurred in a wider diversity of forest types than did northern populations. Fishers were not found in high-elevation subalpine and alpine habitats or in dry, warm, open forest and grassland environments. 2. Fishers are associated with moderate to dense forest canopy. The most consistent predictor of fisher occurrence at large spatial scales was moderate to high amounts of contiguous canopy cover rather than specific habitat type. Research has suggested that inadequate canopy cover limits fisher distribution across forest types and ecoregions. 3. Fisher home ranges typically include a diversity of forest successional stages and plant communities. Composition of individual fisher home ranges generally included a mosaic of forested environments ranging from pure hardwood and mixed conifer-hardwood to pure conifer stands. Although fisher home ranges included a diversity of successional stages, they often included high proportions of mid- to late-successional forests. However, investigators have also found fisher home ranges including or positively associated with younger successional stages, likely due to prey resources associated with those environments. 4. Active fishers are frequently associated with complex forest structure. Active fishers are likely engaged in several behaviors, each of which may be linked to different habitat conditions including foraging for different types of prey, traveling between kill and rest sites, territory defense, and various social interactions. Fishers appeared to be more flexible in their use of various forest successional stages when active than when resting or denning. Nonetheless, active fishers typically avoided nonforested environments and early-successional forest stands that lacked canopy cover. During systematic surveys in the southern Sierra Nevada, fishers were detected at sites that were more structurally complex (greater than expected canopy cover, large trees, a hardwood component and next to a stream) than non-detection sites. 5. Fisher rest sites are strongly associated with moderate to dense forest canopy and elements of late- successional forests. Rest sites are presumably selected to provide one or more advantages to fisher (thermal, protection from predators, secure locations for consuming prey), therefore, forest attributes at these sites reflect the diversity in meeting these needs. In general, fishers selected rest sites near water with dense canopy cover, various attributes of coarse down wood, larger trees and a greater abundance of large trees than were typically available. 6. Fishers typically rest in large, deformed or deteriorating trees and logs. Thermoregulation, minimization of predation risk and the need for secure habitat for consuming prey are likely influences in fisher selection of resting sites. Fishers rested primarily in large live trees, followed by snags and coarse down wood. When resting in live trees, fishers have been observed using deformities associated with mistletoe and broom rust infections, large branches, interlaced branching structures, platform nests and cavities. Fishers primarily used cavities when resting in snags. Research suggests fisher select live trees and snags based on which species are most likely to develop rust brooms, mistletoe brooms, platforms, or cavities. Additionally, the large size of resting structures is likely due to several factors including tree age and the time required to develop various microstructures. 7. Cavities in large trees are a critical resource for reproduction. The strongest and most consistent habitat component reported across all fisher studies throughout the West Coast DPS was the use of

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cavities in large live and dead trees by reproductive females with kits. In general, cavities used by reproductive females for birthing and nursing were created by heartwood decay. Females gained access to the internal cavities through relatively small or narrow openings created by branches breaking away from the bole, cracks in the bole, fire scars and pileated woodpeckers. Reproductive females used a variety of species for den sites, however hardwoods were used for denning more frequently than conifers, even in areas where they were a minor component of the forest. Similar to rest sites, the large size of trees and snags used for denning is likely due to tree age and the time required for heartwood decay to develop and form cavities, and the size of cavities needed to accommodate an adult female and kits.

Southern Sierra Nevada Fisher Habitat

General Habitat is largely restricted to a narrow north-south band on mostly western slopes of mid-elevation forests in the southern Sierra Nevada Mountains (Spencer et al. 2008). It is associated with mesic topographic positions (northern slopes) in areas of lower precipitation (less persistent snow cover), and is concentrated in or near large old stands of mixed conifer, sequoia, and ponderosa pine, especially areas with black oak (Spencer et al. 2008). An analysis of Forest Inventory and Analysis plot data from suitable habitats demonstrated that highly suitable resting microhabitats in the form of clusters of very large trees surrounded by dense canopy are relatively rare (Spencer et al. 2008).

Mazzoni (2002) studied habitat use by fishers in the Kings’ River Project in the southern Sierra Nevada. Ninety percent of fisher rest sites were in large live trees (mean of 37 inches diameter at breast height) and large snags (mean of 40 inches diameter at breast height). Large logs as well as stumps and rock crevices were also used for resting. Selection for resting in white fir, ponderosa pine and black oak was evident, and selection against incense cedar and sugar pine was documented. Compared to random sites, areas of 2.47 acres surrounding rest sites had greater levels of canopy, coarse woody debris, basal area, crown volume and canopy layering. Rest sites were closer to water than random sites, and Mazzoni (2002) suggests this may be an artifact of riparian buffers that retain large structural elements of the habitat and dense canopy. The importance of ecological processes such as decay and disease, especially mistletoe brooms, are highlighted for creating fisher rest structures. This has also been documented in other portions of the fisher’s range (Paragi et al. 1996, Parks et al. 1999).

Zielinski et al. (2004b) found that female rest sites, when compared to random sites, included denser canopies, larger trees, steeper slopes, and greater presence of large conifer snags.

Den Site Selection. It is important to review data most localized to the analysis area because that best reflects availability and use of habitat elements in the specific geographic vicinity being analyzed. Where such data are lacking, expanding the data universe to include the nearest information is an accepted scientific practice. Due to ecological differences, separation of data for conifer and hardwood den trees is recommended (Truex et al. 2008). Natal dens refer to the site where parturition is assumed to have occurred, while maternal dens refer to sites where an adult female was observed resting with one or more kit(s) (Purcell and Thompson pers. comm.).

Den site structural elements must exist in the proper juxtaposition within specific habitats in order to provide a secure environment for birth and rearing of fisher kits. Natal dens, where kits are born, are most commonly found in tree cavities at heights of greater than 20 feet (Lewis et al. 1998). Maternal dens, where kits are raised, may be in cavities closer to the ground so active kits can avoid injury in the event of a fall from the den (Lewis et al. 1998).

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Den tree data collected in the Kings River Project area on the Sierra National Forest between 2007 and 2010, (Thompson et al. 2011, Table 11) included use of black oak, white fir, incense cedar, ponderosa pine, and sugar pine. Live black oaks selected as maternal den sites were among the largest oaks used and averaged 34.2 inches in diameter at breast height, while oaks used as maternal den sites were much smaller and averaged 23.6 inches in diameter at breast height. Live conifers used as natal dens averaged 45.2 inches in diameter at breast height, while those used as natal dens were smaller, averaging 37.9 inches in diameter at breast height. Forty-four of 93 maternal and natal dens (47 percent) were in black oaks, which do not typically leaf out until mid- to late-May, thus providing little canopy cover during actual use periods. Selection of these sites may be driven by their location and associated access to warming morning sun (K. Purcell pers. comm., C. Thompson pers. comm.). All confirmed births through the 2008 field season occurred between March 30 and April 11, and natal dens were occupied for 2 to 8 weeks. Natal dens refer to the site where parturition is assumed to have occurred, while maternal dens refer to sites where an adult female was observed.

Table 11. Natal and maternal den means for female fishers in Kings River Project area of Sierra National Forest through 2010

Den type Substrate Mean DBH (inches) Number of substrates

Natal Live conifer 45.2 11 Natal Live hardwood 33.7 13 Natal Snag conifer 39.2 5 Natal Snag hardwood 26.3 3 Maternal Live conifer 37.2 12 Maternal Live hardwood 25.5 31 Maternal Snag conifer 40.1 19 Maternal Snag hardwood not applicable 0

In 2007 and 2008, den sites in the Kings River Project area occurred in Sierran mixed conifer, montane hardwood-conifer and ponderosa pine forest types (K. Purcell, pers. comm.). Black oak was strongly selected as the den tree (C. Thompson pers. comm.). On the Kings River Project study area, natal dens (n=7) averaged 46 feet high with a range of 6 to 110 feet (K. Purcell, pers. comm.). Maternal dens (n=7) on the Kings River Project averaged 21.6 feet high, with a range of 9 to 41 feet (K. Purcell, pers. comm.). Generally, natal dens were found to be larger than maternal dens, only 1 hardwood snag was used, and conifer snags appear to be used more as maternal dens (K. Purcell, and C. Thompson, pers. comm.). As of 2009, average canopy cover was 74.3 percent (SD = 12.4, range 47.5 to 99.0, n = 51). Moosehorn readings at 2, 5, 10, and 15 meters, in 4 directions were averaged to measure canopy cover (K. Purcell, and C. Thompson, pers. comm.).

As of 1998 (Truex, et al., 1998), natal dens in the southern Sierra were located in white fir or black oak. Subsequently, most natal and maternal dens were in large conifers (white fir, sugar pine or ponderosa pine in southern Sierra) or oaks (California black oak in southern Sierra), generally in live form (Truex et al. 1998, Mazzoni 2002, Zielinski et al. 2004b). All natal dens were established during the last week of March or the first week in April and were occupied for 4 to 7 weeks. The canopy closure surrounding these den trees ranged from 89 percent to 97 percent, measured by spherical densiometer (implying a bias on the high side for remotely sensed canopy coverage, as typically measured by the Forest Service). The mean diameter of breast height of dens in white fir was 49.4 inches, compared to only 26.3 inches in black oak (

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Table 14). Natal dens are where parturition is assumed to have occurred, maternal dens are where an adult female was observed resting with one or more kit(s). Table 12 and Table 13 were derived from Truex et al. (1998). It is important to note the smaller diameter at breast height of oaks used as den trees, inferring that they achieve the requisite structural characteristics at smaller sizes than conifers. Similar information on tree species and size for natal and maternal den structures has been documented on the Sierra Nevada Adaptive Management Project (Table 16).

Table 12. Tree measures for natal and maternal dens with surrounding habitat for female fishers on the Sequoia National Forest from 1992-1996 Diameter Basal area Canopy Tree Tree Individual Den Type at breast (square feet Closure height Species Condition 1 (Inches) per acre) (percent) 1 Natal White fir snag 58.3 140 94 1 Natal Unknown snag 44.1 280 96 2 Natal White fir live 32.3 280 96 3 Natal Black oak live 39.0 260 93 3 Natal Black oak live 29.9 500 97 3 Maternal Black oak live 15.7 101 89 3 Maternal White fir live 57.5 121 93 4 Maternal Black oak live 20.5 262 96 1 Canopy cover was measured using a spherical densiometer held at waist height and therefore included tall shrubs and small trees. These understory layers provide ‘over-fisher’ cover below the overstory canopy that is typically measured in USFS aerial photo interpretation methodology. Thus, these canopy cover figures are considerably higher than remotely-sensed overstory canopy cover data. It is the remotely-sensed data that are used for project analyses.

Table 13. Mean of tree measures for natal and maternal dens with surrounding habitat for female fishers on the Sequoia National Forest from 1992-1996 Diameter Basal area Canopy at breast Study Area Tree Measures (square feet Closure height per acre) (percent)1 (Inches) Study area combined mean n/a n/a 94 Conifer mean 49.4 180 94 Hardwood mean 26.3 351 94 Live tree mean 32.5 254 94 Snag mean 51.2 210 95 1 Canopy cover was measured using a spherical densiometer held at waist height and therefore included tall shrubs and small trees. These understory layers provide ‘over-fisher’ cover below the overstory canopy that is typically measured in USFS aerial photo interpretation methodology. Thus, these canopy cover figures are considerably higher than remotely-sensed overstory canopy cover data. It is the remotely-sensed data that are used for project analyses.

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Table 14. Tree species, size, and height of live trees and snags used by fisher in the Sierra Nevada Adaptive Management Project Fisher Study Area for reproduction Number of Mean diameter at breast height Mean Total Height Tree Species sampes (inches) (meters) Black Oak Live 15 30.7 21.1 Black Oak Snag 2 40.9 8.8 Incense Cedar Live 15 46.1 33.9 Incense Cedar Snag 11 41.3 16.6 White Fir Live 11 40.6 34.8 White Fir Snag 12 44.5 34.3 Sugar Pine Live 3 49.6 37.1 Sugar Pine Snag 1 36.2 34.8 Ponderosa Live 2 35.8 34.9

Table 15. Mean of tree size and height of live trees and snags used by fisher in the Sierra Nevada Adaptive Management Project Fisher Study Area for reproduction Study Area Tree Measures Mean diameter at breast height (inches) Mean Total Height (meters) Study area combined mean 40.6 28.5 Conifer mean 42.0 32.3 Hardwood mean 35.8 15.0 Live tree mean 40.6 32.4 Snag mean 40.7 23.6

Rest Site Selection Large diameter black oaks and canyon live oaks compose almost half of the rest sites used by fishers in the southern Sierra Nevada (Zielinski et al. 2004b) (

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Table 14) while incense cedar were used less than expected. Data are from denning seasons 2008 – 2011. R. Sweitzer (unpublished data). Purcell et al. (Purcell et al. 2009) determined in the Kings River Project study area, fisher rest sites (regardless of species) averaged 37.5 inches for live trees and 46.0 inches for snags (

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Table 17). Additionally, from 2007 to 2011, rest sites of all trees in the Kings River Project area averaged 34.9 inches dbh, ranging from 7.8 to 78.4 inches (n = 283). Conifers used as rest sites averaged 37.6 inches while hardwoods averaged 27.9 inches (C. Thompson pers. comm.).

Table 16. Diameter at breast height (dbh) in inches of rest trees used by fishers on the Sequoia NF 1992 - 1996. Derived from Truex et al. (1998) Number of Mean Standard Deviation Median Tree Type Range (inches) Samples (inches) (inches) (inches) Conifer 181 40.2 19.6 11-171 42 Hardwood 146 25.6 8.5 12-57 25 Combined 317 34.8 not applicable not applicable not applicable

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Table 17. Mean values for fisher rest trees and snags in the Sierra National Forest, 1999-2001 Tree Type and Number Mean diameter Standard Number Mean Standard Fisher Gender sampled for at breast deviation of sampled for height deviation diameter height (inches) diameter height (feet) of height (inches) (feet) All Live 57 37.5 11.5 49 120.3 39.4 Females 47 38.5 11.9 39 122.2 39.3 Males 10 32.8 8.0 10 113.0 40.9 Conifers Only 49 37.2 12.2 342 130.5 32.6 Females 40 38.4 12.7 33 134.0 29.9 Males 9 32.0 8.2 9 117.5 40.6 Snags 12 46.0 18.6 11 55.6 47.3 Females 6 38.9 12.6 6 49.1 39.7 Males 6 53.0 22.0 5 63.4 59.1

Most resting structures used in the Kings River Project area were in live trees (76 percent), 15 percent were in snags, 3 were in logs and 2 each were in stumps and rock crevices (Purcell et al. 2009). Mean canopy cover as measured by moosehorn at rest sites was 73.7 percent, compared to random site canopy cover of 55.3 percent (Purcell et al. 2009). The majority (88.5 percent) of rest sites were in habitat with at least 20 percent canopy cover (Mazzoni 2002).

Examining data from Table 2 in Purcell et al. (2009), resting trees were predominantly ponderosa pine and white fir (

Table 18). In the immediate vicinity of the selected resting structure, ponderosa pine was used more than expected, while incense cedar was used less than expected (Purcell et al. 2009). Habitat at fisher resting sites had higher canopy cover, greater basal area of snags and hardwoods, and smaller and more variable tree sizes compared to random sites. Resting sites were also found on steeper slopes and closer to streams. Canopy cover was consistently the most important variable distinguishing rest and random sites (Purcell et al. 2009).

Table 18. Live resting tree species used by fishers compared to the number of available large trees (greater than 30 inches) in 2.47 acre plots surrounding them on the Sierra National Forest 1999-2001 Species All Females Males Available Incense cedar 4 3 1 222 White fir 20 19 1 236 Sugar pine 8 7 1 144 Ponderosa pine 23 17 6 193 California black oak 9 8 1 78

Estimated Number of Rest and Den Trees Required for Fishers in Each Home Range A review of available literature and anecdotal information was used to develop an estimate of forest structure used by a given fisher during their lifetime. Obviously, these numbers are somewhat speculative, but this provides what we consider to be a minimum number of resting structures that need to be available to fishers post-project. Given that fishers generally use at least one rest site per day, and have been reported to reuse only about 14 percent (range of 3 to 27 percent) of rest site structures (Seglund 1995, Self et al. 2001, Mazzoni 2002, Zielinski et al. 2004b, Yaeger 2005, Aubry et al. 2006), this equates to a minimum of 314 rest trees needed per an average southern Sierra Nevada female home range (2,357

108 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests acres) annually. The large range of 3 to 27 percent appears to be an artifact of divergent assumptions made in data analysis, with the larger numbers being reflective of number of incidents of reuse (counting multiple uses of a given tree as separate instances of reuse even by the same individual), while the lower numbers are the number of individual trees actually used more than once. In the future, we hope to reanalyze the data to a common definition of “reuse”. For project analysis purposes, the most useful number reflects the number of individual trees reused.

Reproductive females also utilize up to five den sites per year for a cumulative total of 319 potentially suitable trees needed per home range (or 0.14 trees per acre). The mean life span for fishers is approximately 10 years, equating to a minimum of 1.4 suitable rest or den trees needed per acre for each female home range over an average life span. Males would also require an estimated 314 rest sites, and with a mean home range of 9,518 acres this equates to 0.3 trees per acre over an average lifetime. Thus for an area to provide sufficient male and female rest and den site trees, more than 1.7 trees per acre are required. Because we don’t know what factors influence a fisher to decide to rest in one location versus another, there is a need to provide sufficient alternate rest and den tree choices to compensate for our lack of knowledge. Therefore we choose to buffer the 1.7 trees per acre by a factor of ten (selected to ensure availability of many more rest structures than are actually used) to maintain up to 17 potential resting/denning trees per acre, where they exist. This number should be re-aggregated to a total number of rest trees available in the home range of a female fisher. The spatial distribution of these structural elements is not uniform; suitable trees may occur individually or in widely scattered discrete clumps or patches. New data are being collected in several areas of the Sierra National Forest beginning in 2007, which will inform and improve this estimate in the future.

Home Range Composition Using data available at the time, Zielinski et al. (2004c) examined the vegetation composition of fisher home ranges in the southern Sierra Nevada as presented in the following paragraph (Table 19, Table 20, and Table 21). Since these figures are merely descriptions of information regarding home range composition selected relative to what is available, it should be noted that fishers may occupy areas that differ somewhat from values presented here. Additionally, the GIS data used in Zielinski et al. (2004c) lacked the spatial resolution to map small inclusions of shrub habitat within the greater mixed-conifer matrix. R. Truex (pers. comm.) believes that this fine grain heterogeneity is important from the perspective of prey diversity.

For the Sequoia National Forest, Sierran mixed conifer, ponderosa pine, and montane hardwood forest types comprised an average of 86 percent of the 12 (8 female and 4 male) fisher home ranges, with size- class 4 stands (11 to 24 inches in diameter at breast height) and canopy closure Class D (60 to 100 percent closure) comprising 61 percent and 66 percent, respectively, of the home ranges (Zielinski, et al., 2004c). CWHR size class 4 stands (11 to 24 inches in diameter at breast height), dense canopy closure (greater than 60 percent), and Sierran Mixed Conifer forest types constituted the greatest proportion of home ranges for female fishers. Home ranges for both sexes, rarely had less than 15 percent Sierran mixed conifer, less than 5 percent area in CWHR size class 5 (greater than 24 inches in diameter at breast height), or less than 53 percent dense canopy closure stands (dense stands included all size classes and vegetation types including live oak, plantation, and shrub layers). The montane hardwood type averaged 12 percent of home range areas for both sexes. For both sexes, CWHR size class 2 (1 to 6 inches in diameter at breast height) stands comprised generally less than 3 percent of home ranges, and less than 10 percent of home ranges supported open canopies (25 to 39 percent).

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Table 19. Female fisher home range CWHR forest type composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons Mean Home Standard CWHR System Forest Type Range Percentage Deviation Sierran Mixed Conifer 39 29 Ponderosa Pine 40 27 Montane Hardwood 14 15 Montane Hardwood-Conifer 6 4 Mixed Chaparral 0.2 0.4 TOTAL OF HOME RANGE 1 99 Not applicable 1 The remaining home range area (1.37 percent) consisted of small percentages of montane chaparral, urban and red fir

Table 20. Female fisher home range CWHR size class composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons Mean Home Standard CWHR System Size Class Range Percentage Deviation 5: med/large tree greater than 24” dbh 13 13 4: small tree 11” to 24” dbh 61 21 3: pole tree 6” to 11” dbh 22 28 2: sapling tree 1” to 6” dbh 2 2 TOTAL OF HOME RANGE1 98 Not applicable 1 The remaining home range size classes (2.14 percent) consisted of Class 1 (less than 1” dbh, 0.34 percent) or were not determined.

Table 21. Female fisher home range CWHR canopy closure composition on the Sequoia National Forest (n = 8), derived from Zielinski et al. (2004b) based on 100 percent minimum convex polygons Mean Home Standard CWHR System Canopy Closure Range Percentage Deviation Dense 60 to 100 percent 72 9 Moderate 40 to 59 percent 20 7 Open 25 to 39 percent 5 3 TOTAL OF HOME RANGE1 97 Not applicable 1 The remaining home range canopy closures were Sparse (10 to 24 percent) or undetermined.

On the Sierra National forest, Mazzoni (Mazzoni 2002) found home range composition by canopy closure class for males and females combined (n = 11), to be dense = 14 percent of the home range, moderate = 39 percent, open = 25 percent, and sparse (canopy closure less than 25 percent) = 21 percent of the home range area. These results differ substantially from those reported from the Sequoia National Forest in Table 19 to Table 21 above.

An Ecosystem Management Strategy for Sierran Mixed Conifer Forests North et al. (North, et al., 2009) describe an ecosystem management strategy for Sierran mixed conifer forests designed to address wildfire concerns as well as the persistence of rare species such as fishers. The following discussion draws from that paper.

Strategic fuels reduction activities can be designed to restore the ecology of the southern Sierra Nevada forests that have been actively managed for over a century. Topographic variables (i.e., slope shape, aspect, and slope position) can be used as a template to vary treatments in order to produce heterogeneous

110 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests stand structures and densities across the landscape. Localized cool or moist areas, where fire would have historically burned less frequently or at lower severity, support higher tree density and canopy cover, which provides pockets of legacy structure and habitat for fishers. In contrast, upper, southern aspect slopes have low densities of large fire-resistant trees. For thinning to achieve fuels reduction, marking rules could be based on crown strata or age cohorts and species, rather than uniform diameter limits. This system emphasizes the ecological role of fire, changing climate conditions, sensitive wildlife habitat, and the importance of forest structural heterogeneity.

Spatial heterogeneity is a key feature in ecological resiliency (Stephens et al. 2009). North et al. (2009) noted that a clumped tree distribution may slow the spread of crown fire, and that this spacing pattern may result from frequent fire in mixed-conifer forests. Cool, frequent fires reduce canopy cover to increase habitat and microclimate heterogeneity at site, stand and landscape levels, all of which benefit fishers by providing a diversity of prey habitats, and generate discrete clumps of large trees with dense canopy cover required for resting and denning.

Threats

Threats to the West Coast Distinct Population Segment The USFWS (USDI-FWS, 2004) identified major threats to fishers in the West Coast Distinct Population Segment (DPS), discussed relative to specified factors for listing under Section 4 of the Endangered Species Act. Only those threats deemed by USFWS (USDI-FWS, 2004) to be “important” to the entire West Coast DPS are summarized in this section. The reader is referred to the Federal Register for the complete discussion.

Factor A. Factor A is the present or threatened destruction, modification or curtailment of the species’ habitats or range. The extent of past and present timber harvest can fragment fisher habitat, reduce it in size, or change the forest structure to unsuitable for fishers. Both fuels reduction activities and effects of wildfire could result in loss and/or fragmentation of habitat. Development, recreation and roads also pose a threat of habitat loss and fragmentation, as well as direct mortality.

Factor B. Factor B is the overutilization for commercial, recreational, scientific or educational purposes. Historical trapping resulted in a severe population decline. Current mortalities or injuries from incidental trapping even where fisher trapping has been eliminated could be frequent and widespread enough to prevent population recovery or re-occupation of suitable habitat.

Factor C. Factor C is disease or predation. There is potential for disease outbreaks to occur in these small, isolated fisher populations with devastating effects. Mortality from predation by mountain lion, bobcat, coyote or large raptors could pose a significant threat to fishers.

Factor D. Factor D is the inadequacy of existing regulatory mechanisms. Some protections are available, but highly variable from jurisdiction to jurisdiction, and limited. Current regulations fail to provide sufficient certainty that conservation efforts will be implemented or that they will be effective in reducing threats to fishers.

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Fishers and Climate Change The Intergovernmental Panel on Climate Change (IPCC 2001) projects a doubling of atmospheric carbon dioxide (sometimes referred to by the acronym CO2) from industrial sources by as early as 2050. Climate responses to increased atmospheric carbon dioxide are expected to vary regionally and topographically, but a universal trend towards warming is expected due to trapping of heat by greenhouse gases. California is thought to be highly vulnerable to the effects of climate change due to coastal and latitudinal orientation, extreme elevational gradients, and the variety of ecosystems present (Snyder et al. 2002). Because California’s ecosystems are already stressed by human growth and agricultural demands, added stress from climate change could substantially alter the current biotic landscape.

Snyder et al. (Snyder et al. 2002) modeled climate change for California based on a projected doubling of atmospheric carbon dioxide and concluded that a warming trend would occur across the state with the greatest temperature changes in the Sierra Nevada (where average annual spring temperatures could increase by as much as 12.7 degrees Fahrenheit). According to Dettinger (2005), the most prevalent prediction of more recent climate change models is that temperatures will warm by about 9 degrees Fahrenheit by 2100, while precipitation will remain similar or slightly reduced compared to present levels. Lenihan et al. (2006) analyzed the responses of vegetation distribution to three future climate scenarios in California and similarly predicted dramatic increases in mid-elevation mixed evergreen forests (conifer or oak), primarily as a result of increased temperatures. This same analysis (Lenihan et al. 2006) predicted decreases in conifer forests (pine/mixed conifer) due to increased fire. In fact, Lenihan et al. (2006) projected that relative to the past century, the annual acreage burned would increase 10 percent to 50 percent for the period 2050 to 2099.

These predicted climate changes may benefit fishers and their habitat in one of three ways:

1. decreased snow levels would open up greater areas of potential habitat, given the animal’s tendency to avoid deep snow (Raine 1983, Arthur et al. 1989, Aubry et al. 1992) 2. increased rainfall during the growing season may result in increased vegetative productivity leading to more food for fishers and their prey and more resting sites 3. an upslope expansion of mixed evergreen forest habitats including more oaks that fisher inhabit in the southern Sierra Nevada and in the Klamath region of northern California could also benefit fishers (Self et al. 1992, Klug 1997, Higley et al. 1998, Zielinski et al. 2004b, Zielinski et al. 2004c, Yaeger 2005) Conversely, the predicted hot dry summers could lead to a great increase in the frequency of uncharacteristically severe stand-replacing wildfires, most notably should wet and warm winters and springs contribute to increased fuel loading, and if current human-caused fire suppression policies are extended into the future. Fire regimes respond rapidly to changes in climate and are likely to drive the short-term responses in terms of vegetation floristics and structure (Flannigan et al. 2000, Dale et al. 2001). Large fires could accelerate habitat fragmentation, especially in coniferous forests, and result in the loss of fisher population viability. Greater incidence of wildfires may reduce the frequency and alter the distribution of important structural features used by fishers such as large trees and high canopy cover (Safford 2006). While these wildfires may result in the temporary creation of snags and coarse woody debris, increased fire frequencies may reduce the availability of these structural features over the long term (Safford 2006).

Lenihan et al. (2006) predicted that due to increased wildfire and changes in moisture regimes, continental coniferous forests would be replaced by more fire-tolerant mixed evergreen forests with oak components. In summary, future climate change may result in an increase of the forest types where fisher are currently detected in the southern Sierra Nevada and northern California at the expense of mixed

112 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests conifer forests in the Sierra Nevada traditionally thought as prime fisher habitat. It should be noted that forest conversion may require the intermediate step of a stand-replacing fire (with either temporary or permanent habitat loss resulting), and the subsequent maintenance (by fire) of a more open-canopied forest less suitable to fishers (basically the conversion to a favorable forest type but an unfavorable forest structure).

Further, climate change may affect the ability of fishers to expand their current range. The extant fisher population in the southern Sierra Nevada exists at the animal’s southernmost range, where increased temperatures are predicted to have the greatest impact. There is a possibility that fisher range may shift upward in elevation to track the forest types that best meet their habitat needs. If the existing population’s ability to expand northward is limited by forest fragmentation or other natural barriers, then climate change may eliminate these populations before the barriers limiting expansion are lifted (e.g., before forest succession improves habitat in fragmented areas, or the episodic freezing or drought conditions occur allowing fishers to cross river barriers).

Therefore, the only supportable conclusion is that it is unclear whether modeled climate change would benefit or adversely affect fishers over the long term. The ultimate fate of the species and its habitat may depend on the interactions of other factors influencing fisher conservation. It does seem likely, however, that the future will challenge the ability of fishers to adapt to a changing Sierra Nevada climate and ecology.

Uncharacteristically Severe Wildfire The cessation of burning by indigenous peoples and the implementation of fire suppression policies has negatively affected many forests in the southern Sierra Nevada. Past fire suppression policies to suppress all human and lightning-caused fires, have resulted in widespread accumulation of forest fuels and have moved forests beyond the natural fire regimes of relatively small, low-intensity fires to larger, more complex high-intensity fires. Subsequently, forests are experiencing changes in plant species composition, reduced productivity and structural heterogeneity, as well as increased susceptibility to insect infestations (Lofroth et al. 2010).

Uncharacteristically severe wildfire is defined as fire occurring beyond the historical range of natural variation in terms of scope, intensity and duration. These stand-replacing fires affect large areas of the landscape, decreasing or removing key fisher structural and habitat elements including large trees, overstory and understory canopy, vegetative diversity, snags, and logs. Landscape permeability for fisher movements at all scales may decrease as a result. As part of the threat evaluation completed for the West Coast Fisher Conservation Assessment (Lofroth et al. 2010), uncharacteristically severe wildfire ranked as high threat in the southern Sierra Nevada geographic area. Conversely, Zielinski et al. (2010a) determined that despite recent fire recorded within the Sequoia National Forest, predicted fisher resting habitat values increased, likely as a result of increased canopy cover values recorded several years post wildfire. However, the type of wildfire was not described, therefore, further investigation is necessary to determine the severity of fire and the habitat types affected that would produce increased canopy cover within a limited time frame.

There is evidence that uncharacteristically severe wildfire effects are increasing, particularly in the southern Sierra Nevada. This is particularly true for the southern portion of the Sequoia National Forest where the project is located. The southern Sequoia National Forest has experienced several large and uncharacteristically severe wildfires. These fires include the Flat (1975; 18,729 acres), Bonita (1977; 7,425 acres), Fay (1987; 12,147 acres), Stormy (1990; 22,182 acres), Jacks (1997; 5,744 acres), Manter (2000; 79,182 acres), McNally (2002; 149,414 acres), and Goldledge (2007; 4,196 acres). The fire history records analyzed within and around the vicinity of the proposed project area indicated some of these fires

113 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests resulted in the stand-replacement of thousands of acres of conifer forest and their associated wildlife habitat. Forest fire history records also indicate that most of the catastrophic fires started burning outside of the analysis area at lower elevations in grass and brush fuel types then have transitioned into conifer forest fuel types like those represented in the proposed project area. Over the last 40 years, sensitive wildlife habitat is being lost faster than it can be grown back. The Tobias Project is surrounded by large wildland fire scars; Stormy Fire (1990), McNally (2002) to the north and within the project area, and the Goldledge (2007) and Bull (2000) to the east (Map 6). A significant proportion of the fires in the last decade have included uncharacteristically severe effects on forested stands within the 3,500 to 7,500 feet elevation band of most value to fisher.

Fragmented landscapes created by uncharacteristically severe wildfires are likely to eliminate fisher habitat linkages, either permanently via vegetative type conversion or temporarily until recovery occurs. Landscape permeability to fishers is decreased. This results in detrimental impacts to fisher daily movements and energy balance, creates barriers to dispersal movements, affects the establishment of home ranges, and prolongs or prevents breeding season movements. These impacts may decrease fisher survival. Overall population fitness is affected by individual survival and mortality. Direct mortality as a result of fire may occur in extreme cases depending upon season (e.g. kit loss in reproductive season, loss of adults in fast-moving canopy firestorms either directly or from potential smoke inhalation) (personal observation of fire related fisher mortality, Pierce Fire 1987, S. Anderson).

Following a wildfire, prey species abundance and community composition will shift. An initial increase in abundance of disturbance-adapted prey species may occur at the expense of species diversity with a gradual reversal of this trend as succession occurs. Although prey abundance may increase, prey availability will not necessarily follow due to fisher reluctance to enter open areas. Extensive burned areas can create dispersal barriers for prey. The West Coast Fisher Biology Team speculated that the abundance of prey available following fire may support pre-fire population levels of fishers that have been compressed into adjusted home ranges. This prey abundance may not persist over time, however, and result in displacement or loss of fishers on the margins of remaining habitat (D. Macfarlane, pers. comm.). Displaced individuals could create within species competition if packed into the remaining habitat, which could, in turn, increase disease transmission.

Fishers exhibit strong selection for rest and den sites based upon forest structure and canopy cover. Changes in the frequency, abundance, and distribution of these habitat elements may create conditions inimical to successful reproduction, as well as survival of the young to recruitment into the population. Lack of well-distributed escape cover may result in increased predation.

It is unknown whether or to what extent fishers exhibit site fidelity. Habitat changes due to uncharacteristically severe wildfire could temporarily disrupt fisher social organization in a manner difficult to conceptualize (D. Macfarlane, pers. comm.). Resident animals may continue to occupy the burned area, but might not be replaced via recruitment of young into the population or via emigration of other adults upon their death. These socially-mediated population impacts may be exhibited as a lag effect. That is, they may require an average fisher lifetime (10 or more years) under a statistically rigorous monitoring program for at least that period of time to become evident.

Vegetation Manipulation to Reduce Risk of Uncharacteristically Severe Wildfire Truex and Zielinski (Truex et al. 2013) developed fisher resource selection functions and resource selection probability functions as described in Zielinski et al. (2004b) to compare rest sites selected and track plate detections to areas not selected or sampled with no detections. These resource selection functions were used to estimate the change in fisher habitat suitability pre- to post-treatment in fuels

114 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests reduction projects at two sites in the Sierra Nevada. The remainder of this section discusses the results of the Truex and Zielinski (Truex et al. 2013) study.

Four primary treatments were applied for effects assessment: control (no treatment); mechanical harvest (usually including mastication following harvest); mechanical harvest followed by prescribed burning; and an area where prescribed burning was the only treatment. Study areas were the Blodgett Forest Research Station and a satellite site at Sequoia and Kings Canyon National Parks.

This study generally concluded that fire and fire surrogate treatments have modest but significant short- term effects to the quality and availability of fisher resting habitat, as well as canopy closure. At Blodgett Forest Research Station, mechanical as well as mechanical plus fire treatments significantly reduced fisher resting habitat and average canopy closure. At the Sequoia and Kings Canyon National Parks site, the late season burn treatment had a significant effect on fisher habitat suitability as well as canopy closure. The short-term treatment effects to foraging habitat at both sites were generally not significant. This may be explained by the broad spectrum of foraging habitat parameters, rendering it less likely to be a limiting factor to fisher than resting habitat.

Although the mechanical and mechanical/fire treatments had greater effects on fisher resting habitat suitability than prescription fire at Blodgett Forest Research Station, these effects can be mitigated by the ability of mechanical treatments to avoid individual habitat elements such as the critically important hardwoods and large trees. The use of prescribed fire alone can be mitigated by raking debris away from key fisher structural elements in the habitat. The effect of greatest magnitude was a reduction in canopy closure. All treatments reduced canopy closure. Canopy closure, however, recovers relatively quickly compared to the loss of large dead or live trees.

Interpretation of these results needs to be cautious and informed by more data in the next decade. In areas where fisher habitat suitability is already low or marginal, the predicted effects may have a disproportionately large impact to habitat recovery. On the other hand, the short-term negative effects of the treatments may result in beneficial effects on subsequent stand development. Future monitoring will be needed to elucidate the exact nature of this relationship.

Another limitation of this study is that it focused upon effects at the individual stand level. As wide- ranging predators, fisher function at larger landscape scales within their habitats. Thus, it is important to analyze the spatial and temporal array of treatments in a landscape context. The more broadly distributed the treatments are over space and time, the lower the likelihood of significant negative effects in a landscape context. It seems likely that such treatments distributed over space and time should have lower impacts than large-scale catastrophic wildfire.

One last caveat offered by Truex and Zielinski (2013) in interpreting the study results is to recognize that a reduction in habitat suitability does not necessarily equate to loss of suitability. Population level implications to localized reductions in habitat suitability have yet to be studied. To decrease effects to fisher habitat suitability, the authors recommend planning treatments to maintain elements important to fisher (e.g. large diameter hardwoods). Early season burns (mid-May or later) timed to follow the fisher denning period seem to have less impact to habitat. However, K. Purcell and C. Thompson (pers. comm.) have noted that by mid-May the kits still have relatively limited mobility; they are still largely dependent on the female until the end of August. Thus, to avoid potential conflict with denning, early season burns (spring burns) should occur prior to mid-March. Planning treatments to occur dispersed over space and time to the extent possible will minimize the effect to individual fishers.

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The Trade-Offs Between Fuels Reduction Activities and Wildfire The Conservation Biology Institute (CBI) conducted a computer simulation study of the interactions between fuels management, forest fires, fisher habitat, and the fisher population in the southern Sierra Nevada (Spencer et al. 2008, Syphard et al. 2011, Scheller et al. 2011). The study area included the Stanislaus, Sierra, and Sequoia National Forests and Yosemite and Sequoia-Kings Canyon National Parks. This transparent, objective, and highly collaborative process was guided and informed by input from a group of independent science advisors and stakeholders.

To assess the status of fishers and to predict future changes in their habitat and population viability under various management and fire scenarios, a GIS landscape level fisher habitat model was created and coupled with:

1. a landscape change model that simulates the effects of fires, forest management actions, and ecological succession on forest vegetation called LANDIS-II (Scheller et al. 2007), and 2. a fisher population model called PATCH (Schumaker 1998). PATCH is a stochastic, spatially-explicit wildlife population simulation model particularly sensitive to landscape quality and pattern. It facilitates exploration of vegetation management effects over time. PATCH tracks the success of females and their young to estimate the likelihood of species persistence on defined landscapes.

Coupling these models facilitated evaluation of likely changes in the quality and distribution of fisher habitat and populations under alternative forest management and fire scenarios projected 50 years into the future. Simulations were replicated within an experimental framework that allowed systematic exploration of the relative and absolute effects on fishers of alternative fuel treatment rates, intensities, locations, and fire regimes. This design allowed investigation of how these factors, alone or in combination, might affect fisher habitat and populations.

Thus, simulations using this regime assumed that fire behavior over the next 50 years would be similar to that observed over the past two decades. Because there has been a trend toward larger, more severe fires in recent decades, and many experts believe this trend will be exacerbated by climate change and other factors, a “high fire” regime was defined by restricting the model to simulating fires only under severe fire weather conditions. These changes in the weather data resulted in the simulation of generally larger, more severe fires. Three fuel treatment rates (2, 4 and 8 percent) were tested, which were defined as the proportion of treatable landscape treated over a five-year time interval. The potentially treatable landscape consisted of lands within the three affected National Forests (Sequoia, Sierra and Stanislaus), but excluded non-treatable areas, such as Wilderness Areas, Wild and Scenic River areas, Research Natural Areas, non-vegetated land, and spotted owl protected activity centers.

Two vegetation management prescription intensities (light and medium) were tested, as defined by USDA Forest Service personnel. Both prescriptions assumed a combination of mechanical treatment followed by prescribed fire on slopes less than 30 percent, and prescribed fire only on slopes greater than 30 percent. In both cases, the mechanical treatments simulated “thinning from below,” which reflected preferential removal of most biomass (about 80 percent) of the smallest trees, and progressively lesser proportions of biomass removed with increasing tree size. The maximum diameter of trees removed under the “light” treatment was 12 inches, while the maximum diameter of trees removed under the “medium” treatment was 30 inches. Neither of the simulated intensity treatments removed substantial biomass of larger trees (e.g., 20 to 30 inches diameter at breast height). In simulations, these treatments were not widely divergent from one another in terms of affecting forest biomass or post-treatment fuels conditions, as the

116 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests medium intensity treatment, on average, removed only about 20 percent more total tree biomass than the light treatment within treated stands.

Results of the simulations, however, demonstrated that treatments may effectively reduce the extent and severity of fire on the landscape over a 50-year time span. Given the right combinations of treatment rate, intensity, and location, the benefits to fishers of reducing fire outweigh the cumulative negative effects of the treatments themselves on fishers. Across the broad spatial scales CBI examined, given specific assumptions disclosed in Spencer et al. (Spencer et al. 2008) about how thinning treatments affect fuel characteristics, fire spread rates, and fire severity, and within the finite combinations of fire regimes and treatments tested, it was concluded that the long-term positive effects of fuel treatments (due to the reduction of fire hazard) outweighed the short-term negative effects of fuel treatments (due to immediate loss of forest biomass) on fisher. This was especially true assuming a more severe fire regime in the future. Truex and Zielinski (2013) documented expected short-term detrimental habitat changes for fishers as a result of fire and fuels reduction treatments. Spencer et al. (2008) places the tradeoffs of short- term habitat degradation for long term benefit in clear context for the southern Sierra Nevada fisher population and habitat as a whole, demonstrating specific conditions where short-term detriment for long- term habitat maintenance is acceptable.

The area of treatment, the intensity of treatments, and the location of treatments all affected the total amount and severity of fire under certain conditions, but these effects did not universally translate into significant effects on the simulated fisher population. If no treatments were simulated under the more severe “high fire” regime, then the modeled fisher populations declined substantially from their baseline population levels due to removal and fragmentation of large amounts of forest biomass due to wildfire. However, no significant declines were seen under “baseline” fire conditions in the absence of treatments.

Although there is scientific uncertainty to what degree fire conditions may become more severe in the future, most experts on Sierra Nevada vegetation and fire believe that wildfires will continue to become larger and more severe relative to recent history due to climate change, fuel accumulation, insects, droughts, and other factors. To the extent that the past two decades of fire histories are reflective of the next 50 years, CBI modeling results suggest that fuels treatments will have little effect on fishers, either positively or negatively, at the regional scale. However, if, as most experts agree, fires will become on average larger and more severe in the future, CBI analysis results suggest that carefully planned and implemented fuels treatments may reduce fire risks and help to sustain fisher habitat and populations.

Simulated fuels treatments inside fisher habitat were generally more effective in protecting fisher habitat from fire than treatments located outside fisher habitat, but adding treatments outside yet adjacent to fisher habitat would likely yield additional benefits to fisher under the more severe fire regime.

Treating only 2 percent of the treatable landscape every 5 years (or up to 10 percent of the treatable landscape over 20 years) had no significant effect on fire or fishers at the landscape level, while treating 4 percent to 8 percent of the treatable landscape every 5 years (or up to 20 to 32 percent of the treatable landscape over 20 years) was effective in reducing fire and benefiting fishers. Other fire modeling studies (e.g., (Finney et al. 2007) have concluded that, for fuels treatments to be effective in reducing fire size over large, forested landscapes, roughly one-third of the landscape needs to be treated over 20 years. This is in rough accord with CBI findings that treating only 10 percent of the landscape over 20 years was ineffective, but treating up to 20 to 32 percent over 20 years was effective in reducing fire spread rates and sizes.

Zielinski et al. (2013) sampled fisher home range-size areas for fisher scats, using scat detector dogs, and found that the areas with the most abundant scats had an average of 2.6 percent of their area disturbed per year (equivalent to 13 percent over a 5-year period) by a combination of vegetation management

117 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests treatments. This exceeds the minimum proportion of treated landscape that may be necessary to reduce the size or severity of future fires (e.g., 8 percent treated per a 5-year period, Syphard et al. 2011; 2 percent annually, Finney et al. 2007). The degree of disturbance within sample units varies widely, and fishers may in some circumstances tolerate even higher rates of disturbance. In one of five high-use units and one of three moderate-use units, about 6.5 percent of the area was disturbed annually on average (Zielinski et al. 2013). They also found evidence of greater tolerance to treatments designed primarily for forest restoration goals than those with more emphasis on commercial timber harvest.

The medium intensity treatment tended to decrease the spread and amount of fire on the landscape more than the light intensity treatment under certain conditions and in certain fire regimes (e.g., at mid and high elevations for the “baseline” regime and high elevations for the “high fire” regime). However, over the range of treatment combinations modeled, there was no significant difference in simulated fisher populations resulting from applying the two treatment intensities, in large part because the range of simulated treatments was too narrow to result in landscape-level differences in effects on fisher habitat (the medium intensity treatment removed only 20 percent more biomass than the light treatment within treated stands) and due to high stochasticity among replicates in the simulations of fires and fisher populations.

The final CBI report (Spencer et al. 2008) concluded that the relative merits of light versus moderate treatment intensities need to be evaluated on a finer scale. CBI recommend that fuel treatments be designed to reduce fire spread rates and severity based on site-specific analysis also taking fisher habitat value in and near the treatment into consideration. Under a heightened fire regime, placing treatments outside fisher habitat is beneficial, and a combination of inside and outside fisher habitat is best. In areas of very high fire hazard and outside of fisher habitat, higher intensity treatments may be warranted. Within fisher habitat, treatments should be designed to balance desired fuel conditions with maintaining sufficient overstory and habitat elements to sustain or encourage occupancy by fishers. Under the baseline fire regime, locating fuels reduction treatments inside fisher habitat provided the greatest benefit to fishers.

CBI concluded that removing larger trees and other essential habitat elements should generally be avoided within fisher habitat, to the degree feasible while meeting fuel reduction and landscape vegetation management goals. Although not specifically stated in Spencer et al. (2008), we assume that the CWHR definition of a medium to large tree as being equal to or greater than 24 inches in diameter at breast height is basically what was intended by “larger trees” in the CBI conclusion; however, an equally strong case could be made for equal to or greater than 30 inches in diameter at breast height based on rest site data (D. Macfarlane, pers. comm.). There could be benefits to fisher of removing, for example, some larger firs or cedars near black oaks to stimulate growth of the oaks, which provide important habitat elements for fishers and their prey. Indeed, the models predicted that the distribution of old black oaks in the project area is likely to increase under all treatment combinations, providing increased rest site opportunities and prey availability for fishers.

CBI simulation results to date cannot be directly applied to the question of whether removing a significant number of larger trees within a stand will reduce fire hazards or help protect fisher habitat. Nevertheless, simulated results did show that removing primarily small trees (“ladder fuels”) in both of the treatment intensities were effective at reducing the extent and severity of fire.

Habitat Fragmentation or Loss of Connectivity Land ownership, human settlement, and timber harvesting patterns have significantly influenced forest landscapes in the southern Sierra Nevada (Lofroth et al. 2010). Population growth and encroachment with its associated infrastructure such as roads, housing, reservoirs, powerlines, have further fragmented the

118 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests forested habitat. In some instances, these lands are permanently alienated and may act as barriers to movement for fisher. Habitat connectivity is a key to maintaining fisher within a landscape. Activities under Forest Service control that result in habitat fragmentation or population isolation pose a risk to the persistence of fishers. Timber harvest, fuels reduction treatments, road presence and construction, and recreational activities may result in the loss of habitat connectivity resulting in a negative impact on fisher distribution and abundance.

Disease Although there have been no documented cases of disease in wild populations causing widespread mortality and subsequent population declines, recent studies from the Kings River Project and Sierra Nevada Adaptive Management Project indicate disease may be more of a significant factor in fisher mortality than previously known (M. Gabriel, pers. comm.). Serology tests of fisher show past exposure to several infectious diseases, including parvovirus, canine adenovirus, toxoplasmosis, and canine distemper (Brown, et al., 2008), (Sierra Nevada Adaptive Management Project from http://snamp.cnr.berkeley.edu/documents/Fisher/ as accessed December 1, 2009). These diseases may lead to mortality due to predation, vehicle accidents, and starvation in addition to direct mortality from disease.

Indirect Effects The strategy for the Southern Sierra fisher was utilized in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3.

Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: The Forest Plan was created with the recommendation from the Strategy for the Southern Sierra Fisher, hence provides a framework that will allow for the persistence of the fisher on the landscape. Habitat components such as those associated with late seral stage components will be maintained or enhanced. Threats such as uncharacteristic wildfires are being ameliorated through the implementation of the plan. Climate change is being addressed through creation and maintenance of corridors and landscape linkages so that the species can move from one location to another.

Determination Statement Inyo NF: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on this forest.

Sequoia and Sierra NFs: Implementation of the Forest Plan may affect fisher but will not lead towards Federal listing or a loss of viability. Impacts to fisher are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags.

119 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests

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Inyo Mountains Slender Salamander (Batrachoseps campi) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The Inyo Mountains slender salamander is a California endemic species. Its limited range is restricted to 16 localities within the Inyo Mountains between Waucoba Mountain and New York Butte (Jennings and Hayes 1994). These sites are within canyons on both the eastern and western slopes east of Lone Pine in Inyo County (Morey 2000). These sites comprise a total occupied habitat of less than twenty hectares; however there are likely a few additional populations yet to be found (Hansen and Wake 2005). The elevational range of this species is from 550 meters at Hunter Canyon to 2,620 meters at Upper Lead Canyon (Jennings and Hayes 1994). All known populations occur on federal lands on either U.S. Bureau of Land Management lands or National Forest System lands (Hammerson 2004). See Figure 10 for a locality map.

Within Region 5 all populations are on the Inyo National Forest.

The abundance of this species is unknown. However, as each known population has both a small range and is isolated from other populations the overall population size is likely small (Hammerson 2004). As a result of habitat degradation by both human and natural causes several populations have decreased significantly in number since the species was first described in the 1970’s (Hansen and Wake 2005). Further study is needed to determine abundance and population status on National Forest System lands.

In the NRIS database, the Inyo NF has 167 records, Sequoia and Sierra NFs have no records. Figure 11 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status All inhabited sites are isolated areas of suitable habitat surrounded by uninhabitable desert or semi-desert terrain. Within the Inyo Mountains these salamanders are typically found within the east to west trending canyons in suitable habitat. Perennial springs and seeps and associated riparian plant communities seem to be a requirement for this species. Common plants in these areas include willows, wild rose, and coyote brush which grow into dense thickets. Inyo Mountain slender salamanders seem to prefer areas where solid-rock cliffs, outcrops or talus are in contact with surface flow. Inyo Mountains slender salamanders have been found under rocks resting on wet substrate, woody debris, within clumps of moist ferns in waterfall spray zones, and in fissures or crevices within adjacent granitic or limestone outcrops. All records except for one are from riparian sites, this lone exception is of an animal that was captured near openings in a rock formation on a ridge between two canyons. This suggests the possibility that these salamanders may not be restricted to canyon bottoms at higher elevations (Hansen and Wake 2005).

Feeding likely occurs opportunistically both above and below ground (Morey 2000). The diet is likely similar to other Batrachoseps salamanders including earthworms, small slugs, various terrestrial arthropods and insects (Stebbins 1951).

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Figure 10. Map of Inyo Mountains slender salamander locations from multiple sources, 2010

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Figure 11. Map of Inyo Mountains slender salamander from NRIS databases, 2016

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Inyo Mountains slender salamanders live in the vicinity of perennial springs and seeps which provide a moist, thermally buffered environment. This environment allows for surface activity to be possible at most known sites year-round. Animals have been found in every month of the year. In higher elevations winter snowfall and low temperatures likely limit surface activity but this is unknown due to the difficulty in accessing these sites during winter (Hansen and Wake 2005). Like the other Batracroseps salamanders, the Inyo Mountains slender salamander is likely nocturnal (Morey 2000).

Slender salamanders have very small home ranges. Studies on a similar localized California slender salamander (B. attenuatus) found that adult salamanders moved an average of approximately 5 feet over a two year period and were found to repeatedly use the same cover object (CWHR 2005). In general slender salamanders demonstrate high site fidelity and rarely move more than 5 to 10 meters over their lifetime (Cunningham 1960). Inyo Mountains slender salamanders likely have a home range of less than one acre (Morey 2000). Home range and territoriality in this species has not been studied. Presumably territories are small or seasonal as breeding sized Batrachoseps spp. are often found under the same cover object in close proximity to one another (CWHR 2005).

Little is known about the reproduction of this species. Nest sites are unknown, however eggs are probably laid in moist crevices within rocky outcrops (Hansen and Wake 2005). Similar to other Batrachoseps salamanders, Inyo Mountains slender salamanders likely undergo direct development (Jennings and Hayes 1994). Some of the type specimens of this species were found to have enlarged yellow ovarian eggs in September and October 1973 (Marlow et al. 1979). Additionally, a female found mid-April contained eggs visible through the abdomen (Hansen and Wake 2005).

Threats As this salamander has a small, limited range, it is vulnerable to stochastic events such as flash floods, fire or climate change. Flash flooding in particular is a threat to this species. In 1985 flash flooding scoured the vegetation from Long John Canyon and in 2001 a similar event occurred in French Spring. In both cases the salamander population was not extirpated, however it was likely significantly impacted (Hansen and Wake 2005).

Anthropological threats primarily involve habitat degradation as a result of water diversion for mining, spring capping, or spring opening for upland game and cattle. All of these activities have had significant negative impacts on this species. In Barrel Spring, the original corridor where salamanders could be found was reduced by 90 percent as a result of water diversion and road construction. Another major threat is habitat degradation caused by feral horses and burros (Marlow et al. 1979). Additional threats to this species include natural predators and disease. Although Batrachochytrium dendrobatidis has been documented on the widespread California slender salamander (Batrachoseps attenuatus), the actual impacts of chytridomycosis on this species is unknown (Weinstein 2009).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Inyo Mountains slender salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and

134 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

Determination Statement Inyo NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Sequoia and Sierra NFs: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on these forests.

Literature Cited – Inyo Mountains Slender Salamander CWHR Program Staff. 2005. Kern Plateau Salamander Batrachoseps robustus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Cunningham, J.D. 1960. Aspects of the ecology of the Pacific Slender Salamander, Batrachoseps pacificus, in southern California. Ecology 41: 88-99.

Hansen, R.W. and Wake, D.B. 2005. Batrachoseps campi Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 31 May 2012.

Hammerson, G. 2004. Batrachoseps campi. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. http://www.iucnredlist.org. Downloaded on 31 May 2012.

Jennings, M. R. and Hayes, M. P. 1994. Inyo Mountains Salamander Batrachoseps campi Brode and Wake 1979. In: Amphibian and Reptile Species of Special Concern in California. California Department of Fish and Game, Sacramento, California.

Marlow, R.W., Brode, J.M., and Wake, D.B. 1979. A new salamander, genus Batrachoseps, from the Inyo Mountains of California, with a discussion of relationships in the genus. Contributions in Science Natural History Museum of Los Angeles County 308: 3-17.

Morey, S. 2000. Inyo Mountains Salamander Batrachoseps campi. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Stebbins, R.C. 2003. A field guide to western reptiles and amphibians. R.T. Peterson. Houghton Mifflin Company, New York, New York.

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Weinstein, S.B. 2009. An aquatic disease on a terrestrial salamander: individual and population level effects of the amphibian Chytrid fungus, Batrachochytrium dendrobatidis, on Batrachoseps attenuatus (Plethodontidae). Copeia 2009(4): 653-660.

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Relictual Slender Salamander (Batrachoseps relictus) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The relictual slender salamander was recently split into several different species based on genetic analysis. After this species split the remaining relictual slender salamander populations are known from only two extant populations which results in the relictual slender salamander having the smallest known range for a described Batrachoseps species (Jockusch et al. 2012).

The two documented populations both exist on the Sequoia National Forest (Figure 12). The first population was documented in 1979 east of Squirrel Meadow. The second population exists on Breckenridge Mountain. Salamanders have been observed at both high and low elevation; however a large portion of the mountain remains unsurveyed. Streamside and seep habitats that drain into the northern slope of Brekenridge Mountain may have additional populations (Jockusch et al. 2012). Population size of this species is unknown.

In the NRIS database, the Inyo NF has no records, Sequoia NF has 17 records, and the Sierra NF has no records. Figure 13 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Both extant populations of relictual slender salamanders are found in a small area of high elevation pine- fir forest (Jockusch et al. 2012). Presumably this species occurs mainly in heavily forested areas consisting of mixed pine-fir-cedar-oak habitat. Animals can be found under logs, rocks, at the edge of springs and seepages and in the water (Hammerson 2004).

The life history of the relictual slender salamander is not well known as most published studies are actually studies of the recently split Greenhorn Mountains slender salamander (B. altasierrae) (Jockusch et al. 2012). Feeding likely occurs opportunistically both above and below ground (Kucera 1997). The diet is likely similar to other Batrachoseps salamanders including earthworms, small slugs, various terrestrial arthropods and insects (Stebbins 1951).

In order to avoid desiccation animals are present under surface cover only when soils are adequately moist. This limits their active season to April to November, with lower elevation animals being active only through May or June. Animals have been collected as early as January indicating that these animals may be active during cold winter months (Hansen and Wake 2005). These salamanders are primarily nocturnal, retreating to burrows or cover objects during the day (Kucera 1997).

Batrachoseps salamanders have very small home ranges. Studies on a similar localized California slender salamander (B. attenuatus) found that adult salamanders moved an average of approximately 5 feet over a two year period and were found to repeatedly use the same cover object (Kucera 1997). In general slender salamanders demonstrate high site fidelity and rarely move more than 5-10 meters over their lifetime (Cunningham 1960). Home range and territoriality in this species has not been studied. Presumably territories are small or seasonal as breeding sized Batrachoseps spp. are often found under the same cover object in close proximity to one another (Kucera 1997).

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Figure 12. Map of relictual slender salamander locations from multiple sources, 2010

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Figure 13. Map of relictual Slender Salamander locations from the NRIS databases, 2016

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Little is known about the reproduction of this species other than their eggs are laid terrestrially (Hansen and Wake 2005). Presumably reproduction is similar to other slender salamanders with 13 to 20 eggs laid where the young emerge fully formed during winter and early spring. The clutch size of this species is unknown as egg deposition sites are unknown (Kucera 1997).

Threats This species is most threatened by degradation to its habitat. Habitat alteration is most likely the cause of presumed extirpation at the type location of this species. At this site in the Lower Kern River Canyon no animals have been found since April 1970. There has been extensive development and change in land use patterns since the late 1960’s which are likely a major factor in at least eight localized population extirpations (Jennings and Hayes 1994). As this species is more closely tied with water than other Batrachoseps species, impacts that affect water quantity or quality likely have a significant impact. Additionally habitat degradation as a result of road maintenance and related construction activities have likely significantly impacted this species (Hansen and Wake 2005). Although fire may impact this species, the impact of the fire itself is likely minimal as these animals spend a majority of their time in vicinity of moist microhabitats or underground. Impacts from fire are more likely indirect impacts caused by fire control mechanisms such as fire retardant use and bulldozers.

Additional threats to this species include natural predators and disease. Although Batrachochytrium dendrobatidis has been documented on the widespread California slender salamander (Batrachoseps attenuatus), the actual impacts of chytridomycosis on this species is unknown. However, any impacts are likely greater in the relictual slender salamander than other slender salamanders as this species spends more time in water (Weinstein 2009). Natural predators of this species likely include: spotted and striped skunks, ringtails, raccoons, gray foxes, ring-necked snakes, and various skinks, moles and shrews (Kucera 1997).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Relictual slender salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

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Determination Statement Inyo and Sierra NFs: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on these forests.

Sequoia NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Literature Cited – Relictual Slender Salamander Cunningham, J.D. 1960. Aspects of the ecology of the Pacific Slender Salamander, Batrachoseps pacificus, in southern California. Ecology 41: 88-99.

Hammerson, G. 2004. Batrachoseps relictus. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. http://www.iucnredlist.org. Downloaded on 31 May 2012.

Hansen, R.W. and Wake, D.B. 2005. Batrachoseps relictus Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 31 May 2012.

Jennings, M. R. and Hayes, M. P. 1994. Relictual slender salamander Batrachoseps relictus Brame and Murray 1968. In: Amphibian and Reptile Species of Special Concern in California. California Department of Fish and Game, Sacramento, California.

Jockusch, E.L., Martinez-Solano, I., Hansen, R.W., and Wake, D.B. 2012. Morphological and molecular diversification of slender salamanders (Caudata: Plethodontidae: Batrachoseps) in the southern Sierra Nevada of California with descriptions of two new species. Zootaxa 3190: 1-30.

Kucera, T. 1997. Relictual Slender Salamander Batrachoseps relictus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Stebbins, R.C. 2003. A field guide to western reptiles and amphibians. R.T. Peterson. Houghton Mifflin Company, New York, New York.

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Kern Canyon Slender Salamander (Batrachoseps simatus) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The Kern Canyon slender salamander is a small salamander endemic to the lower Kern River Canyon in Tulare and Kern Counties in California. The elevational range of this species is from 305 to 1220 meters (Morey and Basey 1990). All known populations occur within Region 5, and most known populations occur on the Sequoia National Forest. However, total adult population is unknown although individual population sizes are likely very small but locally abundant (Hammerson 2010).

Figure 14 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Very little is known about the life history of the Kern Canyon slender salamander. Most of the information below is based on information from other presumably similar Batrachoseps salamanders.

Kern Canyon slender salamanders are found on north facing slopes in valley-foothill hardwood, valley- foothill hardwood-conifer and mixed chaparral habitats (Morey and Basey 1990). Specifically these salamanders inhabit small wooded tributary canyons, oak-pine communities on slopes, willow and cottonwood communities near streams, chaparral, and potentially grasslands adjacent to forests. Salamanders are typically found in crevices in talus slopes or under rocks and logs (Hammerson 2010). During dry times of the year, animals retreat to moist underground areas or move to seepage areas (Morey and Basey 1990).

Feeding likely occurs opportunistically both above and below ground. Foraging primarily occurs under surface objects such as bark, talus rocks or leaf litter, but may also occur in termite tunnels and earthworm burrows (Morey and Basey 1990). The diet is likely similar to other Batrachoseps salamanders including earthworms, small slugs, various terrestrial arthropods and insects (Stebbins 1951).

Slender salamanders have very small home ranges. Studies on a similar localized California slender salamander (B. attenuatus) found that adult salamanders moved an average of approximately 5 feet over a two year period and were found to repeatedly use the same cover object (CWHR 2005). In general Batrachoseps salamanders demonstrate high site fidelity and rarely move more than 5 to 10 meters over their lifetime (Cunningham 1960). Home range and territoriality in this species has not been studied. Presumably territories are small or seasonal as breeding sized Batrachoseps spp. are often found under the same cover object in close proximity to one another (CWHR 2005).

Little is known about the reproduction of this species other than their eggs are laid terrestrially. Presumably reproduction is similar to other Batrachoseps salamanders with 13 to 20 eggs laid where the young emerge fully formed during winter and early spring (Stebbins).

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Figure 14. Map of Kern Canyon slender salamander from multiple sources, 2010

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Threats As this salamander is not well studied, the general lack of knowledge about this animal’s ecology is likely the most significant threat to its persistence. Without knowing more about its basic ecology it will be challenging to manage this species into the future. As these are small populations, stochastic events such as fire, flash flooding, and climate change are likely significant risks to the species. Presumably habitat degradation is a major threat to this species as a result of development, timber harvest, fire suppression activities, mining, grazing and direct damage to localized sites. Although fire may impact this species, the impact of the fire itself is likely minimal as these animals spend a majority of their time in vicinity of moist microhabitats or underground. Impacts from fire are more likely indirect impacts caused by fire control mechanisms such as fire retardant use and bulldozers.

Additional threats to this species include natural predators and disease. Although Batrachochytrium dendrobatidis has been documented on the widespread California slender salamander (Batrachoseps attenuatus), the actual impacts of chytridomycosis on this species is unknown (Weinstein 2009). Natural predators of this species likely include: small snakes including ring-necked snakes, beetle larvae and other predatory arthropods, diurnal birds and small mammals (Morey and Basey 1990).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Kern Canyon slender salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

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Literature Cited – Kern Canyon Slender Salamander Cunningham, J.D. 1960. Aspects of the ecology of the Pacific Slender Salamander, Batrachoseps pacificus, in southern California. Ecology 41: 88-99.

CWHR Program Staff. 2005. Kern Plateau Salamander Batrachoseps robustus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Hammerson, G. 2010. Batrachoseps simatus. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. http://www.iucnredlist.org. Downloaded on 31 May 2012.

Morey, S. and Basey, M. 1990. Kern Canyon slender salamander Batrachoseps simatus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Stebbins, R.C. 2003. A field guide to western reptiles and amphibians. R.T. Peterson. Houghton Mifflin Company, New York, New York.

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Fairview Slender Salamander (Batrachoseps bramei) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The Fairview slender salamander is a recently described species. Genetic work published in 2012, coupled with a difference in morphology and range split this species from the relictual slender salamander (Batrachoseps relictus). All known locations for the Fairview slender salamander are on the Sequoia National Forest and it is likely that the entire species range is within the Sequoia National Forest (Jockusch et al. 2012).

This salamander has a very small range within the Upper Kern River Canyon and along the western shore of Lake Isabella. The southernmost population is from Wofford Heights on the west side of the Kern River approximately 2 kilometers south of the Cannell Creek drainage on the eastern side of the Kern River. The southern extent of this species range is near the junction of the main and south forks of the Kern River. The northern extent is at least 1 kilometers north of the junction of South Falls Creek and the Kern River. The northern portions of this range have not been thoroughly surveyed as a result of the remote location. The elevational range for this species is 860 to 1,280 meters (Jockusch et al. 2012).

Fairview slender salamanders are not found in association with any other slender salamanders. They approach the Kern Canyon slender salamander (Batrachoseps simiatus) in the Lower Kern River Canyon; however these populations are separated by the Kern River as well as inhospitable xeric terrain. Additionally, they approach the Kern Plateau slender salamander (B. robustus) and Greenhorn Mountains slender salamanders (B. altasierrae) but these salamanders are found at different elevations. The population status of this species is unknown; however these animals appear to be somewhat abundant and can be found consistently throughout their range (Jockusch et al. 2012).

Figure 15 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status The entire population of Fairview slender salamanders is found on an uplifted ridge of metamorphic rocks paralleling the Kern River. Localized populations are found on north-facing slopes and talus amongst chaparral. This chaparral plant community consists of the following genera: Ceanothus, Arctostaphylos, Ribes, Chrysothamnus, Pinus and Quercus. However, some individuals have been found in other habitat including open sandy flood plains, grasslands, on the riverbank, and in protected groves. Most animals have been found under rocks at the base of talus slopes, but they also have been found under logs, in gravel and in leaf litter (Jockusch et al. 2012).

Feeding likely occurs opportunistically both above and below ground (Kucera 1997). The diet is likely similar to other Batrachoseps salamanders which include earthworms, small slugs, various terrestrial arthropods and insects (Stebbins 1951). The water requirements are unknown, but presumably these salamanders require moist microhabitats.

Presumably surface activity is limited to the rainy season from November to April (Kucera 1997). These salamanders have been observed active at night and are typically found active in spring. Fairview slender salamanders have been collected under cover objects with snow on the ground to temperatures as low as 35 degrees Fahrenheit (Jockusch et al. 2012).

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Figure 15. Map of Fairview slender salamander locations from multiple sources, 2010

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Slender salamanders have very small home ranges. Studies on a similar localized California slender salamander (B. attenuatus) found that adult salamanders moved an average of approximately 5 feet over a two year period and were found to repeatedly use the same cover object (Kucera 1997). In general slender salamanders demonstrate high site fidelity and rarely move more than 5 to 10 meters over their lifetime (Cunningham 1960). Territoriality in this species has not been studied. Presumably territories are small or seasonal as breeding sized Batrachoseps spp. are often found under the same cover object in close proximity to one another (Kucera 1997).

Little is known about the reproduction of this species. Presumably reproduction is similar to other Batrachoseps salamanders where eggs are laid terrestrially where the young emerge fully formed. The clutch size of this species is unknown as egg deposition sites are unknown. Presumably due to this species’ association with talus slopes the eggs are deposited deep within the talus piles where temperatures and humidity levels are stable. It is likely that courtship and oviposition occurs following the first heavy rains in the fall (Stebbins 2003).

Threats As this species has a very limited range they are vulnerable to stochastic events such as floods, fire and climate change. However, habitat loss or degradation is likely the main threat to this species as it is with many of the other limited range Batrachoseps species. This is especially true as some of these populations are in close proximity to Mountain Highway 99 and exist in the road-edge habitat. Although fire may impact this species, the impact of the fire itself is likely minimal as these animals spend a majority of their time in vicinity of moist microhabitats in talus or underground. Impacts from fire are more likely indirect resulting from fire control mechanisms such as fire retardant use and heavy equipment (Jockusch et al. 2012).

Additional threats to this species include natural predators and disease. Although Batrachochytrium dendrobatidis has been documented on the widespread California slender salamander (Batrachoseps attenuatus), the actual impacts of chytridomycosis on this species is unknown (Weinstein 2009). Natural predators of this species likely include: spotted and striped skunks, ringtails, raccoons, gray foxes, ring- necked snakes, and various skinks, moles and shrews (Kucera 1997).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan (species listed as B. relictus).

Alternative B, C, and D: Fairview slender salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

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Literature Cited – Fairview Slender Salamander Cunningham, J.D. 1960. Aspects of the ecology of the Pacific Slender Salamander, Batrachoseps pacificus, in southern California. Ecology 41: 88-99.

Kucera, T. 1997. Relictual slender salamander Batrachoseps relictus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Stebbins,R.C. 2003. A field guide to western reptiles and amphibians. R.T. Peterson. Houghton Mifflin Company, New York, New York.

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Kings River Slender Salamander (Batrachoseps regius) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species account Kings River slender salamanders are limited in distribution to the lower reaches of the Kings River in Fresno County, California (Jockusch et al. 1998). The elevational range of this species is from 335 to 2,470 meters. There have not been any formal population studies for this species, however populations appear stable. The majority of the known range of this species is restricted to Region 5 lands on the Sierra and Sequoia National Forests with the remainder of the population in Kings Canyon National Park (Hammerson 2004).

Genetic analysis split this species from Batrachoseps pacificus relictus in 1998. Both gregarious slender salamander (B. gregarious) and relictual slender salamander (B. relictus) have been found near the range boundary of this species, however no known sympatry with other slender salamanders occurs (Jockusch et al. 1998). Kings River slender salamanders can be found in sympatry with Sierra Nevada ensatina (Ensatina eschscholtzii platensis) and California newt (Taricha torosa) (Jockusch et al. 1998; Hansen and Wake 2005).

Figure 16 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Kings River slender salamanders are found along streams and moist canyons, in valley foothill riparian habitat, blue oak woodland and mixed conifer woodland (Kucera 2005). This type habitat for this species is well-shaded, mixed chaparral on north-facing slopes. Vegetation in the area included Aesculus californicus, Umbellularia californica, Quercus wislizenii, Pinus ponderosa, Pinus sabiniana, and Quercus douglasii. Animals utilizing cover objects and are often found under rocks in talus near roadsides (Jockusch et al. 1998).

Feeding likely occurs opportunistically both above and below ground (Kucera 1997). The diet is likely similar to other Batrachoseps salamanders including earthworms, small slugs, various terrestrial arthropods and insects (Stebbins 2003).

In order to avoid desiccation animals are present under surface cover only when soils are adequately moist. This limits their active season to April to November, with lower elevation animals being active only through May or June. Slender salamanders are primarily nocturnal, retreating to burrows or cover objects during the day (Kucera 2005).

Batrachoseps salamanders have very small home ranges. Studies on a similar localized California slender salamander (B. attenuatus) found that adult salamanders moved an average of approximately 5 feet over a two year period and were found to repeatedly use the same cover object (Kucera 1997). In general Batrachoseps salamanders demonstrate high site fidelity and rarely move more than 5-10 meters over their lifetime (Cunningham 1960). Home range and territoriality in this species has not been studied. Presumably territories are small or seasonal as breeding sized Batrachoseps spp. are often found under the same cover object in close proximity to one another (Kucera 1997).

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Figure 16. Map of Kings River slender salamander locations from multiple sources, 2010

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Little is known about the reproduction of this species other than their eggs are laid terrestrially (Hansen and Wake 2005). Presumably reproduction is similar to other Batrachoseps salamanders with 13 to 20 eggs laid where the young emerge fully formed during winter and early spring. The clutch size of this species is unknown as egg deposition sites are unknown (Kucera 1997).

Threats As this species has a very limited range they are vulnerable to stochastic events such as floods, fire and climate change. Additionally, it appears that the known populations have been in isolation from each other for a long time further compounding the risk of stochastic events (Hansen and Wake 2005). Habitat loss or degradation is another threat to this species as it is with many of the other limited range Batrachoseps species (Jockusch et al. 2012). Many of the known sites for this species occur directly adjacent to a road and are at risk for impacts associated with road construction and maintenance (Hansen and Wake 2005). Although fire may impact this species, the impact of the fire itself is likely minimal as these animals spend a majority of their time in vicinity of moist microhabitats or underground. Impacts from fire are more likely indirect resulting from fire control mechanisms such as fire retardant use and heavy equipment (Jockusch et al. 2012).

Additional threats to this species include natural predators and disease. Although Batrachochytrium dendrobatidis has been documented on the widespread California slender salamander (Batrachoseps attenuatus), the actual impacts of chytridomycosis on this species is unknown (Weinstein 2009). Natural predators of this species likely include: spotted and striped skunks, ringtails, raccoons, gray foxes, ring- necked snakes, and various skinks, moles and shrews (Kucera 1997).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of aquatic areas.

Alternative B, C, and D: Kings River slender salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

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Sequoia NF: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species is only located within the National Monument.

Sierra NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Literature Cited – Kings River Slender Salamander Cunningham, J.D. 1960. Aspects of the ecology of the Pacific Slender Salamander, Batrachoseps pacificus, in southern California. Ecology 41: 88-99.

Hammerson, G. 2004. Batrachoseps regius Amphibiaweb account. In: IUCN 2012. IUCN Red List of Threatened Species. Version 2012.1. http://www.iucnredlist.org. Downloaded on 6 August 2012.

Hansen, R.W. and Wake, D.B. 2005. Batrachoseps regius Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 6 August 2012.

Jockusch, E.L., Wake, D.B., and Yanev, K.P. 1998. New species of slender salamanders, Batrachoseps (Amphibia: Plethodontidae), from the Sierra Nevada of California. Contributions in Science 472: 1-17.

Kucera, T. 2005. Kings River slender salamander Batrachoseps regius. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Stebbins, R.C. 2003. A field guide to western reptiles and amphibians. R.T. Peterson. Houghton Mifflin Company, New York, New York.

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Black Toad (Anaxyrus exsul) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species account The black toad is found over a range of only fifteen hectares and has one of the smallest ranges of any amphibian worldwide (Fellers 2005). Black toads are endemic to the vicinity of Deep Springs Valley, Inyo County, California (Figure 17). The majority of the species occurs on land owned by Deep Springs College (Murphy et al. 2003). The Bureau of Land Management also owns property that supports a population of black toads (Hammerson 2004). These toads are only known from the vicinity of Corral Springs, Buckhorn Springs, Bog Mound Springs and Antelope Springs. Antelope Springs is approximately 5 kilometers north of the other springs and is situated on a hillside near the Inyo National Forest (Wang 2009). It is possible but unlikely that the species is found in similar habitat in areas adjacent to the known populations.

When black toads were first discovered the population was estimated at 700 total animals (Myers 1942). However, a more thorough census in 1978 found the population to be between 7,897 and 9,744 animals. Later, a re-census in 1999 found a similar population of 8,419 with an error of plus or minus 2,795 animals (Murphy et al. 2003). The population may be as high as 24,000 animals as there are unsurveyed areas in the vicinity of Bog Mound Springs and Buckhorn Springs (Hammerson 2004). Although the Antelope Springs population was thought to be introduced, genetic analysis does not support this theory. Genetic analysis also found that the effective population size of this species is very small (Antelope: 7 to 14, Bog Mound: 7 to 16, Corral: 24 to 30; and Buckhorn: 23 to 28), but gene flow does occur between springs less than 1.5 kilometers apart (Wang 2009).

In the NRIS database, the Inyo NF has 6 records, Sequoia and Sierra NFs have no records. Figure 18 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status This toad is extremely aquatic, and is rarely found out of water. They are restricted to wet areas in the vicinity of permanent springs (Myers 1942). Each of the four sub-populations are isolated by at least 1.5 kilometers of arid desert scrub (Wang 2009). The immediate riparian habitat varies; the valley springs are typically unshaded, but the Antelope Springs site is heavily shaded by Salix spp. and Chrysothamnus spp. This shading alters daily air temperatures where shaded areas are typically 8 degrees Celsius cooler than the open sites (Schuierer 1963). Water temperatures at the spring origin are typically 20 degrees Celsius with a pH level of 7. At several springs hydrogen sulphide is present in the water but this does not appear to impact the toads (Schuierer 1961). Both larvae and juvenile animals seem to prefer shallow areas and adults preferring cooler, deeper water (Schuierer 1961).

These toads consist of approximately 77 percent water, and can survive desiccation of up to 40 percent by weight. Black toads do not appear to drink surface water, but instead have the ability to rapidly absorb water through their skin (Straw 1958). Foraging typically occurs close to water, but adults have been found foraging up to twelve meters from water. When weather is hot, foraging occurs primarily in the early and late hours of the day. Diet consists of approximately 80 percent Hymenoptera, 10 percent Diptera and 10 percent various other invertebrates (Schuierer 1961).

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Figure 17. Distribution map of black toad from 2010 from multiple sources

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Figure 18. Map of black toad from NRIS Databases, 2016

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Both tadpoles and adults can survive high temperatures, the aquatic critical thermal maximum is approximately 104 degrees Fahrenheit. Adults can survive air temperatures up to 114.8 degrees Fahrenheit if means of evaporative cooling are available (Straw 1958). Daily activity patterns are primarily diurnal, but nocturnal activity occurs during warmer summer nights (Schuierer 1961, Schuierer 1963, Murphy et al. 2003).

The critical thermal minimum of black toads is 24.8 degrees Fahrenheit (Brattstrom 1968). Animals typically enter hibernation in November, but may remain active during this month. Hibernation occurs in rodent burrows approximately one to two meters above the water line, with animals clustering into groups of 20 to 30 animals to retain body heat. Females typically hibernate near the spring source and males usually hibernate near their most recent breeding site (Murphy et al. 2003). Emergence begins in March and coincides with the start of the breeding season (Morey 1990). Home range and territoriality have not been evaluated in this species (Fellers 2005).

Breeding is typically started in late March and is usually completed by late April. Toads gather in large numbers at breeding sites with neither sex gathering before the other. Temperature seems to have little effect on the breeding start date; instead day length seems to be the primary trigger (Schierer 1961). Clutches are typically 120 to 150 eggs (Morey 1990). Hatching can occur as quickly as 4 to 5 days at 68 degrees Fahrenheit with a larval period of three to five weeks. Metamorphosed individuals are typically seen in early June and may be sexually mature as early as the second year (Schierer 1961; Morey 1990). Multiple male amplexus clusters have been reported, but only in modified, artificial canals and not in the more typical marsh habitat (Murphy et al. 2003).

Threats The black toad is a relictual species that was once abundant during the moister Pleistocene. Over time the area dried and become one of the most arid regions in North America. Black toads then became restricted to the few perennial water sources left in the Deep Springs Valley (Myers 1942).

There are two major potential future threats to this species, both of which are a factor of the extremely limited range. As these toads were isolated from the western toad (Bufo boreas) by the drying of the desert over the past 10,000 years, further changes which favor drying may cause spring failure and population extinction. Additionally, amphibian chytridiomycosis has not been found in this species but poses a huge threat to this densely packed, highly aquatic species (Murphy et al. 2003).

Until the late 1960’s water diversion and cattle trampling resulted in notable mortality to this species (Murphy et al. 2003). Annual recanalization of the water ways caused population numbers to decrease from 3,600 in 1954 to 1,200 in 1961 in one population (Schuierer 1961). Efforts were undertaken by the main landowner (Deep Springs College) to mitigate for these mortalities by the use of cattle exclosures and stopping agricultural activities which impacted the toad. Although the cessation of agricultural water use has clearly benefited the toads, cattle exclosures may actually negatively impact the toad. In areas where cattle have been permanently excluded, there is less open water and dense stands of aquatic vegetation. This may impact the toad’s utilization of the habitat, but specific impacts are unknown. Further, it appears that winter grazing does not directly impact the hibernating toads, and may in fact benefit toad populations by keeping vegetation low (Murphy et al. 2003).

Other potential impacts to this species includes: collection for scientific research, illegal collection for the pet trade, and predators. Introduced carp (Cyprinus carpio) are present in the Corral Springs watershed, but their impacts on black toads is unknown (Murphy et al. 2003). Natural predators are not well documented, but likely include aquatic invertebrates, birds and mammals (Morey 1990). Common ravens (Covus corax) have been observed eviscerating adult toads (Fellers 2005).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will provide for the protection of the riparian zone that supports this species.

Alternative B, C, and D: Black toads were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Removal of invasive species is proposed across each alternative, thus providing a reduced risk to the toad.

Literature Cited – Black Toad Brattstrom, B.H. 1968. Thermal acclimation in anuran amphibians as a function of latitude and altitude. Comp. Biochem. Physiol. 24: 93-111.

Fellers, G.M. 2005. Bufo exsul Meyers 1942(a) Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 26 June 2012.

Hammerson, G. 2004. Anaxyrus exsul. In: IUCN 2012. IUCN Red List of the Threatened Species. Version 2012.1. http://www.iucnredlist.org. Downloaded on 26 June 2012.

Morey, S. 1990. Black toad Bufo exsul. In Zeiner, D.C., W.F.Laudenslayer, Jr., K.E. Mayer, and M. White, eds. 1988-1990. California's Wildlife. Vol. I-III. California Department of Fish and Game, Sacramento, California.

Murphy, J.F., Simandle, E.T. and Becker, D.E. 2003. Population status and conservation of the black toad, Bufo exsul. The Southwestern Naturalist, 48(1): 54-60.

Myers, G.S. 1942. The black toad of Deep Springs Valley, Inyo County, California. Occassional Papers of the Museum of Zoology, University of Michigan 460: 1-13.

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Schuierer, F.W. 1961. Remarks upon the natural history of Bufo exsul Myers, the endemic toad of Deep Springs Valley, Inyo County, California. Herpetologica 17(4): 260-266.

Schuierer, F.W. 1963. Notes on two populations of Bufo exsul Myers and a commentary on speciation within the Bufo boreas group. Herpetologica 18(4): 262-267.

Straw, R.M. 1958. Experimental notes on the Deep Springs toad Bufo exsul. Ecology 39(3): 552-553.

Wang, I.J. 2009. Fine-scale population structure in a desert amphibian: landscape genetics of the black toad (Bufo exsul). Molecular Ecology 18: 3847-3856.

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Limestone Salamander (Hydromantes brunus) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species account Limestone salamanders are an endemic salamander species found in a small area in Mariposa County, California (Basey and Morey 2000). The total known extent of this species range is approximately sixteen to seventeen kilometers in length along the Merced River. Specifically, this salamander occurs from the vicinity of the type locality on state route 140 west to Hell Hollow and slightly up the North Fork of the Merced River (Wake and Papenfuss 2005).

There are two conservation areas created to protect the state threatened limestone salamander. The first is the Limestone Salamander Ecological Reserve managed by the California Department of Fish and Game which protects 120 acres of habitat including the type location. The second, the Limestone Salamander Area of Critical Environmental Concern, is managed by the Bureau of Land Management which consists of 1600 acres of both confirmed and potential habitat (Hammerson and Wake 2004).

Within Region 5 this salamander is found in foothill areas in the Sierra and Stanislaus National Forests. Although few localities are known from within these forests, contiguous, suitable habitat exists along the North, Middle and South Forks of the Merced River and may contain additional populations (Figure 19).

Other salamander species which can be found in sympatry include arboreal salamanders (Aneides lugubris), Hell Hollow slender salamanders (Batrachoseps diabolicus), yellow-eyed ensatinas (Ensatina eschscholtzii xanthoptica) and Sierra newts (Taricha torosa sierrae). Although there have been no formal population studies published on this species, there is no indication that the extent or density of this species has changed from the historical values (Wake and Papenfuss 2005).

Habitat Status These salamanders are typically found in association with limestone. They can also be found under slate slabs, irregularly shaped limestone pieces, moss-covered and barren talus, in rock crevices and in abandoned mine tunnels. Typically animals are found on steep slopes, especially those which are north and east-facing, but can be found on level ground as well (Basey and Morey 2000; Wake and Papenfuss 2005). Vegetation at these sites is either mixed chaparral or gray pine-oak woodland (Hammerson and Wake 2004). At the type location the dominant flora consists of: digger pine, toyon, California laurel, manzanita, chamise, buck brush, yerba santa, phacelia, and California wood fern (Gorman 1954). California buckeye may serve as an indicator species for optimal habitat for limestone salamanders (Basey and Morey 2000).

Limestone salamanders presumably feed on insects and other small invertebrates (Basey and Morey 2000). In captivity limestone salamanders have eaten Batrachoseps salamanders (Gorman 1954). Water requirements are unknown; however water needs are probably met by rain and subterranean sources (Basey and Morey 2000).

Limestone salamanders are active on the surface when soil is moist and air temperatures are cool. This limits surface activity to winter and early spring, however animals likely remain active underground throughout the year. The holotype of this species was collected in February and animals have been observed active in mine shafts in July. Observations have been made at temperatures ranging from 10 to 14 degrees Celsius with an average temperature of 11.4 degrees Celsius. Home range size and territoriality in this species remains unknown (Wake and Papenfuss 2005).

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Figure 19. Map of limestone salamander based on multiple sources, 2010

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Little is known about reproduction in this species. Reproduction is terrestrial with eggs likely laid in cracks and crevices below the surface in limestone talus. Egg development has not been observed, but is likely direct development like similar species in the genera. Ovarian eggs were enlarged in the holotype in late February. Eggs are probably laid in late spring (Wake and Papenfuss 2005).

Threats As with other species with a limited range, stochastic events are a significant threat to the persistence of this species. Events such as fire, flood, disease, habitat alteration, or climate change can significantly impact a limited range animal. Fire likely has only minimal impact to this species, however fire suppression activities may disturb habitat. No studies have investigated the impact of Batrachochytrium dendrobatidis on this species; however its highly terrestrial lifecycle puts it less at risk for serious impact. Habitat alteration such as development for mining, road widening or construction, limestone quarrying and dam building likely pose the greatest threat to this species (Hammerson and Wake 2004). As few studies have investigated this species, additional research needs to be conducted to determine what threats are most significant for this species.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Limestone salamander were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species.

Determination Statement Sierra NFs: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Inyo and Sequoia NFs: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on these forests.

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Literature Cited – Limestone Salamander Basey, H. and Morey, S. 2000. Limestone Salamander Hydromantes brunus. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Gorman, J. 1954. A new species of salamander from central California. Herpetologica 10(3) 153-158.

Hammerson, G. and Wake, D. 2004. Hydromantes brunus. In: IUCN 2011. IUCN Red List of Threatened Species. Version 2011.2. http://www.iucnredlist.org. Downloaded on 4 June 2012.

Wake, D.B. and Papenfuss, T.J. 2005. Hydromantes brunus Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 4 June 2012.

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Foothill Yellow-legged Frog (Rana boylii) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account Foothill yellow-legged frogs have suffered significant population declines across the majority of the known range. Historically this frog was found across most of southwestern Oregon west of the Cascades Mountains crest south through California to Baja California (Fellers 2005, Jennings and Hayes 1994). Specimens collected from the Sierra San Pedro Martir of Baja California in 1961 were lost in transit and represented a population almost 300 miles south of the nearest known population (Loomis 1965). The foothill yellow-legged frog is found in most of northern California west of the Cascade Mountains crest, in the Coast Ranges from the California-Oregon border south to the Transverse Mountains in Los Angeles County and along the western slope of the Sierra Nevada Mountains south to Kern County (see Figure 20 and Figure 21). Isolated populations have been reported from the San Joaquin Valley and the mountains in Los Angeles County. This frog can be found from near sea level to 6,370 feet where habitat is suitable (Morey 2000). Within Region 5 this frog is found on, or could occur on, all national forests except for the Cleveland, Inyo, Modoc, and Lake Tahoe Basin National Forests.

Populations of foothill yellow-legged frogs in the Pacific Northwest are considered to be the most stable with approximately 40 percent of streams occupied, 30 percent are occupied in the Cascade Mountains, 30 percent in the south Coast Range south of San Francisco and 12 percent in the Sierra Nevada foothills (Fellers 2005). Populations in and south of the Tehachapi Mountains have probably been extirpated (Santos-Barrera et al. 2004). The last verifiable record from this area is a series of animals which were collected in 1970 however unverifiable observations occurred through the late 1970’s (Jennings and Hayes 1994). Any remaining populations in Mexico are protected by Mexican law under the “Special Protection” category (Santos-Barrera et al. 2004). While there are no recognized subspecies of foothill yellow-legged frogs, recent genetic studies indicate that there is a genetic break along the transverse mountains (Lind et al. 2011). Although there are numerous occupied streams, only 30 of the 213 known populations in California have populations of at least 20 individual adults. These frogs are most numerous in the northern coast range with six populations of at least 100 adults and an additional nine populations of at least 50 adults (Fellers 2005).

In the NRIS database, the Inyo NF has 2 records, Sequoia NF has no records, and the Sierra NF has 2,258 records. Figure 22 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Foothill yellow-legged frogs are found in partially shaded rocky streams in a variety of habitats including: valley-foothill hardwood, valley-foothill hardwood-conifer, valley-foothill riparian, ponderosa pine, mixed conifer, coastal scrub, mixed chaparral and wet meadows and appear to be highly dependent on free water for all life stages (Morey 2000).

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Figure 20. Map of foothill yellow-legged frog locations in the southern range based on multiple sources, 2010

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Figure 21. Map of foothill yellow-legged frog locations in the central range based on multiple sources, 2010

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Figure 22. Map of foothill yellow-legged frog in the NRIS Databases, 2016

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The habitat characteristics of non-breeding adult foothill yellow-legged frogs have not been fully evaluated. Leidy et al. (2009) observed a group of six adults aggregated on a vertical ledge over a meter away from water in late summer. These animals were easily approached and did not respond to touch. This suggests that foothill yellow-legged frogs become inactive during late summer and autumn. The location of this aggregation also indicates that adults may migrate up tributaries consisting of large-sized boulders and bedrock to utilize the cooler air and water temperatures, and to avoid predators and high water flows (Leidy et al. 2009). Overwintering behavior is completely unknown, but adults are commonly found in tributaries prior to being found in the mainstem waterway. They are rarely seen more than a few meters away from water, but it remains unknown if they utilize upland areas during winter months (Kupferberg 1996). Habitat use of juvenile frogs is also largely unknown. Some evidence indicates that they potentially use smaller waterways such has springs or small tributary streams (Lind et al. 2011).

Breeding habitat is typically classified as a stream with riffles containing cobble-sized or larger rocks as substrate (Morey 2000). These streams are further defined by having low-water velocities near tributary confluences in shallow reaches and are wider and shallower than non-breeding sites, have emergent rocks and are typically asymmetrical with cobble or small boulder bars (Wheeler and Welsh 2008; Kupferberg 1996). Egg attachment sites are usually cobbles or boulders, but frogs may sometimes utilize bedrock or vegetation. These sites are often on the lee side of rocks or beneath overhangs such that the site has a narrow range of low-water velocity. Coarse sediment enables frogs to choose the best oviposition site to shield egg masses from high-flows. The reproductive strategy of the foothill yellow-legged frog is well suited to rivers with predictable winter flooding and summer droughts (Kupferberg 1996).

Wheeler and Welsh (2008) found that approximately 68 percent of adult male foothill yellow-legged frogs in their study were site faithful. These animals had an average breeding home range of 0.58 square meters and home range size was directly linked with the frequency of aggressive behavior and calling activity. Males were not actively guarding future oviposition sites, but were guarding a specific, but generalized, patch of habitat within the breeding site (Wheeler and Welsh 2008).

Larval foothill yellow-legged frogs primarily consume algae and will preferentially graze on epiphytic diatoms as this food item allows them to grow more rapidly (Jennings and Hayes 1994). Post-metamorphs likely consume both aquatic and terrestrial insects but there is little research on the subject (Jennings and Hayes 1994). Adult diet is thought to include: flies, moths, hornets, ants, beetles, grasshoppers, water striders and snails with a terrestrial arthropod composition of 87.5 percent insects and 12.6 percent arachnids (Fellers 2005).

Breeding can occur as early as April 7 but may start as late as May 8 and typically continues at least a month with an average duration of 49.5 days between first and last egg depositions (Wheeler and Welsh 2008, Kupferberg 1996). Breeding occurs earlier in low-base flow years and begins when stream flow is at or below 0.6 meters per second and between 0.04 and 0.17 meters per second at the microhabitat scale (Wheeler and Welsh 2008). Eggs are typically laid in shallow areas ranging from 4 to 43 centimeters at varying distances from shore. When base-flow is low frogs will oviposit further from shore (Kupferberg 1996, Lind et al. 1996). Prior to egg deposition but while in amplexus, females will scrape potential attachment sites with their hind-feet in order to remove any debris and make egg adhesion stronger. This reduces the likelihood of the clutch being detached by a change in water velocity (Rombough and Hayes 2005). Females lay a single annual clutch of between 300 and 2,000 eggs (Jennings and Hayes 1994, Kupferberg 1996). Reproductive output is typically 18.8 clutches per breeding site, with a measurement error of plus or minus 1.9 clutches (Kupferberg 1996). The critical thermal maximum for embryos is approximately 26 degrees Celsius, however eggs are typically found from 9 to 21.5 degrees Celsius (Jennings and Hayes 1994, Kupferberg 1996). Incubation lasts approximately two weeks (5 to 37 days) depending on water temperature and position within the clutch. Eggs near the attachment point and eggs

168 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests in the center of the clutch typically hatch later than eggs on the periphery of the clutch (Kupferberg 1996, Fellers 2005). After hatching, tadpoles move away from the egg mass. As with egg development, larval development is temperature dependent with metamorphosis typically occurring 3 to 4 months after hatching with no documented overwintering of larvae. Foothill yellow-legged frogs metamorphose at a size of 1.4 to 1.7 centimeters in length. Reproductive maturity is thought to occur the second year after metamorphosis, but can occur as early as six months after metamorphosis. Longevity for this species is unknown (Fellers 2005).

Threats High mortality in this species occurs during the egg and larval life stages. The main causes of mortality in eggs are hydrologic in nature. Eggs are usually killed by either desiccation or scour (Kupferberg 1996, Lind et al. 1996). Tadpole mortality can also occur as a result of irregular stream flows. The main critical velocity for tadpoles is 20 centimeters per second but flows as low as 10 centimeters per second can displace large tadpoles. This results in slower growth and development, greater exposure to predators and possible mortality. The seasonal pulses of high water flows used in many regulated rivers have a significant negative impact on recruitment for this species (Kupferberg et al. 2011).

Loss of genetic diversity due to habitat loss is a major threat to foothill yellow-legged frogs. Populations which are more than 10 kilometers apart are prone to genetic drift and barriers such as dams or habitat fragmentation may prevent dispersal between isolated populations (Dever 2007). In one study area, 94 percent of downstream bar habitat and potential breeding habitat was lost after the installation of a dam. The encroachment of riparian vegetation created stable sandy berms which caused the river to become narrower and deeper and thus unsuitable for use by foothill yellow-legged frogs (Lind et al. 1996).

Pesticides can impact these frogs in both original and derived forms. Chloroxon (the oxon derivative of chlorpyrifos) killed all tadpoles exposed to it in Sparling and Fellers (2007) study and was at least 100 times more lethal than the parent chemical. Air-borne pesticides are implicated as the most significant threat to this species, especially for Sierra Nevada populations which are directly impacted by pesticide drift from the central valley (Fellers 2005).

Predation by non-native, introduced fishes is a major threat to this species. Smallmouth bass (Micropterus dolomieu) in particular readily consume both larvae and adult frogs and are capable of directly affecting populations of foothill yellow-legged frogs. As foothill yellow-legged frog larvae are unable to avoid predation by novel predators, populations are prone to localized extinction when new predators are introduced into the system (Paoletti et al. 2011). Additionally, predation or competition with introduced American bullfrogs (Rana catesbiana) likely impact this species (Fellers 2005). Native garter snakes (Thamnophis spp.) feed heavily on all life stages of this frog (Morey and Papenfuss 2000).

Parasites pose an additional threat to foothill yellow-legged frogs. The parasite, Ribeiroia has been shown to cause severe limb deformities in other frog species and has been found in the vicinity of foothill yellow-legged frogs. Another parasite, Anchor Worm (Lernaea cyprinacea), is non-native and typically infects fish but can infect larval foothill yellow-legged frogs which can cause deformities or mortality. During periods of warm water, declining discharge and high host density infestation rates are high and outbreaks can occur. However, limb malformation occurs in less than 25 percent of infected individuals and if animals are infected after leg development, little to no deformation is observed (Kupferberg et al. 2009). Perhaps the most significant parasite that impacts this species is Batrachochytrium dendrobatidis which causes amphibian chytridomycosis. This parasite has been found in this species and has had significant impacts to the similar mountain yellow-legged frog (Rana sierra and Rana muscosa) and other amphibian species worldwide (Fellers 2005).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Foothill yellow-legged frog was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species on the foothill yellow-legged frog.

Determination Statement Sequoia and Sierra NFs: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Inyo NF: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on this forests.

Literature Cited – Foothill Yellow-legged Frog Dever, J.A. 2007. Fine-scale genetic structure in the threatened foothill yellow-legged frog (Rana boylii). Journal of Herpetology 41(1): 168-173.

Fellers, G.M. 2005. Rana boylii Amphibiaweb account. In: Amphibiaweb http://www.amphibiaweb.org. Downloaded on 26 June 2012.

Jennings, M. R. and Hayes, M. P. 1994. Foothill yellow-legged frog Rana boylii Baird 1854. In: Amphibian and Reptile Species of Special Concern in California. California Department of Fish and Game, Sacramento, California.

Leidy, R.A. Gonsolin, E. and Leidy, G.A. 2009. Late-summer aggregation of the foothill yellow-legged frog (Rana boylii) in central California. The Southwestern Naturalist 54(3): 367-368.

Lind, A.J., Spinks, P.Q., Fellers, G.M. and Shaffer, H.B. 2011. Rangewide phylogeography and landscape genetics of the Western U.S. endemic frog Rana boylii (Ranidae): implications for the conservation of frogs and rivers. Conservation Genetics 12: 269-284.

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Lind, A.J., Welsh Jr., H.H. and Wilson, R.A. 1996. The effects of a dam on breeding habitat and egg survival of the foothill yellow-legged frog (Rana boylii) in northwestern California. Herpetological Review 27(2): 62-67.

Loomis, R.B. 1965. The yellow-legged frog, Rana boylii, from the Sierra San Pedro Martir, Baja California Norte, Mexico. Herpetologica 21(1): 78-80.

Kupferberg, S.J. 1996. Hydrologic and geomorphic factors affecting conservation of a river-breeding frog (Rana boylii). Ecological Applications 6(4): 1332-1344.

Kupferberg, S.J., Catenazzi, A., Lunde, K., Lind, A.J., and Palen, W.J. 2009. Parasitic copepod (Lernaea cyprinacea) outbreaks in foothill yellow-legged frogs (Rana boylii) linked to unusually warm summers and amphibian malformations in northern California. Copeia 2009(3): 529-537.

Kupferberg, S.J., Lind, A.J., Thill, V. and Yarnell, S.M. 2011. Water velocity tolerance in tadpoles of the foothill yellow-legged frog (Rana boylii): swimming performance, growth, and survival. Copeia 2011(1): 141-152.

Morey, S. and Papenfuss, T. 1990. Foothill yellow-legged frog Rana boylii. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Paoletti, D.J., Olson, D.H. and Blaustein, A.R. 2011. Responses of foothill yellow-legged frog (Rana boylii) larvae to an introduced predator. Copeia 2011(1): 161-168.

Rombough, C.J. and Hayes, M.P. 2005. Novel aspects of oviposition site preparation by foothill yellow- legged frogs (Rana boylii). Northwestern Naturalist 86: 157-160.

Santos-Barrera, G., Hammerson, G. and Fellers, G. 2004. Rana boylii. In: IUCN 2012. IUCN Red List of the Threatened Species. Version 2012.1. http://www.iucnredlist.org. Downloaded on 26 June 2012.

Sparling, D.W. and Fellers, G. 2007. Comparative toxicity of chlorpyrifos, diazinon, malathion and their oxon derivatives to larval Rana boylii. Environmental Pollution 147: 535-539.

Wheeler, C.A. and Welsh Jr., H.H. 2007. Mating strategy and breeding patterns of the foothill yellow- legged frog (Rana boylii). Herpetological Conservation and Biology 3(2): 128-142.

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California Golden Trout (Oncoryhnchus mykiss aguabonita) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species Account California golden trout are endemic to the South Fork of the Kern River and Golden Trout Creek, both located on an area referred to as the Kern Plateau in the Golden Trout Wilderness on the Inyo National Forest. They currently occupy all historic habitat within Golden Trout Creek, but only occupy about 25 percent of their historic habitat within the South Fork of the Kern River, plus a population in Mulkey Creek that was transplanted above a natural barrier. California golden trout populations suffered declines during the 19th and first half of the 20th century from overfishing and heavy grazing (Stevens 2004). Knapp (1990) estimated that golden trout streams typically support 8 to 58 fish per 100 m of stream, although a more recent estimates for Mulkey Creek, a tributary to the South Fork Kern River, found densities of 472 fish per 100 meters (Carmona-Catot and Weaver 2006), 350 fish per 100 meters in grazed sections, and 330 fish per 100 meters in an un-grazed area (Weaver and Mehalick, 2008). Fish were estimated in three different stream reaches in Big Whitney Meadow, rested from grazing for 8 years at the time, with estimates of 646, 351 and 216 fish per 100 meters in different stream sections (Weaver and Mehalick 2008).

Figure 23 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Key ecological conditions necessary for persistence of this species: California golden trout rely on water that is clear and mostly cold, although summer temperatures can fluctuate from 3 to 20 degrees Celsius (Knapp and Dudley 1990), and they can survive maximum temperatures up to 24 degrees Celsius. California golden trout generally prefer pool habitat and congregate near emergent sedges and undercut banks (Matthews 1996). They feed on both terrestrial and aquatic invertebrates, mostly adult and larval insects, taking whatever is most abundant.

Threats Three main factors have been identified as threats to the California golden trout: hybridization with rainbow trout, predation/competition by brown trout, and habitat modification due to historic and current grazing. Non-native trout species, including rainbow and brown trout, were stocked into the lower portions of the South Fork of the Kern River, and these fish moved upstream into the last remaining habitat of the golden trout. Several chemical treatments, along with the construction of two barriers (Templeton and Schaeffer barriers), were undertaken in the 1960’s to 1990’s to remove the non-native fish and protect the remaining golden trout. Golden trout in the South Fork of the Kern River show increasing hybridization levels towards their downstream range, whereas fish in the Golden Trout Creek watershed show overall very low hybridization levels (Cordes et al. 2006). Genetically ‘pure’ populations exist in only a few kilometers of streams. Brown trout are still present in the South Fork Kern River, between the upper Templeton Barrier and the lower Schaeffer Barrier, which compete for habitat niches and predate on golden trout.

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Figure 23. Map of California golden trout from NRIS Aquatic Survey Database, 2016

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Historic un-regulated grazing during the last half of the 19th century is thought to have contributed to significant habitat modification in many portions of the California golden trout’s range. Grazing that occurred in the mid-1800s and early 1900s consisted of many herds of cattle and horses, and bands of sheep that travelled through the area or stayed for the summer. The resulting effects of continuous use in the meadows may have contributed to the creation of large incisions through the meadows, dropping water tables after high run-off events that occurred in the early 20th century, changing the habitat of the golden trout. This left the deep and unconsolidated meadow material exposed and prone to de-stabilizing. Massive efforts of the Forest Service to repair these incised meadows were implemented in the 1930’s and continue to the present, which proved successful as deep gullies filled in and “new” meadows were created. Improved cattle management strategies implemented in the latter half of the 20th century built on the previous restoration as meadows regained their function and recovered desirable habitat for the golden trout. Current cattle management focuses on restoring the hydrologic and vegetative function of meadows in golden trout habitat, which has been demonstrated by current monitoring efforts. Although extensive restoration in golden trout habitat has been successful at recovering the streams through the meadows, there are still a few areas, such as Templeton and Monache meadows, where recovery is much slower, predominately due to the geomorphic processes, even though in these areas grazing has been severely restricted, or rested from grazing up to 20 years (Inyo NF files).

Because these fish are restricted to only a small portion of their endemic habitat, they are more susceptible to stochastic events, such as large-scale flood events, wildfire, drought, etc., that could alter significant portions of their habitat. Changes in climate with predicted earlier run-off times could impact the function of streams and connected watersheds, but it is difficult to speculate how these changes would affect fish habitat.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: California golden trout were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

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Alternative B, C and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Sierra NF: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on this forest.

Literature Cited – California Golden Trout Carmona-Catot, G., and J. Weaver. 2006. Golden trout report 2006. California Department of Fish and Game Heritage and Wild Trout Program, Sacramento, California.

Cordes et al. 2006

Knapp, R.A. and T.L. Dudley. 1990. Growth and longevity of golden trout, Oncorhynchus aquabonita, in their native streams. California Fish, Game 76:161-173.

Knapp 1990

Matthews, K.R. 1996. Habitat selection and movement patterns of California Golden trout in degraded and recovering stream sections in the Golden trout wilderness, California. North American Journal of Fisheries Management, 16:579-590.

Stephens, S. J., C. McGuire, C. and L. Sims. 2004. Conservation Assessment and Strategy for the California Golden Trout (Oncorhynchus mykiss aguabonita), Tulare County, California. Calif. Dept. Fish and Game.

Weaver and Mehalick, 2008

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Hardhead Minnow (Mylopharodon conocephalus) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account This species occurs in scattered tributaries of the San Joaquin River but not in the valley reaches of the river (Moyle 2002). Elevational range is 10 to 1,450 meters (http://calfish.ucdavis.edu). Hardhead minnows are found in the Kern River Upper Tehachapi-Grapevine Watershed; South Fork Kern Watershed; and rivers along the east side of the San Joaquin Valley.

California’s populations of hardhead minnow, a fish species of special concern, have experienced population declines, possibly due to habitat perturbations, including dam construction with consequent temperature changes and the introduction of non-native species to California’s mid- to low-elevation streams especially in the southern part of their range (Moyle 2002).

Habitat Status Hardhead are typically found in small to large streams in a low to mid-elevation environment. Within a stream hardhead tend to prefer warmer temperatures than salmonids and they are often found associated with pikeminnows and suckers (Moyle 2002). Their preferred stream temperature might easily exceed 20 degrees Celsius, though these fish do not favor low dissolved oxygen levels (Kaufman et al. 2013). Therefore the hardhead minnow is usually found in clear deep streams with a slow but present flow. Most hardhead reach sexual maturity at 3 years and spawn in the spring around April-May, though spawning may take place as late as August. In small streams hardhead tend to spawn near their resident pools, while fish in larger rivers or lakes often move up to 30 to 75 kilometers to find suitable spawning grounds. Though spawning may occur in pools, runs, or riffles, the bedding area will typically be characterized by gravel and rocky substrate. Upon hatching, young larval hardhead remain under vegetative cover along stream or lake margins. As the juveniles grow they may move to deeper water or be swept downstream to larger rivers below. Adult hardhead may live up to 9 or 10 years.

Threats Widely distributed in California but declining and vulnerable to local extirpation from habitat alteration (stream flow quantity and quality changes) and introduced fishes (NatureServe). Hydropower peaking at odd times can negatively influence hardhead. Hardhead have evolved to cope with seasonal shifts in water temperatures and flows. Altered flow conditions could force hardhead into smaller habitat areas (river sections), especially when river temperatures increase excessively (Kaufman et al. 2013,). It is clear from thermal preference experiments that hardhead prefer cool water (Kaufman et al. 2013).

Dams and diversions have eliminated habitat and left many populations isolated and vulnerable to local extinction due to unsuitable stream temperatures and flows (Moyle 2002). Centrarchid fishes (bass, sunfish) threaten populations in foothill streams. Reservoir populations should not be regarded as "safe" populations since these also are vulnerable to large declines caused by increased populations of introduced smallmouth bass and other centrarchid basses (Moyle 2002).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of aquatic areas.

Alternative B, C, and D: Hardhead were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

Literature Cited – Hardhead Kaufman, R.C., Coalter, R., Nordman, N.L., Cocherell, D., Cech Jr, J.J., Thompson, L.C. and Fangue, N.A., 2013. Effects of temperature on hardhead minnow (Mylopharodon conocephalus) blood- oxygen equilibria. Environmental biology of fishes 96(12):1389-1397.

Moyle, P.B., 2002. Inland fishes of California. Univ. of California Press.

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Kern brook lamprey (Lampetra hubbsi) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The Kern brook lamprey was originally described in the genus Entosphenus. The taxonomic status of the genus Lampetra is under debate (see Vladykov and Kott 1976); Robins et al. (1991) retained Lampetra as the genus and regarded Entosphenus as a subgenus. Apparently, L. hubbsi was derived from parasitic L. tridentata (Lee et al. 1980). Other nonparasitic species in this genus occur in south central California and in the Pit and Klamath River drainages in northern California. L. hubbsi apparently is distinctive from all others (Starnes 1995). See Moyle et al. (1989) for comparative morphological data on California Lampetra.

The Kern brook lamprey (Vladykov and Follett 1976) is endemic to the east side of the San Joaquin Valley, California; Friant-Kern Canal, east of Delano, Kern County, California, which apparently provides ammocoete habitat but not spawning habitat; lower reaches of the Merced, Kaweah, Kings, and San Joaquin rivers (Moyle et al. 1989, Moyle 2002). Lampreys with low numbers of trunk myomeres (i.e. mussel subunits) reported from the upper San Joaquin River between Millerton Reservoir and Kerckhoff Dam, as well as those collected in the Kings River above Pine Flat Dam (Fresno County), also may be L. hubbsi (Moyle et al. 1989, Moyle 2002). Apparently the species is thinly distributed throughout the San Joaquin drainage, with populations isolated from one another, at elevations of 30 to 327 meters (Moyle et al. 1989, Moyle 2002). The California Fish Website (http://calfish.ucdavis.edu/species/?uid=39&ds=241) lists 8 watersheds for this species: Middle San Joaquin-Lower Chowchilla Watershed, Middle San Joaquin-Lower Merced-Lower Stanislaus Watershed, Mill Watershed, Tulare-Buena Vista Lakes Watershed, Upper Dry Watershed, Upper Kaweah Watershed, Upper King Watershed, and Upper Merced Watershed.

Moyle (2002) rated Kern brook lamprey as "special concern; the species is in decline, so species management is needed to keep it from becoming threatened or endangered." Moyle (2002) specified that relatively few unequivocal collections of this species have been made since it was first discovered in 1976. This is because most collections are ammocoetes that cannot be reliably distinguished from those of western brook lamprey, a more broadly distributed species. Probable populations are thinly scattered throughout the San Joaquin drainage and isolated from one another (Brown and Moyle 1993). This fragmented distribution makes local extirpations likely, without the potential for recolonization, followed by eventual extinction. The probability of local extirpation is increased because all known populations but one are below dams, where regulated discharges result in fluctuations or sudden drops in flows that may strand or desiccate ammocoetes.

Moyle et al. (2011) rated Kern brook lamprey as 2.0 (vulnerable) which means “sufficiently threatened to be on a trajectory toward extinction if present trends continue.” Jelks et al. (2008) list the species as threatened and declining.

Although existing data are sparse, Nawa (2003) noted that each of the four species of lamprey from the west coast of North America (Pacific lamprey, river lamprey, western brook lamprey, and Kern brook lamprey) is likely to become extinct or endangered with extinction in the foreseeable future throughout all or parts of their range in the coterminous United States.

Distribution Nawa (2003) compiled the following distribution for Kern brook lamprey:

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 Friant-Kern Canal and Merced River, CA - Historical collections of the Kern brook lamprey are known only from the Friant-Kern Canal in Merced County and the Merced River (Froese and Pauly 2002).  Kings River, CA - Seven Kern brook lamprey were documented in November 1995 in Kings River, below the North Fork New River, near Trimmer, California (USGS 2002). Lampreys collected in the Kings River above Pine Flat Dam (Fresno County) also may be L. hubbsi (Moyle et al. 1989, Moyle 2002).  Kaweah River, CA - Lower reaches of Kaweah River (Moyle et al. 1989, Moyle 2002).  San Joaquin River, CA - Lampreys reported from the upper San Joaquin River between Millerton Reservoir and Kerckhoff Dam may be L. hubbsi (Moyle et al. 1989, Moyle 2002).

Life History Moyle (2002) indicated that the principle habitats of the Kern brook lamprey are silty backwaters of rivers emerging from the Sierra foothills (mean elevation of 135 meters with a range from 30 to 327 meters). Ammocoetes are usually found in shallow pools and along edges of runs where water velocity is low. Ammocoetes favor substrates that are a mixture of sand and mud ranging in depth from 30 to 110 centimeters, where summer temperatures rarely exceed 25 degrees Celsius (Brown and Moyle 1993). This habitat also characterizes the lightless siphons of the Friant-Kern Canal, where ammocoetes are abundant at times. Presumably, siphon populations do not contribute to the survival of the species, because adults derived from them would wind up in the aqueduct itself. Adults in natural environments seek riffles with gravel for spawning and rubble for cover. Based on the times at which adults are collected, Kern brook lampreys undergo metamorphosis in fall and spawn in spring. Other aspects of its life history are not known, but are presumed to be similar to those of the western brook lamprey.

According to Kostow (2002), lamprey eggs are sticky and dense, and are deposited in redds by spawning adults. Upon completion of a redd, the eggs are buried beneath sand and gravel. The length of egg incubation lasts between ten and twenty days, with longer incubation times when water is cooler. Upon hatching early larva spend another week to a month in the redd. They eventually emerge from the natal redd at night and move downstream to areas with fine silt deposits and a mild current where they burrow into the silt. At this age they are about 10 mm long. Optimal spawning grounds appear to be riffle/gravel areas in close proximity to pools or other silt deposits so that the tiny larvae find substrates suitable for burrowing during their initial movement from the redd. The burrow is U-shaped, with the lamprey’s mouth at the surface of one end from where it filter feeds. For the next three to seven years (depending on species and regional variation in temperature and productivity), lamprey ammocoetes will remain in burrows filter feeding on suspended algae, mostly diatoms. They will move gradually downstream, primarily at night, seeking courser sand and silt substrates and deeper water as they grow. Older ammocoetes tend to congregate in lower basins and flood plains.

Growth rate may vary seasonally, influenced by water temperature and food supply. The most rapid increase in length occurs in the first years (1 to 3 years) during which ammocoetes of most species reach about 10 centimeters in length. Lipid accumulation begins at about that size, and growth rate in length declines, in preparation for the non-feeding period of metamorphism. The age of ammocoetes is very difficult to determine because the species lacks bony structures, which allow aging in bony fishes.

Courtship and spawning behaviors of northwest lampreys have been described for only a few species and in some cases only in captivity. Pletcher (1963) described these behaviors for western brook lamprey, L. richardsoni, in both captivity and the wild. Courtship occurs on spawning gravels and involves redd building and mutual displays that are tactile and probably chemical. Solitary males, and perhaps some females, may begin by preparing multiple rudimentary nests. Either gender may initiate courtship by way

179 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests of a “courtship glide”, where one lamprey slithers along the body of a prospective mate. A receptive mate will accompany the initiator to the rudimentary nest. Initial nests may be communal, occupied by as many as a dozen individuals of both genders. Pletcher (1963) believed that receptive females emitted a chemical stimulus that attracted other lamprey.

Communal courtship generally breaks into pairs or smaller groups and disperses to separate nests before actual spawning begins. The female lays in the rudimentary nest and undulates while the male performs most of the nest building. Lamprey will carry smaller rocks to the edge of the nest in their oral disks. Larger rocks may be pushed and finer substrates may be expelled by rapid swimming motions. When the lampreys are ready to spawn, the male grasps the female by the back of her head with his oral disk, then twists his tail around her. They simultaneously vibrate while clasped together, depositing, fertilizing and burying eggs. Spawning is mostly done by pairs, but may include additional lamprey. Both polygamous and polyandrous group matings have been observed. A female will deposit about 100 to 500 eggs in each spawning bout. Between bouts, the female rests while the male briefly departs. He resumes nest building upon return, enlarging the nest upstream so that previous egg deposits are minimally disturbed. Another spawning bout will be followed by another rest. A female probably deposits all her eggs in about 12 hours. Most authors believe that all lamprey die soon after spawning. The extreme physiological changes, in particular the atrophy of the gut and the filling of the body cavity with gonadal materials, seems to make life after spawning unlikely. Females have been described as living only a few hours to a week after spawning; males perhaps for a few weeks.

Habitat Status Generally, lamprey redds may be identified as round depressions in gravel or cobble substrates. Rocks moved by lampreys during redd construction are typically piled on all sides of the depression rim and are lighter colored due to exposure of non-algal-covered surfaces. A plume of finer substrate carried by the current may be apparent downstream of the depression as a result of spawning and egg burial. However, in places with low velocities, these finer substrates may not be apparent. In many instances, a few large rocks are left in the bottom of the depression (Gunckel et al. 2009). Because Kern brook lamprey redds are small (15 to 20 centimeters in diameter), large rubble or cobble substrates serve as adequate cover under which spawning adults can hide.

The Kern brook lamprey was first discovered in the Friant-Kern Canal, but it has since been found in the lower reaches of the Merced, Kaweah, Kings, and San Joaquin rivers, as well as in the Kings River above Pine Flat Reservoir and the San Joaquin River above Millerton Reservoir (Brown and Moyle 1987, 1992, 1993). In 1988, ammocoetes and adult lampreys were found in several siphons of the Friant-Kern Canal, when they were poisoned during an effort to rid the canals of white bass (Morone chrysops). The "low- count" lampreys (i.e., low numbers of trunk myomeres) reported from the upper San Joaquin River between Millerton Reservoir and Kerckhoff Dam by Wang (1986) are also most likely L. hubbsi, as are similar appearing ammocoetes from the Kings River above Pine Flat Reservoir.

Populations of this species are thinly scattered throughout the San Joaquin drainage and isolated from one another. Such a fragmented distribution makes local extirpations likely, with no potential for recolonization, followed by eventual extinction of the species. The probability of local extirpation is increased by the fact that all known populations, with one exception, are located below dams, where stream flows are regulated without regard to the needs of the lampreys. Fluctuations or sudden drops in flow may isolate or dry up ammocoetes.

Recorded observations of Kern brook lamprey in Region 5 include the Sierra NF and Sequoia NF (Kings River upstream from Pine Flat Reservoir).

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Habitat associations - Principal habitats of Kern brook lamprey are silty backwaters of large rivers in the foothill regions (mean elevation is 135 meters with a range from 30 to 327 meters). In summer, ammocoetes are usually found in shallow pools along edges of run areas with slight flow (L.R. Brown, pers. comm.) in substrates reaching depths of 30 to 110 centimeters where water temperatures rarely exceed 25 degrees Celcius. Common substrates occupied are sand, gravel, and rubble (average compositions being 40 percent, 22 percent, 23 percent, respectively). Ammocoetes seem to favor sand/mud substrate where they remain buried with the head protruding above the substrate and feed by filtering diatoms and other microorganisms from the water. This type of habitat is apparently present in the siphons of the Friant-Kern Canal. Adults require coarser gravel-rubble substrate for spawning. Temperature requirements for Kern brook lamprey are not known but the fact they are present almost entirely in reaches where summer temperatures rarely exceed 24 degrees Celcius is suggestive of a cool- water requirement.

Threats Lampreys have experienced declines in abundance throughout the Northern Hemisphere due to human disturbances (Renaud 1997). Lamprey on the Oregon coast, Columbia Basin, and elsewhere have declined along with salmonids occupying the same habitat, which suggests that the same habitat disturbances that have caused the federal listing of anadromous salmonids also affect lamprey (NMFS 1996 and 1998, Close et al. 2002).

Dams and other artificial barriers - Altered hydrographs for flood control or power generation may harm lamprey. For example, lamprey are passive swimmers and increased migration time due to manipulated discharge in the Columbia system may affect them adversely as it does salmonids (for example, through increased exposure to predation) (Kostow 2002, BioAnalysts 2000). Rapid artificial drawdown of streamflow can cause ammocoetes to be stranded in exposed bars and mudflats. The fall dewatering of irrigation screens at Savage Rapids Dam on the Rogue River causes ammocoetes to be stranded in their burrows. Water level fluctuations in the mainstem North Umpqua River caused by hydro-power generation temporarily dewater stream margin habitats used by ammocoetes resulting in observed mortalities (Pacificorp 1998). Artificial features such as tide gates, hatchery weirs, and stream diversion structures may also be barriers to upstream migration (Kostow 2002).

Road Culverts - Similar to dams, culverts that pass adult salmonids are often barriers to lamprey. A systematic survey of lamprey in the Alsea Basin, Oregon found lampreys were often absent above road culverts (Kostow 2002).

Water Diversions - Stream diversions can kill juvenile and adult lamprey by stranding due to artificial lowering of the water level, or because the diversions are unscreened or the lamprey can get under or through the screens (Kostow 2002; BioAnalysts 2000). Low flows or no flow can kill ammocoetes rearing below water diversions and exacerbate high temperature and sediment effects (Close 2002; BioAnalysts 2000). Production from adult lamprey may be lost when they spawn in irrigation ditches. For example, an irrigator in the Entiat Valley, Washington found ammocoetes in the irrigation system when the intake was in the Entiat River (BioAnalysts 2000). A constant head orifice at Stony Creek and the Tehama-Colusa Canal (Sacramento River basin) entrained and killed 165 lamprey larvae from January 28 to May 4, 1994, but only 6 were entrained from September 9 to September 22, 1994 (Brown 1994). Lamprey ammocoetes were the third most common fish entrained by experimental pumps on the Sacramento River, but survival upon recovery from the screens ranged from 95 to 100 percent (Borthwick et al. 1999).

Chemical and Organic Waste Poisoning - Bridge crossings, roads, and irrigation ditches make eradication from accidental spills or intentional chemical treatment a high-risk threat. Lampreys are particularly

181 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests vulnerable to chemical spills because populations in a basin may concentrate in one stream (Kostow 2002, Nawa 2003). Since lamprey ammocoetes take up to six years before metamorphosing, six years of production are lost during a chemical poisoning. If all or a substantial amount of the stream’s ammocoetes are killed, adult lamprey may not be drawn to it to spawn, resulting in local extinction. In 1998, Kern brook lamprey ammocoetes and adults were collected from the siphons of the Friant-Kern canal when they were poisoned as part of an effort to eradicate white bass from the system (Moyle 2002). A 1999 spill of a herbicide in the lower portions of Fifteenmile Creek, Oregon killed thousands of lampreys. Chemical spills in the Willamette basin have also killed lamprey (Pringle Creek, Salem, Oregon). A 2001 spill of liquid cow manure into Gee Creek near Ridgefield, Washington killed lamprey and other fish (Washington Dept. Ecology 2001). A probable chlorine spill in upper Alameda Creek in the Sunol Valley in April 2002 killed at least 24 to 36 Pacific lampreys (Nawa 2003). From the late 1940s through the late 1980s the Oregon Fish Commission killed non-game fishes across the state with rotenone (Close et al. 1995). The use of rotenone in John Day River by Oregon Department of Fish and Wildlife killed thousands of lampreys in 1969 and 1982. About 90 and 85 miles of the Umatilla River were poisoned in 1967 and 1974 to eradicate non-game fish. The 1967 treatment killed one million fish, which was about 95 percent of the fish in the treated area. September poisoning in the Umatilla would have decimated several age classes of larvae, young adults, and adults returning to spawn (efforts are now underway to reestablish lamprey in the Umatilla River). The Oregon Department of Fish and Wildlife purposely exterminated a unique lake dwelling sub-population of the Miller Lake lamprey with chemical treatment of Miller Lake (Bond and Kan 1973). Miller Lake lampreys present in Miller Creek have failed to colonize the lake. The Environmental Protection Agency recently detected high levels of polychlorinated biphenyls (PCBs) in lampreys collected from the Columbia River (Kostow 2002). Lamprey ammocoetes tend to concentrate in the lower portions of streams and rivers where stream gradients are low. These same areas are often heavily polluted by industry, agriculture, and urbanization (for example, Bear Creek in the Rogue Basin, Willamette River, and Umatilla River). Because ammocoetes spend 4 to 7 years filter feeding in the benthos they would be vulnerable to chemical toxicants that bioaccumulate (BioAnalysts 2000, Close et al. 2002).

Dredging - Kostow (2002) reports that most lamprey die after passing through dredges. Suction dredging for gold would also likely kill developing eggs and ammocoetes (Nawa 2003).

Channelization and destruction of riparian vegetation - Lamprey species depend on muddy bottoms, backwater areas, and low gradient areas during their larval life stage. Lampreys are greatly affected by loss of wetlands, side channels, back eddies, and beaver ponds (PSMFC 1997). Channelization, floodplain filling, and destruction of riparian vegetation is widespread in low-gradient stream areas favored by lamprey for spawning and rearing. River channelization negatively impacts larval lamprey habitat by increasing stream velocity, thereby reducing depositional areas favored by larval lamprey (Close et al. 2002). High stream temperatures resulting from the destruction of riparian vegetation are a likely limiting factor because lampreys prefer temperatures below 20 degrees Celcius (BioAnalysts 2000). Although most streams in Oregon exceed the Department of Environmental Quality Standard of 17 degrees Celcius, few streams reach 82 degrees Fahrenheit, the temperature at which lamprey ammocoetes begin to die (van de Wetering and Ewing 1999). Other aspects of elevated stream temperature may adversely affect lamprey survival, such as increased metabolic rates during metamorphosis to young adults and decreased stream microbial activity during the summer (van de Wetering and Ewing 1999).

Predation by non-native fish - Hubbs (1967) reported "no evidence of any lamprey occurrence at any time in coastal streams south of the Los Angeles Plain. Here, as over much of the West depletion of the water and introduction of more aggressive eastern [USA] fishes [such as bass of the family Centrachidae] have rapidly brought the native fish fauna to or beyond the brink [of extinction]." Kern brook lampreys are

182 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests vulnerable to predation by alien fishes, especially where conditions are favorable for predator fish from eastern states.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of aquatic areas.

Alternative B, C, and D: Kern brook lamprey was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

Literature Cited – Kern Brook Lamprey BioAnalysts, Inc. 2000. A Status of Pacific Lamprey in the Mid-Columbia Region. Rocky Reach Hydroelectric Project FERC Project No. 2145. Prepared for Public Utility District No. 1 of Chelan County, Wenatchee, Washington.

Bond, C. E. and T. T. Kan. 1973. Lampetra (Entosphenus) minima n. sp., a dwarfed parasitic lamprey from Oregon. Copeia 1973 (3): 568-574.

Borthwick, S.M., R.R. Corwin, C.R. Liston. 1999. Investigations of fish entrainment by archimedes and internal helical pumps at the Red Bluff research pumping plant, Sacramento River California; February 1997-June 1998. U.S. Bureau of Reclamation, Red Bluff, California.

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Brown, L.R. and P.B. Moyle. 1987. Survey of fishes of mid-elevation streams of the San Joaquin Valley. Unpublished Report, California Department of Fish and Game.

Brown, L.R. and P.B. Moyle. 1992. Native fishes of the San Joaquin Valley drainage: status of a remnant fauna and its habitats. Pages 89-98, In: D.F. Williams, T.A. Rado and S. Byrde, eds., Proceedings of the Conference on Endangered and Sensitive Species of the San Joaquin Valley, California. California Energy Commission, Sacramento.

Brown, L.R. and P.B. Moyle. 1993. Distribution, ecology, and status of the fishes of the San Joaquin River drainage, California. Calif. Dept. Fish & Game 79:96-113.

Brown, M.R. 1994. Fishery impacts from reverse operations of the constant head orifice at Stony Creek and the Tehama-Colusa Canal, California. USFWS AFFI-FRO-94-12.

Close, D., M. Fitzpatrick, H. Li, B. Parker, D. Hatch, and G. James. 1995. Status report of the Pacific lamprey (Lampetra tridentata) in the Columbia River basin. Project No. 94-026, Contract No. 95BI9067. Report to the U.S. Department of Energy, Bonneville Power Administration, Portland, Oregon.

Close, D.A., M.S. Fitzpatrick, and H.W. Li. 2002. The ecological and cultural importance of a species at risk of extinction, Pacific lamprey. Fisheries 27(7):19-25.

Froese, R. and D. Pauly. Editors. 2002. FishBase. World wide web electronic publication, www.fishbase.org.

Gunckel et al. 2009

Hubbs, C.L. 1967. Occurrence of the Pacific lamprey, Entosphenus tridentatus, off Baja California and in streams of southern California; with remarks on its nomenclature. Transactions San Diego Society Natural History 14(21):303-311.

Jelks, H.L., S.J. Walsh, N.M. Burkhead, S. Contreras-Balderas, E. DÌaz-Pardo, D.A. Hendrickson, J. Lyons, N.E. Mandrak, F. McCormick, and J.S. Nelson. 2008. Conservation status of imperiled North American freshwater and diadromous fishes. Fisheries 33:372-386.

Kostow, K. 2002. Oregon lampreys: natural history status and analysis of management issues. Oregon Department of Fish and Wildlife, Portland, Oregon.

Lee, D. S., C. R. Gilbert, C. H. Hocutt, R. E. Jenkins, D. E. McAllister, and J. R. Stauffer, Jr. 1980. Atlas of North American freshwater fishes. North Carolina State Museum of Natural History, Raleigh, North Carolina. i-x + 854 pp.

Moyle, P. B., J. E. Williams, and E. D. Wikramanayake. 1989. Fish species of special concern of California. Final report submitted to California Dept. of Fish and Game, Inland Fisheries Division, Rancho Cordova. 222 pp.

Moyle, P.B. 2002. Inland fishes of California. Regents of the University of California, Berkeley and Los Angeles, CA.

Moyle, P.B., J. V. E. Katz and R. M. Quiñones. 2011. Rapid decline of California's native inland fishes: a status assessment. Biological Conservation 144: 2414-2423.

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Nawa, R. 2003. A Petition for Rules to List: Pacific Lamprey (Lampetra tridentata); River Lamprey (Lampetra ayresi); Western Brook Lamprey (Lampetra richardsoni); and Kern Brook Lamprey (Lampetra hubbsi) as Threatened or Endangered under the Endangered Species Act. Letter to the U.S. Fish and Wildlife Service, Department of the Interior.

NMFS (National Marine Fisheries Service) 1996. Factors for Decline [of west coast steelhead]. Portland, Oregon. http://www.nwr.noaa.gov/1salmon/salmesa/pubs/sthffd/pdf.

NMFS (National Marine Fisheries Service) 1998. Factors contributing to the decline of Chinook salmon: An addendum to the 1996 west coast steelhead factors for decline report. Portland, Oregon. http://www.nwr.noaa.gov/1salmon/salmesa/pubs/chinffd/pdf.

PacifiCorp. 1998. North Umpqua Cooperative Watershed Analysis. North Umpqua Hydroelectric Project, FERC Project No. 1927, Douglas County, Oregon. Portland, Oregon.

Pletcher, F.T. 1963. The Life History and Distribution of Lampreys in the Salmon and Certain Other Rivers in British Columbia, Canada. Master of Science Thesis. University of British Columbia, Vancouver B.C. 195 p.

PSFMC (Pacific States Marine Fisheries Commission). 1997. Anadromous fish of the Pacific Northwest. Habitat Education Program, Pacific States Marine Fisheries Commission. Available at: http://www.psmfc.org/habitat/edu_lamprey_fact.html

Renaud, C.B. 1997. Conservation status of northern hemisphere lampreys (Petromyzontidae). Journal of Applied Ichthyology 13:143-148.

Robins et al. (1991)

Starnes, W. C. 1995. Taxonomic validation for fish species on the U.S. Fish and Wildlife Service Category 2 species list. 28 pp.

USGS (U. S. Geological Survey) 2002. National water quality assessment program, San Joaquin-Tulare basins study unit. World wide web publication, www.water.wr.usgs.gov/sanj_nawqa van de Wetering, S.J. and R.E. Ewing. 1999. Lethal temperatures for larval Pacific lamprey, Lampetra tridentata. Confederated tribes of the Siletz Indians. Siletz, Oregon.

Vladykov, V. D., and E. Kott. 1976. A new nonparasitic species of lamprey of the genus Entosphenus Gill, 1862, (Petromyzonidae) from south central California. Bulletin of the Southern California Academy of Science 75:60-67.

Wang, J.C.S. 1986. Fishes of the Sacramento-San Joaquin estuary and adjacent waters, California: A guide to the early life histories. Interagency Ecological Study Program for the Sacramento-San Joaquin Estuary, Tech. Report 9.

Washington Dept. of Ecology. 2001. Ridgefield dairy owner fined for Gee Creek manure spill. News Release dated October 12, 2001.

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Kern River Rainbow Trout (Oncorhynchus mykiss gilberti) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species account The taxonomic status of Kern River rainbow trout is controversial, because of its complex evolutionary history and exposure to introduced varieties of rainbow trout, with which it may breed, degrading genetic purity. D. S. Jordan's 1894 designation of this fish as a distinctive subspecies of rainbow trout was accepted until Schreck and Behnke (1971) described it as a population of golden trout. Their decision was based mostly on comparisons of lateral scale counts and on aerial surveys that led them to believe that there were no effective barriers on the Kern River which might have served to isolate trout in the Kern River from those in the Little Kern River. However, in a subsequent analysis, Gold and Gall (1975) determined that golden trout populations were effectively isolated genetically and physically. Meristic (Gold and Gall 1975) and genetic (Berg 1987) characteristics of Kern River rainbow trout were regarded as sufficiently distinctive to warrant its subspecific status (Berg 1987). Bagley and Gall (1998), using mitochondrial and nuclear DNA, found that the Kern River rainbow was distinctive, but probably originated as the result of an early (natural) invasion of coastal rainbow trout that hybridized with Little Kern golden trout, creating a new genome. This has been more or less confirmed by analysis of genetic variation by amplified fragment length polymorphism markers for populations of rainbow trout statewide (Stephens 2007). This analysis indicates that Kern River rainbow trout represent a distinct lineage that is intermediate between coastal rainbow trout and Little Kern golden trout, although there is also some evidence of recent hybridization with coastal rainbows, presumably of hatchery origin. It is also possible that the mixed nature of the genome was the result of planting of the other two golden trout subspecies into Kern River rainbow trout waters (Bagley and Gall 1998).

According to NatureServe (2012): Behnke (1992) grouped the Kern and Little Kern golden trout as one subspecies (gilberti) of O. mykiss, and stated that they could be recognized as separate subspecies (gilberti and whitei, respectively) provided they are kept together in the same species (O. mykiss). Behnke indicated that whitei may be indistinguishable from gilberti. Behnke (2002) treated these forms as three subspecies: Golden Trout Creek golden trout or California golden trout (O. mykiss aguabonita), Kern River rainbow trout (O. mykiss gilberti), and Little Kern River golden trout (O. mykiss whitei). This nomenclature was also used by Moyle (2002).

Kern River rainbow trout (Jordan 1894) is endemic to the Kern River system, California. Once widely distributed in the Kern River system (probably downstream as far as Keyesville in the mainstem Kern River and downstream to Onyx on the South Fork of the Kern River), it now occurs in the Kern River from Durrwood Creek upstream to Junction Meadow. Populations established through transplantation occur in Rattlesnake Creek, Osa Creek, upper Ninemile Creek and possibly upper Peppermint Creek (Moyle 2002). Additionally, there are introduced populations of Kern River rainbows in the Kaweah-Kern River and Chagoopa Creek, which appear to have maintained their genetic integrity (CalTrout 2008).

Bagley and Gall (1998), using a variety of genetic techniques, determined that several populations, mostly located in the middle section of the Kern River drainage appeared to be unhybridized Kern River rainbow trout: Rattlesnake Cr. (in Sequoia National Park), Kern River at Kern Flat, Kern River above Rattlesnake Creek, Boreal Creek, Chagoopa Creek, Kern River at Upper Funston Meadow, Kern River above Redspur Creek, and Kern River at Junction Meadow. These populations in the middle of the historic range lacked apparent influence from California golden trout (either anthropogenic or natural) that was seen in the upper sections of the Kern and also lacked apparent rainbow trout hybridization seen in the lower sections. While Behnke (2002) doubts that pure Kern River rainbow trout still exist in their native range,

186 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests recent genetic analyses suggest that at least some unhybridized populations exist as indicated above. Much of their remaining habitat is in Sequoia National Forest (at least 29 kilometers) and Sequoia National Park (over 40 kilometers). In addition, there are distinctive introduced populations in the Kaweah-Kern River and Chagoopa Creek, which have maintained their genetic identity (M. Stephens 2007).

The Kern River rainbow trout has a high probability of extinction in the next 50 to 100 years if present trends continue. It is listed as a species of special concern by both the U.S. Fish and Wildlife Service and the California Department of Fish and Game. A multi-agency management plan for the upper Kern River basin lists as its goals to “restore, protect, and enhance the native Kern River rainbow trout populations so that threatened or endangered listing does not become necessary.” The Edison Trust Fund, established as mitigation for a hydropower generating station, provides at least $200,000 each year to implement the management plan and improve fish populations in the upper Kern Basin. Funding has been provided for developing a conservation hatchery for Kern River rainbow trout, for increasing patrols of wardens in areas where the trout are fished, and for funding genetic studies (CalTrout 2008).

Moyle (2002) rated Kern River rainbow trout as " threatened or endangered; the species is likely to become extinct or extirpated in the near future (less than 25 years), unless steps are taken to save it." Moyle et al. (2011) rated Kern River rainbow trout as "endangered," meaning “in danger of extinction in the near future if present trends continue." Jelks et al. (2008) list the species as threatened and declining.

Life History According to Moyle (2002), no life history studies have been done on this subspecies, but its life history is no doubt similar to other rainbow trout populations in large rivers. Most studies on golden trout subspecies have been on populations of California golden trout. The biology of the other two subspecies (including Kern River rainbow trout) is believed to be similar.

Because the valleys of the Kern Plateau were not subject to Pleistocene glaciation and are highly erosive due to the presence of decomposing granite, they are broad, flat, and filled with alluvium, resulting in wide meadows through which the streams meander. Native habitat for golden trout (including Kern River rainbow trout) is primarily found at high elevations (greater than 2,300 meters). The principle stream habitats of golden trout, are wide, shallow and exposed, with limited riparian vegetation to provide cover. The bottoms consist largely of sand, gravel and some cobble. The water is clear and usually cold, although summer temperatures can fluctuate from 3 to 22 degrees Celcius on a daily basis (Knapp and Dudley 1990). Preferred habitats of the trout are pools and areas associated with undercut banks, aquatic vegetation, and clumps of sedges (Matthews 1996a and 1996b).

Individual golden trout tend to remain in small stretches of stream measuring 16 to 18 meters; they rarely move more than 5 meters in a day. Long distance movements seem to take place mainly at night (Matthews 1996a and 1996b). Golden trout feed both day and night on a wide variety of items, especially aquatic insects. Lack of predators may account for their diurnal feeding habits. Despite increased vulnerability to birds and mammals, males develop especially bright colors during the breeding season. Golden trout may live up to 9 years, reaching 10 to 11 centimeters in standard length (body length minus the tail fin) by the end of their third summer. Their growth rate slows to 1 to 2 centimeters per year thereafter and they may eventually reach 19 to 20 centimeters in standard length. Historically, fish found in the mainstem Kern River grew to larger sizes, as much as 71 centimeters in total length (body length including longest lobe of the tail fin) and 3.6 kilograms in weight (Behnke 2002), although fish over 25 centimeters in total length are rare today (Stephens et al. 1995). Golden trout reach sexual maturity in 3 to 4 years and spawn in late spring or early summer when water temperatures range from 10 to 15 degrees Celcius. Spawning peaks in the afternoon when water temperatures reach their daily maximum. Females

187 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests dig wide shallow redds in riffles with small gravel particles (4 to 12 millimeters gravel size), shallow depths (5 to 20 centimeters), and water velocities of 30 to 70 centimeters per second (Knapp and Vredenburg 1996, Stefferud 1993). Females lay 300 to 2,300 eggs, depending on body size. The eggs hatch in approximately 20 days at 14 degrees Celcius. Fry remain in the substrate for 2 to 3 weeks before emerging into the main water column.

Trends in Abundance Kern River rainbow trout were once very abundant and widespread in the upper Kern Basin, and were subject to intensive fisheries as a consequence. Since the 19th century overexploitation combined with habitat degradation from grazing and, most importantly, hybridization with other trout, have reduced populations to a small fraction of historic numbers, probably less than 5 percent. In 1992, a study of Kern River rainbow trout abundance in the Kern River in Sequoia National Park, indicated there were about 600 to 1,400 trout per mile of all sizes (Stephens et al. 1995). There is no data on current abundance but if it assumed they currently persist in 20 kilometers of small streams, with 400 to 900 trout per kilometer, the total numbers would be 8,000 to 18,000 fish total. If 10 percent of these fish were capable of reproduction, then the effective population size would probably be less than 1,000 fish. These estimates are highly questionable, given natural variation in numbers, smallness of sample sizes on which they are based, and uncertainties about the actual distribution of Kern River rainbow trout, but they do suggest that absolute numbers in the wild are low and vulnerable to reduction by natural and human-caused events. Thus the status of Kern River rainbow trout could deteriorate rapidly considering the limited number of local populations.

Habitat Status The construction of Isabella Dam eliminated much of the Kern River rainbow’s historic habitat. This barrier resulted in massive introductions of hatchery trout into the river and heavy fishing pressure which led to the elimination of most of the native population (CalTrout 2008).

The Kern River rainbow trout is a subspecies endemic to the Kern River and tributaries, Tulare County. It was once widely distributed in the system; in the mainstem it probably existed downstream well below where Isabella Dam is today and upstream in the South Fork as far as Onyx (Stephens et al. 1995). It has been extirpated from the Kern River at least from the Johnsondale Bridge (about 16 kilometers above Isabella Reservoir) on downstream. Today, remnant populations live in the Kern River above Durrwood Creek, in Upper Ninemile, Rattlesnake and Osa Creeks, and possibly upper Peppermint Creek, and others (Stephens et al. 1995). Bagley and Gall (1998), using a variety of genetic techniques, determined that several populations, mostly located in the middle section of the Kern River drainage, appeared to be unhybridized Kern River rainbow trout: Rattlesnake Cr. (in Sequoia National Park), Kern River at Kern Flat, Kern River above Rattlesnake Creek, Boreal Creek, Chagoopa Creek, Kern River at Upper Funston Meadow, Kern River above Redspur Creek, and Kern River at Junction Meadow. These populations in the middle of the historic range lacked apparent influence from California golden trout (either anthropogenic or natural) that was seen in the upper sections of the Kern and also lacked apparent rainbow trout hybridization seen in the lower sections. While Behnke (2002) doubts that pure Kern River rainbow trout still exist in their native range, recent genetic analyses suggest that at least some unhybridized populations exist as indicated above. Much of their remaining habitat is in Sequoia National Forest (over 29 kilometers) and Sequoia National Park (over 40 kilometers). In addition, there are distinctive introduced populations in the Kaweah-Kern River and Chagoopa Creek, which have maintained their genetic identity (Stephens 2007).

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Threats The primary threats to remaining populations of Kern River rainbow trout involve interactions with non- native trout and include, (1) hybridization with hatchery rainbow trout (O. mykiss), which are still planted in the upper Kern Basin, (2) hybridization with golden trout planted or moving into their waters, and (3) competition from brown trout (Salmo trutta), brook trout (Salvelinus fontinalis), and hatchery rainbow trout. Further introductions by anglers of hatchery rainbow, brown or brook trout into the remaining small isolated streams are possible. In addition, continued grazing in riparian areas and heavy recreational use of the basin, including angling, can degrade the trout’s fragile habitat. Random natural events, such as floods, drought, and fire, can also exacerbate these problems (CalTrout 2008), especially in combination with rain-on-snow flooding associated with climate change (Herbst and Cooper 2009).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of aquatic areas.

Alternative B, C, and D: Kern River Rainbow Trout were considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

Literature Cited – Kern River Rainbow Trout Bagley, M.J. and G.A.E. Gall. 1998. Mitochondrial and nuclear DNA sequence variability among populations of rainbow trout (Oncorhynchus mykiss). Molecular Ecology 7:945-961.

Behnke, R. J. 1992. Native trout of western North America. American Fisheries Society Monograph 6. xx + 275 pp.

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Behnke, R.J. 2002. Trout and Salmon of North America. New York: The Free Press. 359 pp.

Berg, W. J. 1987. Evolutionary genetics of rainbow trout, Parasalmo gairdnerii (Richardson). Ph.D. Diss., Univ. California, Davis.

California Trout. 2008. SOS: California’s Native Fish Crisis - Status of and solutions for restoring our vital salmon, steelhead and trout populations. California Trout, San Francisco, California 94102, Pages 70-71. Available at http://caltrout.org/pdf/Kern%20River%20Rainbow%20Trout.pdf

Gold, J. R., and G. A. E. Gall. 1975. The taxonomic structure of six California high Sierra golden trout (Salmo aquabonita) populations. Proc. California Acad. Sci. 40:243-263.

Herbst, D.B. and S.D. Cooper. 2010. Before and after the deluge: rain-on-snow flooding effects on aquatic invertebrate communities of small streams in the Sierra Nevada, California. Journal of the North American Benthological Society 29:1354-1366.

Knapp, R.A. and T.L. Dudley. 1990. Growth and longevity of golden trout, Oncorhynchus aquabonita, in their native streams. California Fish, Game 76:161-173.

Knapp, R.A. and V.T. Vrendenburg. 1996. Spawning by California golden trout: characteristics of spawning fish, season and daily timing, redd characteristics, and microhabitat preferences. Transaction American Fisheries Society 125:519-531.

Matthews, K.R. 1996a. Diel movement and habitat use of California golden trout in the Golden Trout Wilderness. Transactions, American Fisheries Society, 125:78-86.

Matthews, K.R. 1996b. Habitat selection and movement patterns of California Golden trout in degraded and recovering stream sections in the Golden trout wilderness, California. North American Journal of Fisheries Management, 16:579-590.

Moyle, P.B. 2002. Inland Fishes of California. Regents of the University of California, Berkeley and Los Angeles, CA.

Moyle, P.B., J. V. E. Katz and R. M. Quiñones. 2011. Rapid decline of California's native inland fishes: a status assessment. Biological Conservation 144: 2414-2423.

NatureServe. 2012. NatureServe Explorer: An online encyclopedia of life [web application]. Version 7.1. NatureServe, Arlington, Virginia. Available http://www.natureserve.org/explorer. (Accessed: January 14, 2013).

Schreck, C. B., and R. J. Behnke. 1971. Trouts of the upper Kern River basin, California, with reference to systematics and evolution of western North American SALMO. J. Fish. Res. Board Can. 28:987-998.

Stefferud, J.A. 1993. Spawning season and microhabitat use by California golden trout (Oncorhynchus mykiss aguabonita) in the southern Sierra Nevada. California Fish and Game 79:133-144.

Stephens, M.R. 2007. Contribution of population genetic structure, hybridization, and cultural attitudes to the conservation of threatened native trout. PhD Dissertation. Davis, University of California.

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Stephens, S., D.P. Christenson, M. Lechner, and H. Werner. 1995. Upper Kern basin fishery management plan. California Department of Fish and Game, Sequoia National Forest, and Sequoia National Park.

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Tehachapi fritillary butterfly (Speyeria egleis tehachapina) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account The Tehachapi fritillary subspecies has a very limited distribution in the Pacific Southwest Region. Emmel and Emmel (1973) assessed it as “one of the rarest butterflies in North America.” NatureServe indicates that it is found only in Kern County, in the Tehachapi Mountains and Piute Mountains, at elevations between 7,000 to 8,400 feet. Davenport (1983) states that the subspecies is limited to summit peaks and ridges in those mountain ranges. On public lands, the butterfly subspecies has only been found in the Sequoia National Forest. The Great Basin fritillary (Speyeria egleis) is the nominate species and is widely distributed throughout the western United States (Figure 24).

Population trends are uncertain, but Davenport (2004) states that this “rare subspecies appears to be in a serious decline” and indicates that there have been no records for the butterfly in the Tehachapi Mountains since August 1, 1998. However, he indicates that there have been records from the Piute Mountains since that date. NatureServe (2011) indicates that the subspecies has a very limited range in two mountain ranges in Kern County but states that the butterfly apparently is “fairly common in both ranges.” NatureServe (2011) further states that the “distribution data for U.S. states and Canadian provinces is known to be incomplete or has not been reviewed for this taxon” and that collectors may be a threat to the populations of this butterfly.

Distribution of the nominate species Speyeria egleis in the United States. The map (Figure 24) is reproduced from the web site for Butterflies and Moths of North America (BOMONA), http://www.butterfliesandmoths.org/species/Speyeria-egleis. The Tehachapi fritillary butterfly is restricted to localized populations in the Tehachapi and Piute mountain ranges within the Transverse Range of southern California.

Habitat Status Typical habitats where the Tehachapi fritillary occurs are mountain meadows, forest openings and exposed rocky ridges. According to Davenport et al. (2006) and Emmel and Emmel (1973) the subspecies flies from late July to early August. All fritillary butterflies produce a single brood produced each year over the summer months (Shapiro 2011). Males patrol during the day for females, which lay eggs on leaf litter near violets. First-stage or instar caterpillars hibernate unfed until the following spring, when they feed on violet leaves. The larval food plant is thought to be Viola purpurea xerophyte, however Viola adunca, V. nuttallii and V. walteri may also be host plants.

Shapiro (1996) noted that Speyeria egleis may commonly be seen exhibiting “hilltopping” behavior where individuals form aggregations, a strategy utilized by localized and low-density populations to more effectively find mates for reproduction. Males may patrol a territory or perch, basking in the sun. Females are transient visitors and depart after mating.

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Figure 24. Distribution of the Great Basin fritillary, the nominate species for Tehachapi fritillary, in the western United States

Threats The violet food plants and habitat of the Tehachapi fritillary butterfly are susceptible to destruction from wildfire. Occurrences of the violet species associated with this butterfly should be identified and protected. Miller and Hammond (2007) provide some general guidelines on how to manage habitats occupied by a variety of butterfly species.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of riparian areas.

Alternative B, C and D: Tehachapi fritillary butterfly was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian areas and meadows is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

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Sequoia NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian and meadow areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Literature Cited – Tehachapi Fritillary Butterfly Davenport, K. 1983. Geographic distribution and checklist of the butterflies of Kern County, California. Journal of the Lepidopterists’ Society 37(1): 64.

Davenport, K. 2004. A concise update of the information provided in The Butterflies of Southern California (1973) by T.C. Emmel & J.F. Emmel. The Taxonomic Report of the International Lepidoptera Society 4(7): 11.

Davenport, K.E., R.E. Stanford, & R.L. Langston. 2006. Flight periods of California butterflies for “resident species”, subspecies and most strays to the State. The International Lepidoptera Survey Newsletter 7: 51.

Emmel, T.C., & J.F. Emmel. 1973. The Butterflies of Southern California. Natural History Museum of Los Angeles County, Science Series #26. p. 31.

Miller, J.C. and P.C. Hammond. 2007. Butterflies and Moths of Pacific Northwest Forests and Woodlands: Rare, Endangered, and Management-Sensitive Species. Forest Health Technology Enterprise Team, Technology Transfer Species Identification. FHTET-2006-07. See http://www.fs.fed.us/foresthealth/technology/

NatureServe. 2011. NatureServe Explorer species database records. Located at http://www.natureserve.org/explorer/servlet/NatureServe?searchSciOrCommonName=Speyeria+e gleis+tehachapina&x=9&y=10

Shapiro, A.M. 1996. Status of Butterflies. Chap. 27 In: Sierra Nevada Ecosystem Project: Final Report to Congress, Vol. II: Assessments and scientific basis for management options. University of California, Centers for Water & Wildlife Resources, Davis, CA.

Shapiro, A. 2011. Art Shapiro’s Butterfly Site: Monitoring butterfly populations across Central California for more than 35 years. http://butterfly.ucdavis.edu/

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Apache Silverspot Butterfly (Speyeria nokomis apacheana) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account Apache silverspot butterfly Speyeria nokomis apacheana i is found on the east slope of the Sierra Nevada Mountains in Alpine, Inyo, and Mono Counties where it occurs in marshes and wet meadows near springs, seeps and riparian areas. There are records from the Inyo, Stanislaus and Toiyabe national forests. The Apache silverspot butterfly represents a member derived from basically a desert fauna that has been in retreat since the drying trend began at the end of the Pleistocene Epoch about 1.5 million years ago (Shapiro 1996).

Distribution of the nominate species Nokomis fritillary Speyeria nokomis in the western United States. The map (Figure 25) is reproduced from the web site for Butterflies and Moths of North America (BOMONA), http://www.butterfliesandmoths.org/species/Speyeria-nokomis. The Apache silverspot butterfly subspecies is restricted to a few localized sites in the eastern Sierra Nevada Mountains.

Figure 25. Distribution of the Nokois fritillary, the nominate species for Apache silverspot butterfly, in the western United States

Habitat Status The Apache silverspot butterfly occurs on the east slope of the Sierra Nevada Mountains and inhabits marshes and wet meadows near springs, seeps and riparian areas (Fleishman et al 2002, Britten et al. 2003). Since these habitats are highly localized, minimal migration occurs between populations.

Typical habitats where Apache silverspot adults may be observed from late July to early September (Emmel and Emmel 1973, Davenport et al. 2006) are mountain meadows, forest openings and exposed rocky ridges. All fritillary butterflies produce a single brood each year over the summer months (Shapiro

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2011). Males patrol during the day for females, which lay eggs on leaf litter near their violet larval food plants. First-stage or instar caterpillars hibernate without previous feeding until the following spring, when they feed on violet leaves. The larval food plant is the violet, Viola nephrophylla.

Fleishman et al (2002) developed a logistic regression model to identify environmental variables, listed in order of importance, that were most strongly associated with the occurrence of Apache silverspot butterflies: 1) distance to nearest occupied location (negative correlation), 2) distance to water (negative correlation), 3) high vernal equinox insolation (positive correlation), 4) east facing (positive correlation), and 5) site is relatively flat (positive correlation). They also observed that bull thistle Cirsium vulgare tended to be present in colonized patches. Patches that went extinct were in close proximity to other extirpated sites, lacked the lavender thistle Cirsium neomexicanum, and had a higher percent cover of live vegetation and litter. Presence of both Carcuus nutans and Cirsium scariosum thistle species had no significant association with the presence of Apache silverspot butterflies.

Threats Fleishman et al. (2002) suggest some management considerations to mitigate impacts to this butterfly such as switching periods of heaviest livestock grazing from early summer to late summer or early autumn to prevent soil compaction, depression of the water table and conversion from mesic to xeric vegetation.

The violet food plants and habitat of the Apache silverspot butterfly are susceptible to destruction from wildfire (Shapiro 1996), water diversions such as in the Owens Valley (Hammond and McCorkle 1983) and grazing (Fleishman et al. 2002), or other activities that alter riparian vegetation. Habitat for the Oregon silverspot butterfly, S. zerene hippolyta, which occurs in meadows along the Oregon coast, was compromised when fire was excluded, allowing encroachment by trees and accumulation of grass thatch that smothered violet food sources (Hammond and McCorkle 1983). Occurrences of the violet species associated with this butterfly should be identified and protected. Miller and Hammond (2007) provide some general guidelines on how to manage habitats occupied by a variety of butterfly species.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and it allows for the continued protection of aquatic areas.

Alternative B, C, and D: Apache silverspot butterfly was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

196 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests contributed to the existing baseline and are described in the draft EIS. Therefore only future actions would be considered. Future actions will be addressed in the specific project level NEPA.

Determination Statement Inyo NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

Literature Cited – Apache Silverspot Butterfly Britten, H.B., E. Fleishman, G.T. Austin, and D.M. Murphy. 2003. Genetically effective and adult population sizes in the Apache silverspot butterfly, Speyeria nokomis apacheana (Lepidoptera: Nymphalidae). Western North American Naturalist 63(2): 229-235.

Davenport, K.E., R.E. Stanford and R.L. Langston. 2006. Flight periods of California butterflies for “resident species”, subspecies and most strays to the State. The International Lepidoptera Survey Newsletter 7: 51.

Emmel, T.C., & J.F. Emmel. 1973. The Butterflies of Southern California. Natural History Museum of Los Angeles County, Science Series #26. p. 31.

Fleishman, E., D.D. Murphy and P. Sjögren-Gulve. 2002. Modeling species richness and habitat suitability to taxa of conservation interest. Chaper 45 in Scott, J.M, P.J. Heglund and M.L. Morrison, eds. Predicting Species Occurrences: Issues of Accuracy and Scale. Island Press, Washington, D.C. 873 pp.

Hammond, P.C. and D.V. McCorkle. 1983. The decline and extinction of Speyeria populations resulting from human environmental disturbances (Nymphalidae: Argynninae). The Journal of Research on the Lepidoptera 22:217-224.

NatureServe. 2011. NatureServe Explorer species database records. Located at http://www.natureserve.org/explorer/servlet/NatureServe?searchSciOrCommonName=Speyeria+e gleis+tehachapina&x=9&y=10

Shapiro, A. 2011. Art Shapiro’s Butterfly Site: Monitoring butterfly populations across Central California for more than 35 years. http://butterfly.ucdavis.edu/

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Western Pearlshell (Margaritifera falcata) Regional Foresters Sensitive Species Proposed Species of Conservation Concern

Species Account The western pearlshell is a freshwater mussel belonging to the family Margaritiferidae and is the only species in this genus present in the western United States. NatureServe (http://www.natureserve.org/explorer/, updated 2/19/08) reports the status of western pearlshell is G4G5 and short-term trends are “rapidly declining to declining (decline of 10-50 percent).”

In North American, western pearlshell (Gould, 1850) was historically found in Pacific drainages from southern Alaska, through British Columbia and south to central California (Figure 26). To the east, the distribution extended into western Montana, western Wyoming and northern Utah, where it occurs east of the Continental Divide in the headwaters of the Missouri River (Nedeau et al. 2009). Many populations appear to have been severely reduced in size from dense beds to a few isolated individuals in flow refugia at many historical sites (Vannote and Minshall 1982, Hovingh 2004, Strayer et al. 2004, Howard 2008, 2010). Hovingh (2004) concluded that western pearlshell had been extirpated from eastern California, Nevada and Utah based on from field collections and museum records from over 2,900 sites in the Great Basin.

According to Taylor (1981), in California, this species historically occurred in the Lower Klamath, Pit, Upper Kern and Sacramento rivers, Clear and Goose lakes, Pajaro-Salinas system (only in streams of the southern Santa Cruz Mts.), Lahontan Basin, Central Valley and north coast streams.

Howard (2008, 2010) compiled museum and literature records for western pearlshell at 27 historical sites, five of which are on National Forest lands. She resurveyed 20 of these sites and found that the species still occurred at six, but had apparently been extirpated, since no mussels were observed, from 14 historical sites. On National Forest lands, 15 historical sites were resurveyed: six were still occupied, the remaining nine appear to have been extirpated. She also noted that there were no historical records of western pearlshell in southern California (i.e. outside of the Central Valley). The southern-most historical records for western pearlshell in California are from the Kern River, just west of the border of the Sequoia National Forest. A map displaying the historical (Figure 26) and post-1995 discovered (Figure 27) sites in California was compiled by Howard (2010). In Figure 26, the different colored dots represent the various species found at the historical sites. Note that western pearlshell was not historically found in Southern California. National Forest lands are displayed in green; urban areas in gray. Freshwater ecoregional boundaries are defined by Abell et al. (2008).

In addition to the strategic surveys of historical sites described above, a broad-scale survey of suitable habitat in river main stems, lakes and reservoirs at 244 sites in northern Sierra Nevada national forests were surveyed was undertaken from 2003 to 2007. Howard (2008) surveyed 131 historical and recent sites on the Eldorado, Plumas and Tahoe National Forests and Lake Tahoe Basin Management Unit, and 113 sites were surveyed on the Lassen National Forest (Brim Box et al. 2005, Howard 2010). Howard (2008) reported that western pearlshells were found at 21 sites located on the Eldorado (2 sites), Lassen (7 sites), Plumas (6 sites) and Tahoe (6 sites) National Forests. Davis (2008) reported western pearlshell occurring in 13 of 55 river reaches surveyed in the Klamath and Salmon Rivers in northern California. While dense populations still occur in the Upper Truckee River (Entrix 2007), the dense Margaritifera beds in the Lower Truckee River described by Murphy (1942) appear to have been decimated with a decline from an estimated population size of at least 20,000 to no more than 150 individuals (Howard 2008).

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Figure 26. Map of 113 historical freshwater mussel sites in California based on 400 historical records compiled by Howard (2010)

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Figure 27. Map of 105 sites recently surveyed or compiled by Howard (2010)

In Figure 27 above, the different colored dots represent the species found at the sites. Note that the black dots indicate that no mussels were found at those sites.

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Figure 28. Distribution of western pearlshell (Gould, 1850) in North America based on historical and current records compiled by the Xerces Society, Portland, Oregon

Figure 28 provided by Sarina Jepson, Endangered Species Program manager for the Xerces Society, Portland, Oregon. Photo of western pearlshell from Long Creek, Lake County, Oregon by Tom Grace on August 5. 2006. It is 80 millimeters in length.

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Howard et al. (2015) resurveyed 22 of the 31 sites where western pearlshell historically occurred, and found live individuals at 13 of those sites (59 percent). Although surveys for M. falcata have not been conducted on all 18 national forests in California, there is a high likelihood that additional occupied sites occur in areas of suitable habitat in river systems in close proximity to historical sites that contain salmonid fish species because of the close association between mussel larvae, known as glochidia, and their fish hosts, which provide a source of sustenance and a means of dispersal for immature clams.

Habitat Status Western pearlshell occurs in habitats ranging in size from small creeks, 1 to 2 meters wide, up to large rivers, where ever substrates are primarily composed of clean, coarse gravel, cobble and boulders. Optimal habitats for western pearlshell are low gradient (i.e. less than 4 percent, Howard 2010) pools with velocities ranging from about 25 to 30 centimeters per second and depths from 20 to 60 centimeters (Howard and Cuffey 2003, Stone et al. 2004). In a study on the South Fork of the Eel River, Howard and Cuffey (2003) found that these mussels were almost exclusively confined to pools and near banks, particularly where sedge root mats (Carex nudata) occurred. They are often associated with eddies behind boulders or submerged logs. Historically, this species formed dense mussel beds composed of several thousand individuals. Currently, populations are in decline, often resulting in the presence of a few, scattered individuals, mostly confined to flow-refugia in the main river channel where shear stress is minimal and sedge root mats at the river margin. Flow refugia that are protected from scouring flows are essential to maintain suitable habitat because individual mussels are incapable of repositioning themselves after displacement from the substrate during high flows or after burial by sediment deposits (Vannote and Minshall 1982, Strayer 1999, Allen and Vaughn 2010). Western pearlshell is often more abundant on the outside of river bends, where water velocities were elevated, and individuals are often tightly grouped between or downstream of large, woody debris (Spring Rivers Ecological Sciences 2007).

The lifespan is usually several decades and individuals are capable of living for over 100 years. Populations may be either dioecious (with separate males and females), or much less often, hermaphroditic (combined male and female) (Nedeau et al. 2009). In the one section of the Pit River, during 2004 to 2005 gravid (mature egg-bearing) individuals were observed in April, June and July (Ellis and Haley 2005, Haley 2007, Spring Rivers Ecological Sciences 2007). It appeared that only 30 to 40 percent of individuals in a population reproduced in any given season. During these periods, the flow regime was usually stable at 200 cubic feet per second, although flows up to 2,000 cubic feet per second, occurred during 2005; water temperatures ranged from 50 to 60.8 degrees Fahrenheit. Spawning M. falcata have been observed in April in the South Fork Eel River in California, when water temperatures exceeded 50 degrees Fahrenheit, and mid-May in the Truckee River, California, when water temperatures exceeded 45 degrees Fahrenheit (Murphy 1942, Spring Rivers Ecological Sciences 2007). In cold spring- fed areas of Hat Creek, individuals with mature gonads occurred only in July; none were observed from February to June. Development of embryos apparently was delayed in Hat Creek because of cooler water temperatures.

During spawning, thousands of larval clams, known as glochidia, are extruded by spawning adults as crescent-shaped packets or glutinates onto the adjacent substrate. Fish are attracted to the glochidial deposits, which are ingested and broken open allowing the glochidia larvae to attach to the fish. The glochidia usually attach to fish belonging to the family Salmonidae (i.e. cutthroat, rainbow, brook or brown trout, salmon) to complete their metamorphosis to an adult clam, however they may also attach to other species such as speckled dace, Tahoe sucker, Lahontan redside (Howard 2008). It is not known whether clams are capable of development on all of these species. Since these mussels have an obligate dependence on their fish hosts for development and dispersal, they will only occur in streams and rivers where these fish are present. After residing as a parasite on the fish host for up to 36 days at 57.6 degrees Fahrenheit (Murphy 1942), larval clams detach from the fish host and fall to the bottom and take up a

202 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests benthic existence as filter feeders. The period of encystment may be shorter at higher water temperatures. Once individuals have positioned themselves in the substrate, further movement is minimal.

Threats Threats to western pearlshell and its habitat include eutrophication due to agricultural runoff and urbanization, sedimentation that smothers mussel beds, water diversions that reduce instream flows and reduce available oxygen, introduction of exotic species, grazing, and water impoundments that reduce current velocities, manipulate water levels and allow for sediment deposition (Frest and Johannes 1995, Howard 2008). Manipulations of flow regimes associated with operation of hydropower dams also pose a threat (Spring Rivers Ecological Sciences 2007). The decline or extirpation of host fish species, such as Lahontan cutthroat and other trout, which serve as dispersal agents for the larvae, have also negatively impacted the species. As described by Howard (2010) and Howard and Revenga (2009), fishes in all freshwater ecoregions have suffered severe declines and have high percentages of imperiled fishes. For example, 89 percent of the fish in the Death Valley ecoregion are considered imperiled by the American Fisheries Society; 75 percent in the Sacramento San Joaquin ecoregion; 65 percent in the Lahontan; 60 percent in the Southern California region; 50 percent in Northern California and 40 percent in the Colorado ecoregions. Because fishes serve as hosts for larval freshwater mussels, this degree of imperilment of fishes is likely to have a great impact on recruitment and will therefore likely result in declining populations of mussels.

Impacts to the western pearlshell from human-caused activities include eutrophication due to agricultural runoff and urbanization, sedimentation that smothers mussel beds, water diversions that reduce and alter instream flow regimes, mining, including suction dredge operations, introduction of exotic species, grazing, and water impoundments that reduce current velocities and allow for sediment deposition (Hovingh 2004, Lydeard et al. 2004, Strayer et al. 2004, Strayer and Downing 2006, Krueger et al. 2007). This mussel species depends on salmonid fish hosts to sustain and disperse larval clams. Since many salmonid species such as rainbow trout and salmon have experienced severe declines, western pearlshell mussels have declined as well.

For aquatic mollusks and many aquatic invertebrates in general, the following types of disturbances should be considered as threats:

 Impaired water quality and loss of suitable habitat due to excessive sediment deposition or erosion, and elevated water temperature or low dissolved oxygen.  Water diversions for irrigation, hydropower generation and livestock watering and feeding in riparian and aquatic habitats, resulting in altered discharge, vegetative cover and loss of suitable habitat.  Dam construction which alters flow regimes, unnaturally fluctuates discharge, often lowers the availability of oxygen and allows fine sediments to accumulate.  Excessive sedimentation from a variety of activities such as logging, mining, road and railroad grade construction, and grazing may smother substrates causing death by preventing feeding and movement, and obstructing gills (Hovingh 2004, Vannote and Minshall 1982, Webb et al. 2008, Bettaso and Goodman 2010).  Margaritifera falcata is noteworthy for an inability to reassume a feeding position after it has been dislodged (Vannote and Minshall 1982). As a result, this species is limited to flow refugia between large boulders and in root wads at river margins where shear stresses are attenuated (Strayer 1999, Allen and Vaughn 2010).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Western pearlshell was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

Determination Statement Inyo and Sequoia NFs: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

Literature cited – Western Pearlshell Abell, R., M.I., Thieme, C. Revenga, M. Bryer, M. Kottelat, N. Bogutskaya, B. Coad, N. Mandrak, S. Contreras, W. Bussing, M.I. Stiassny, P. Skelton, G.R. Allen, P. Unmack, A. Naseka, R. Ng, N. Sindorf, J. Robertson, E. Armijo, J.V. Higgins, T.J. Heibel, E. Wikramanayake, D. Olson, H.I. Lopez, R.E. Reis, J.G. Lundberg, M.H. Sabaj Perez, and P. Petry. 2008. Freshwater ecoregions of the world: A new map of biogeographic units for freshwater biodiversity conservation. Bioscience 58: 403-414.

Allen, D.C and C.C. Vaughn. 2010. Complex hydraulic and substrate variables limit freshwater mussel species richness and abundance. Journal of the North American Benthological Society 29:383- 394.

Bettaso, J.B. and D.H. Goodman. 2010. A comparison of mercury contamination in mussel and ammocoete filter feeders. Journal of Fish and Wildlife Management. Available at http://www.fwspubs.org/doi/pdf/10.3996/112009-JFWM-019.

Brim Box, J., S. Chappell, M. McFarland and J. Furnish. 2005. The aquatic mollusk fauna of the Lassen National Forest in northeastern California. USFS PSW Regional Office Report, Vallejo, CA. 117 pp.

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Davis, E. 2008. Freshwater Mussel Abundance, Distribution, and Habitat Preference in two Northern California Rivers within Karuk Ancestral Territory. Undergraduate Thesis, Departments of Biology and Environmental Studies, Whitman College, Walla Walla, WA.

Ellis, M.J. and L.E. Haley. 2005. Reproductive timing of freshwater mussels and the potential impacts of pulsed flows on reproductive success, Spring Rivers Ecological Sciences, Cassel, CA. Paper presented at the Pulsed Flow Conference, July 15, 2005, Davis, CA.

Entrix. 2007. Draft Freshwater Mussel Report for Sunset Stables Restoration and Resource Management Plan Project. A report Prepared for the California Department of General Services Real Estate Services Division and California Tahoe Conservancy, Sacramento, CA. 24 pp.

Frest, T.J. and E.J. Johannes. 1995. Freshwater mollusks of the upper Sacramento system, California, with particular reference to the Cantara Spill. 1994 yearly report by Deixis Consultants (Seattle, WA) to California Department of Fish and Game, iii+88 pages and appendices.

Haley, L.E. 2007. Reproductive timing of freshwater mussels and potential impacts of pulsed flows on reproductive success. PowerPoint presentation at the Pulsed Flow Program Conference, UC Davis, September 18, 2007. Available at http://animalscience.ucdavis.edu/Pulsedflow/Haley%202007.pdf

Hovingh, P. 2004. Intermountain freshwater mollusks, USA (Margaritifera, Anodonta, Gonidea, Valvata, Ferrissia): geography, conservation, and fish management implications. Monographs of the Western North American Naturalist 2:109-135.

Howard, J. 2008. Strategic inventory of freshwater mussels in the northern Sierra Nevada Province. Final Report by Western Mollusk Sciences, San Francisco, CA to PSW Regional Office, Vallejo, CA. 65 pp.

Howard, J. 2010. Sensitive freshwater mussel surveys in the Pacific Southwest Region: Assessment of Conservation Status. Final Report by Western Mollusk Sciences, San Francisco, CA to U.S. Forest Service, PSW Regional Office, Vallejo, CA. 60 pp

Howard, K. A. and K. M. Cuffey. 2003. Freshwater mussels in a California north Coast Range river: occurrence, distribution, and controls. Journal of the North American Benthological Society 22: 63-77.

Howard, J.K. and C. Revenga. 2009. California’s freshwater biodiversity in a continental context. Science for Conservation Technical Brief Series. The Nature Conservancy of California. San Francisco, CA. 29 pp + Appendices.

Howard, J.K., J.L. Furnish, J. Brim Box, and S. Jepsen. 2015. The decline of native freshwater mussels (Bivalvia: Unionoida) in California as determined from historical and current surveys. California Fish and Game 101(1):8-23.

Krueger, K., P. Chapman, M. Hallock and T. Quinn. 2007. Some effects of suction dredge placer mining on the short-term survival of freshwater mussels in Washington. Northwest Science 81:323-332. Available at http://www.bioone.org/doi/pdf/10.3955/0029-344X-81.4.323.

Lydeard, C., R.H. Cowie, W. F. Ponder, A. E. Bogan, P. Bouchet, S.A. Clark, K.S. Cummings, T.J. Frest, O. Gargominy, D.G. Herbert, R.Hershler, K.E. Perez, B. Roth, M. Seddon, E.E. Strong, and F.G. Thompson. 2004. The global decline of nonmarine mollusks. Bioscience 54:321-330. This article

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is available at http://taddeo.ingentaselect.com/vl=4683764/cl=29/fm=docpdf/nw=1/rpsv/cw/aibs/00063568/v54 n4/s9/p321.

Murphy G. 1942. Relationship of the freshwater mussel to trout in the Truckee River. California Fish Game 28:89–102.

Nedeau, E.J., A.K. Smith, J. Stone, and S. Jepsen. 2009. Freshwater Mussels of the Pacific Northwest. Second Edition. The Xerces Society, Portland, OR. 60 pp.

Spring Rivers Ecological Sciences. 2007. Reproductive Timing of Freshwater Mussels and Potential Impacts of Pulsed Flows on Reproductive Success. California Energy Commission, PIER Energy Related

Environmental Research Program. CEC-500-2007-097 Available at http://animalscience.ucdavis.edu/Pulsedflow/project03.htm

Stone, J., S. Barndt, and M. Gangloff. 2004. Spatial distribution and habitat use of the western pearlshell mussel (Margaritifera falcata) in a western Washington stream. Journal of Freshwater Ecology 19: 341-352.

Strayer, D. 1999. Use of flow refuges by unionid mussels in rivers. Journal of the North American Benthological Society 18:468-476.

Strayer, D. L., J.A. Downing. 2006. Challenges for freshwater invertebrate conservation. Journal of the North American Benthological Society 25:271–287.

Taylor, D.W. 1981. Freshwater mollusks of California: a distributional checklist. California Fish and Game 67:140-163.

Vannote, R.L., and G.W. Minshall. 1982. Fluvial processes and local lithology controlling abundance, structure, and composition of mussel beds. Proceedings of the National Academy of Sciences 79:4103-4107.

Webb, K., C. Craft and E. Elswick. 2008. The evaluation of the freshwater western pearl mussel, Margaritifera falcata (Gould, 1850), as a bioindicator through the analysis of metal partitioning and bioaccumulation. Northwest Science 82:163-173.

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Owens Valley Springsnail (Pyrgulopsis owensensis) Regional Foresters Sensitive Species

Species of Conservation Concern

Species Account Owens Valley Springsnail is found only on the Inyo National Forest. The species is found along escarpments of the White and Inyo Mountains on the east side of the Owens Valley.

Habitat Status This species occurs in small springs and spring runs, typically on watercress (Rorippa), travertine deposits or stones in the foothills of the eastern Sierra Nevada Mts. This species co-occurs with Wong’s springsnail (P. wongi) at Batchelder Springs.

Threats For aquatic mollusks and many aquatic invertebrates in general, the following types of disturbances should be considered as threats:

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Owen’s Valley springsnail was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

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Figure 29. Areas where Owens Valley springsnail occurs. Map was prepared by Kurt Fesenmyer of Trout Unlimited, Boise, ID in 2013

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Determination Statement Inyo NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. This species is considered stable across the Inyo NF, but it only known from this area. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

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Wong’s Springsnail (Pyrgulopsis wongi) Regional Foresters Sensitive Species

Proposed Species of Conservation Concern

Species Account This species is only know from the Inyo National Forest. Wong’s springsnail has a widespread distribution in the Owens Valley along the eastern escarpment of the Sierra Nevada Mountains. They range from Pine Creek south to Little Lake, and along the eastern side of the valley from French Spring to Marble Creek in the Inyo Mountains. It is also found in a few sites in Long, Adobe, and Deep Springs Valleys. It is considered locally abundant by highly restricted distribution.

Figure 30 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot. The distribution was originally thought to be limited in the Great Basin of California and Nevada (Hershler 1998). In fact, it occurs in Owens Valley (where most of the populations occur and nine other drainage basins from Rose Valley in central California to Mono Lake and the Carson River Valley in Nevada (Liu and Hershler 2007). The map (Figure 31) was prepared in 2013 by Kurt Fesenmyer of Trout Unlimited, Boise, ID.

Habitat Status Habitat for this species includes seeps and spring-fed streams of small to moderate size. Temperature requirements range from 49.1 to 71.6 degrees Fahrenheit. The snails are typically found commonly in watercress (Rorippa) and /or on small bits of travertine and stone (Hershler 1989). It is only known to occur in flowing water. Spring habitat that has previously been altered by spring-improvements, grazing or other impacts would alter the water quality of the spring and would preclude occurrence of this species. Each population of snail is endemic to the spring it inhabits, and since these snails are obligatory aquatic throughout their entire life, they cannot disperse to other springs, nor can springs where snails have been extirpated be re-colonized. This species co-occurs with Owens Valley springsnail (P. owensensis) at Batchelder Springs.

Threats Historically, there were major impacts to habitats occupied by this species from grazing, and water diversions for mining operations. Water impoundments degrade habitats because this snail requires running water.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Wong’s sringsnail was considered as a species of conservation concern in the creation of the draft EIS, as stated in Chapter 2, and shown in Chapter 3. Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

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Figure 30. Map of Wong’s springsnail from the NRIS Aquatic Survey Database, 2016

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Figure 31. The distribution of Wong’s springsnail in the Great Basin of California and Nevada

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Determination Statement Inyo NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. This species is considered stable across the Inyo NF, but it only known from this area. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

Literature Cited – Wong’s Springsnail Hershler, Robert. 1989. Springsnails (Gastropoda: Hydrobiidae) of Owens and Amargosa River (Exclusive of Ash Meadows) Drainages, Death Valley System, California-Nevada. Proc. Biol. Soc. Wash. 102(1), 1989, pp 176-248.

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Sensitive Species not selected as Species of Conservation Concern

Northern Goshawk (Accipiter gentilis) Regional Forester Sensitive Species

Species Account Northern goshawks occupy boreal and temperate forests throughout the Holarctic zone (Squires and Reynolds 1997). The northern goshawk is a year-round resident throughout many higher elevation areas of California. Within California, this species occurs in the northern Coast Ranges, the Klamath and Siskiyou Mountains, across the Cascades, Modoc Plateau, and Warner Mountains, and south through the Sierra Nevada (Shuford and Gardali 2008, Zeiner et al. 1990). Goshawks may also inhabit suitable habitats in the Transverse Ranges and other mountainous areas in southern California (Zeiner et al. 1990).

In the NRIS database, the Inyo NF has 557 records, Sequoia NF has 331 records, and the Sierra NF has 848 records. Figure 32 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status The northern goshawk is associated with the use of older-age conifer, mixed, and deciduous forests. Forest stands with high suitability contain an availability of large live trees for nesting, a closed canopy for protection and thermal cover, and open space in the understory for maneuverability and flight (Hargis et al. 1994, Squires and Kennedy 2006). Northern goshawks forage within a wider range of forest types and conditions. Large snags and downed logs are considered important components within foraging habitat because such features benefit various prey species (Reynolds et al. 1992).

Nesting chronology varies annually and by elevation. In general, nesting is initiated in February with nest construction, egg-laying, and incubation occurring through May and June (Dewey et al. 2003). The average incubation period is approximately 33 days and the nestling period typically extends from early June through early July, with most young fledged by mid-July. Young birds hatch and begin fledging in late June and early July and are independent by mid-September. Woodbridge and Hargis (2006) state the post-fledging dependency period extends until mid/late August. Goshawk nests are generally constructed in live trees and are usually among the largest trees in the stand. Nest trees averaged 32 inches in diameter at breast height in the Lake Tahoe region, 34 inches in diameter at breast height in the Inyo National Forest, and 51 inches in diameter at breast height in Yosemite National Park (USDA 2001). Nest sites located in protected activity centers near the TRRP Project have similarly occurred in large live trees where 4 nest sites had tree diameters at breast height of 102, 69, 55 and 137 inches.

Northern goshawk nesting habitat at the nest stand scale has consistently greater canopy cover, greater basal area, greater numbers of large diameter trees, fewer small diameter trees, less understory cover, and gentle to moderate slopes relative to non-used, random sites (USDA 2001). Canopy cover, based on mean values reported, the range extends from 31 percent (sd =13) reported on the Inyo National Forest (20 sites) to 70.4 percent (se =3.1) reported in the Lake Tahoe region (35 sites). Other estimations reported for eastside pine on the Lassen and Modoc National Forests also fall this range with a mean 64 percent. In Yosemite National Park, Maurer (2000) found that northern goshawk nest sites (33 sites) averaged 65 percent (sd =15, range 39 to 100 percent).

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Figure 32. Map of northern goshawk records within the NRIS Wildlife Database, 2016

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McGrath et al. (2003) found that goshawks in the Interior Northwest nested, at the 0.4 acre scale, on the lower one-third or bottom of north facing slopes in stands characterized by relatively higher basal area, higher quadratic mean diameter, greater canopy closure, and greater live stem densities, compared to random sites. Goshawks nesting in the relatively open-canopied and xeric stands found on the eastern slopes of the Sierra Nevada in the Inyo National Forest selected nest stands with a mean canopy closure of 29 percent (Hargis et al. 1994). Variability in the structural characteristics of nest stands between studies appears to be related to differences in vegetation type and geographic region. Within the Lake Tahoe region of the Sierra Nevada, Keane (1999) found that nest-site areas (0.25 acre) were characterized by high canopy closure (mean=70.4 percent, se=3.1, canopy measured above 9.8 feet), high densities of live trees in greater than 24-40 inch (mean=22.1 trees per acre, se=3.2) and greater than 40 inch (mean=15.8 trees/acre, se=2.2) diameter at breast height classes, high densities of dead trees in the greater than 24 to 40 inch (mean=3.6 trees/acre, se=0.7) class, low densities of 2 to 12 inch diameter at breast height live trees (mean=121.4 trees/acre, se=12.3), and low shrub/sapling and ground cover (mean=9.9 percent, se=2.0). No difference in slope aspect was detected for nest sites (Ibid.).

The home range increases in size from the breeding season to the non-breeding season and is generally larger for males than for females throughout the year. During the breeding season, the average home range of goshawks in the Lake Tahoe area is 6,745 acres for males and 5,040 acres for females. Non- breeding season home ranges averaged 23,448 acres for males and 13,888 acres for females (Keane 1999). The mean breeding home range size for females from the Inyo National Forest estimated at 3,300 acres (USDA 2001). Home ranges include areas with a greater proportion of larger tree size classes and higher density classes than that randomly available across the landscape. The area within the home range, but outside the PFA, is often referred to as the foraging area (Reynolds et al. 1992). Maintaining requisite habitat elements can be best accomplished by managing large tracts of forests as sustainable ecological units where forest successional processes are continually moving a number of stands, within the natural range of variability, through the late seral stages preferred by this species (Graham et al. 1999, DeStefano et al. 1994).

Northern goshawks have evolved morphological adaptations for capturing prey in forested environments, but are also capable of ambushing prey in open habitats (Squires and Reynolds 1997). Moderately dense, mature conifer forests are generally the preferred foraging habitat for this species (Ibid.). However, goshawks also forage in a variety of other forest age-classes, structures and compositions, and into openings and along forest edges (summarized in Reynolds et al. 2006). In California, mature and old growth habitat (at least 20.8 inches diameter at breast height, canopy closure at least 40 percent) were used, whereas open habitats such as meadows and early seral areas were avoided in mixed-conifer forests (Austin 1993). In Arizona, Beier and Drennan (1997) found that goshawks foraged in stands that had higher canopy closure, greater tree density, and a greater density of large trees (greater than 16.2 inches diameter at breast height) than on contrast plots. Prey availability rather than prey abundance, within suitable foraging habitats, appears to be more important to habitat use by this species (Reynolds et al. 2006).

Reynolds and Meslow (1984, in USDA 2001) found that the goshawk is a height zone generalist, taking prey from the ground-shrub and shrub-canopy layers. Some authors suggest that goshawks may forage along edge environments created between dense forests and adjoining habitats such as brush fields, plantations, meadows, streams, and some instances along roads.

Northern goshawks are known to prey on over 50 species of birds and mammals throughout their western range (Graham et al. 1994). Prey size varies little between geographic regions (Boal and Mannan 1994). The key species or species groups that are more prevalent in goshawk diets in the Sierra Nevada include Douglas squirrel, Spermophilus spp. (golden-mantled squirrel, Belding squirrel, and California ground

216 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests squirrel), chipmunks (Tamias spp.), Steller’s jay, northern flicker, and American robin (USDA 2001). In the Lake Tahoe region, primary prey species include Douglas squirrel (Tamiasciurus douglasii), Steller’s jay (Cyanocitta stelleri), northern flicker (Colaptes auratus), and ground squirrel (Spermophilus spp.). Other prey species include American robin (Turdus migratorius), blue grouse (Dendragapus obscurus), other woodpeckers, and other squirrels (Keane1999). Many of these species are ground dwellers or spend a proportion of their time near the ground. Important components for foraging habitats also include an availability of snags (minimum of 3 per acre greater than 18 inches diameter at breast height) and downed logs (minimum of 5 per acre greater than 12 inches diameter at breast height) for prey populations. It has been hypothesized that relatively open shrub and lower canopy layers within forested stands may facilitate prey detection and capture by northern goshawks (USDA 2001).

The following CWHR classes provide high capability nesting habitat for this species: Jeffrey Pine, Lodgepole Pine, Montane Hardwood, and Subalpine Conifer (4M, 4D, and 5D); Montane Hardwood- Conifer, Montane Riparian, Sierran Mixed Conifer, and White Fir (4M, 4D, 5D, and 6); and Red Fir (5D). Within CWHR, size class 6 is only recognized for a subset of the forest vegetation types (Sierran Mixed Conifer, White Fir, Montane Hardwood-Conifer, Montane Riparian, and Aspen). The following vegetation types and strata provide moderate capability nesting habitat for goshawks: Aspen (4M, 4D, 5D, and 6), Eastside Pine (3M, 3D, 4M, 4D, and 5D), Lodgepole Pine (3M and 3D), Red Fir (4M and 4D), and Subalpine Conifer (3M and 3D).

The following CWHR classes provide high capability perching habitat for this species: Jeffrey Pine, Lodgepole Pine, Montane Hardwood, Montane Hardwood-Conifer, Montane Riparian, Sierran Mixed Conifer, Subalpine Conifer, and White Fir (4M and greater size and density classes); and Red Fir (5M and 5D). The following CWHR types and strata provide moderate capability perching habitat for this species: Aspen and Eastside Pine (3M and greater size and density classes); Jeffrey Pine, Lodgepole Pine, Sierran Mixed Conifer, Subalpine Conifer, and White Fir (3M, 3D, 4S, and 4P); Montane Hardwood, Montane Hardwood-Conifer, and Montane Riparian (4S and 4P); and Red Fir (4M, 4D, 5S, and 5P).

The following CWHR classes provide high capability foraging habitat for goshawk: Alpine Dwarf-Shrub (all strata); Eastside Pine (4D, 5S, 5P, 5M, and 5D); Jeffrey Pine, Lodgepole Pine, Montane Hardwood, Montane Hardwood-Conifer, Montane Riparian, Sierran Mixed Conifer, Subalpine Conifer, and White Fir (4M and greater size and density classes); and Red Fir (5M and 5D). The following vegetation types and strata provide moderate capability foraging habitat for goshawks: Aspen (3M and greater size and density classes); Eastside Pine (1, 2S, 3S, 3P, 3M, 3D, 4S, 4P, and 4M); Jeffrey Pine, Montane Hardwood, Montane Hardwood-Conifer, Montane Riparian, Sierran Mixed Conifer and White Fir (4P and below); Juniper and Pinyon Juniper (3S and greater); Lodgepole Pine and Subalpine Conifer (1, 2S, 3S, 3P, 3M, 3D, 4S, and 4P); and Red Fir (3M, 3D, 4S, 4P, 4M, 4D, 5S, and 5P).

Threats Habitat loss and degradation are the primary known threats to northern goshawks (Squires and Kennedy 2006). As a result of timber harvest and fire suppression policies over the past century, contemporary California forests are strikingly different in structure, composition, and function compared to the range of forest conditions resulting from the historic, natural disturbance regime. These recent management policies have likely degraded goshawk habitat quality by fragmenting forests, reducing the amount and distribution of mature and old-growth forest stands and large trees, increasing understory tree density, and changing tree species composition, resulting in broad-scale reduction of the proportion of pine in forest stands. Nest sites and territories have been lost from logging in nest stands and from stand replacement fires. There is also increased risk of loss of habitat to such fires because past management policies have increased fuel loads.

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Some of the threats facing northern goshawk include habitat loss and fragmentation (e.g., loss of large diameter trees), forest structure changes and changes in prey populations due to fire suppression and climate change, risk of habitat loss due to stand-replacing fires, and disturbance from human activity in and near territories. A study conducted by Morrison et al. (2011) in the Lake Tahoe Basin indicated that northern goshawks are susceptible to human disturbance; human activity was twice as high within infrequently occupied territories as compared to frequently occupied territories. Many kinds of human activities have been documented to affect raptors by altering habitats; physically harming or killing eggs, young, or adults; and by disrupting normal behavior (Postovit and Postovit 1987, Delany et al. 1999 as cited in Morrison et al. 2011). A recent study on nesting northern goshawk response to logging truck noise found that while goshawks alerted (turned their head in the direction of the noise) to the noise they did not flush and response was inversely proportional to the distance of the nest from the road (Grubb et al. 2012).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Late seral stage habitat is within each alternative and the combination of large trees as well as snags is provided for in each alternative. Climate change concerns are being considered and allowing for late seral habitat corridors and landscape linkages. Risks due to uncharacteristic wildfires is ameliorated by the restoration that is proposed.

Determination Statement Inyo, Sequoia and Sierra NFs: Implementation of the Forest Plan may affect northern goshawk but will not lead towards Federal listing or a loss of viability. Impacts to northern goshawk are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas while still maintaining large trees and snags.

References – Northern Goshawk Austin, K.K. 1993. Habitat use and home range size of breeding northern goshawks in the southern Cascades. M.S. Thesis, Oregon State Univ. Corvallis. 56pp.

Beier, P. and J.E. Drennan. 1997. Forest structure and prey abundance in foraging areas of northern goshawks. Ecological Applications 7(2):564-571.

Boal, C.W. and W. Mannan. 1994. Northern goshawk diets in ponderosa pine forests on the Kaibab Plateau. Studies in Avian Biology No. 16:97-102.

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Collins, B.M., J. D. Miller, A. E. Thode, M. Kelly, J. W. van Wagtendonk, S. L., Stephens. 2009. Interactions among wildland fires in a long-established Sierra Nevada natural fire area. Ecosystems. 12: 114-128.

Delaney, D.K., T.G.Grubb, P. Beier, L.L. Pater, and M. H. Reiser. 1999. Effects of helicopter noise on Mexican Spotted Owls. Journal of Wildlife management 63:60-76.

DeStefano, S. Daw, S.K. Desimone, S.M. and E.C. Meslow. 1994. Density and productivity of northern goshawks: implications for monitoring and management. Studies in Avian Biology No. 16:88-91.

Detrich, P.J. and B. Woodbridge. 1994. Territory fidelity, mate fidelity, and movements of color-marked northern goshawks in the southern Cascades of California. Studies in Avian Biology No. 16:130- 132.

Dewey, S.R., P. Kennedy, and R. Stephens. 2003. Are Dawn Volcaization Surveys Effective for Monitoring Goshawk Nest-Area Occupancy? The Journal of Wildlife Management. Vol. 67, No 2. Pp. 390-397.

Fretwell, S.D. and H.L. Lucas. 1970. On territorial behavior and other factors influencing habitat distribution in birds. Acta Biotheoretica, 19:16-36.

Graham, R.T. R.L. Rodriguez, K.M. Paulin, R.L. Player, A.P. Heap, and R. Williams. 1999. The northern goshawk in Utah: habitat assessment and management recommendations. USDA Forest Service, Rocky Mountain Research Station, General Technical Report RMRS-GTR-22.

Graham, R.T. Reynolds, R.T. Reiser, M.H. Bassett, R.L. and D.A. Boyce. 1994. Sustaining forest habitat for the northern goshawk: a question of scale. Studies in Avian Biology No. 16:12-17.

Grubb, T. G., A. E. Gatto, L. L. Pater, D. K. Delaney. 2012 (May 8). Response of nesting northern goshawks to logging truck noise, Kaibab National Forest, Arizona. Final Report to Southwest Region 3 (USDA Forest Service)

Hargis, C.D. C. McCarthy, and R.D. Perloff. 1994. Home ranges and habitats of northern goshawk in eastern California. Studies in Avian Biology, 16:66-74.

Keane, J. J. 1999. Ecology of the northern goshawk in the Sierra Nevada, California. PhD. Dissertation. University of California, Davis, CA. 123 pp.

Lenihan, J.M., Drapek,R., Bachelet, D., and R. P. Neilson. 2003. Climate change effects on vegetation distribution, carbon, and fire in California. Ecological Applications 13: 1667-1681.

Maurer, J.R., 2000. Nesting habitat and prey relations of the northern goshawk in Yosemite National Park California.

McGrath, M.T. DeStefano, S. Riggs, R.A. Irwin, L.L. and G.J. Roloff. 2003. Spatially explicit influences on northern goshawk nesting habitat in the Pacific Interior Northwest. Wildlife Monographs, No. 154.

Miller, J. D., H. D. Safford, M. Crimmins, and A. E. Thode. 2009. Quantitative evidence for increasing forest fire severity in the Sierra Nevada and southern Cascade Mountains, California and Nevada, USA. Ecosystems 12:16-32.

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Morrison, M. L., R. J. Young, J. S. Rosmos, R. Golightly. 2011. Restoring forest raptors: Influence of human disturbance and forest condition on northern goshawks. Restoration Ecology 19(2): 273- 279.

National Research Council. 2011. Climate stabilization targets: emissions, concentrations, and impacts over decades to millennia. The National Academies Press, Washington, DC, USA.

Postovit, H.R. and Postovit, B.C., 1987. Impacts and mitigation techniques.Raptor management techniques manual. National Wildlife Federation, Washington, DC Scientific Technical Series, 10, pp.183-213.

Reynolds, R.T. and Meslow, E.C., 1984. Partitioning of food and niche characteristics of coexisting Accipiter during breeding. The Auk, pp.761-779.

Reynolds, R.T. Graham, R.T. and D.A. Boyce. 2006. An ecosystem-based conservation strategy for the northern goshawk. Studies in Avian Biology, No. 31:299-311.

Reynolds, R.T. Joy, S.M. and D.G. Leslie. 1994. Nest productivity, fidelity, and spacing of northern goshawks in Arizona. Studies in Avian Biology No. 16:106-113.

Reynolds, R.T. R.T. Graham, M.H. Resier and others. 1992. Management recommendations for the northern goshawk in the Southwestern United States. Gen. Tech. Rep. RM-217. Fort Collins, CO. USDA, Forest Service, Rocky Mountain Forest and Range Exp. Station.

Shuford, W. D., and Gardali, T., editors. 2008. California Bird Species of Special Concern: A ranked assessment of species, subspecies, and distinct populations of birds of immediate conservation concern in California. Studies of Western Birds 1. Western Field Ornithologists, Camarillo, California, and California Department of Fish and Game, Sacramento.

Squires, J.R. and P. Kennedy. 2006. Northern goshawk ecology: an assessment of current knowledge and information needs for conservation and management. Studies in Avian Biology 31:8-62

Squires, J.R. and R.T. Reynolds. 1997. Northern goshawk (Accipiter gentilis), the birds of North America online (A. Poole, Ed.). Ithaca: Cornell Lab of Ornithology; Retrieved from the Birds of North America Online: http://bna.birds.cornell.edu/bna/species/298

Stephens, S. L., D. Fry, E. Franco-Vizcano. 2008. Wildfire and Spatial Patterns in Forests in Northwestern Mexico: The United States Wishes It Had Similar Fire Problems. Ecology and Society 13(2):10

Sugihara, N.G., J.W. van Wagtendonk, K.E. Shaffer, J. Fites-Kaufman, and A.E.Thode. 2006. Fire in California's ecosystems. University of California Press, Berkeley, CA.

USDA, Forest Service. January 2001. Sierra Nevada Forest Plan Amendment Final Environmental Impact Statement and Record of Decision. Pacific Southwest Region, San Francisco, California.

Westerling, A. L., H. Hidalgo, D. R. Cayan, and T. Swetnam. 2006. Warming and earlier spring increases western US forest wildlife activity. Science. 313:940-943.

Westerling, A.L. and B. P. Bryant. 2008. Climate change and wildfire in California. Climatic Change 87 (Suppl 1): S231-S249.

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Whitlock, C., Shafer, S.L., Marlon, J., 2003. The role of climate and vegetation change in shaping past and future fire regimes in the northwestern US and the implication for ecosystem management. Forest Ecology and Management 178: 5-21.

Woodbridge, B. and C.D. Hargis. 2006. Northern goshawk inventory and monitoring technical guide. Gen. Tech. Rep. WO-71. Washington, DC: U.S. Department of Agriculture, Forest Service. 80 pp.

Young, R.J. and M.L. Morrison. 2007. Assessment of goshawk territories within the Lake Tahoe basin; March 2004 – September 2005. Unpubl. report. 61 pp.

Zeiner, D.C. W.F. Laudenslayer Jr. K.E. Mayer, and M. White, (eds.). 1990. California’s Wildlife. Volume 2. Birds. California Statewide Wildlife Habitat Relationships System, California Department of Fish and Game, Sacramento, CA.

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Pallid Bat (Antrozous pallidus) Regional Foresters Sensitive Species

Species Account Pallid bat was originally described in 1856 as Vespertilio pallidus, but has had the genus name of Antrozous since 1862, and has most commonly been recognized as Antrozous pallidus (Barbour and Davis 1969, Hermanson and O’Shea 1983). There are currently two subspecies recognized in California (A. p. pacificus and A. p. pallidus) (Hall 1981, Simmons 2005). Recent work by Wilkinson et al. (2010) using amplified fragment length polymorphism in 187 individuals from 29 localities suggests that the California population (A. p. pacificus) is isolated and distinct from the more eastern (A. p. pallidus) and British Columbian populations, but that some gene flow has been occurring through male-mediated dispersal and gene flow since populations diverged during the Pliocene.

Pallid bat is considered a Mammal Species of Special Concern by California Department of Fish and Game, and a Sensitive Species by both Region Five of the U.S. Forest Service and BLM. The Western Bat Working Group designated it as a High Priority for conservation measures for most of its range in California.

The team developing this conservation plan ranked A. pallidus as one of the species of greatest concern). Along with Townsend’s big-eared bat, it received the highest cumulative score of 50 when scored by conservation issue. It also tied with Townsend’s big-eared bat in receiving the largest number of Number 5 scores for conservation issues (total of 7). When each team member selected the five species considered to be of greatest concern using California Department of Fish and Game criteria, pallid bat ranked third, receiving 13 out of a possible 14 votes.

Pallid bat is distributed throughout much of the west, from southern British Columbia to central Mexico, and as far east as western portions of Kansas, Oklahoma, and Texas, with an isolated subspecies in Cuba (Hermanson and O'Shea 1983; Simmons 2005).

Pallid bat is primarily a low to mid-elevation species, with an elevation record of 8,000 feet in the mountains of New Mexico (Black 1974). In California it is found from sea level up to approximately 7,400 feet (Baker et al. 2008, Pierson et al. 2001, 2009), although it is most commonly found below 5,900 feet (Barbour and Davis 1969, Orr 1954, Pierson et al. 2001 & 2009), and there is a record from below sea level in Death Valley (Orr 1954). It is found along the coast, in the coast ranges, the Central Valley, up to mid-elevation in the Sierra Nevada and Cascade ranges, and in the more xeric and desert habitats east of the Sierra Nevada and in southern California.

There is mounting evidence from various places in the state indicating range contractions for this species, particularly in areas of increasing urbanization. In the Museum of Vertebrate Zoology at the University of California, Berkeley and the Smithsonian Institution’s National Museum of Natural History, there are 51 museum records from 14 localities, the earliest being from 1883 and the most recent from the Hearst Mining building on the Berkeley campus in 1941. A colony of 38 was documented in Berkeley residence in 1919 (Storer 1931). E. Pierson and W. Rainey have been monitoring bat sightings in Berkeley since 1979, and have had no reports of pallid bats. Likewise, a colony of pallid bats known from Encina Hall at Stanford University in 1951 (Orr 1954), is known to have been gone for at least 30 years (Alan Launer, pers. comm.).

In Santa Clara and San Diego counties, where high rates of urbanization and land conversion have occurred, Johnston and Stokes (2007) reported that populations of pallid bats (A. p. pacificus) have continued to decline since being designated as a California species of special concern in 1986. Krutzsch

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(1948) identified 23 sites where A. pallidus occurred in the mid-1940s in San Diego Co; however, Johnston and Stokes (2007) reported that 12 of those sites where pallid bats previously occurred were considered extirpated by 2006 based on extensive surveys conducted between 1995 and 2004 in western San Diego County. Two new maternity colony roosts identified after 1948 have been eliminated, one during a building renovation and the other lost to a major fire (Johnston and Stokes 2007). In Santa Clara County, the only two colonies known prior to 1948 have since been extirpated (Johnston and Stokes 2007). Since 1992, 10 new maternity colony roosts have been located countywide; however, two maternity colonies have been destroyed, one colony was extirpated from a building during the maternity season, and five colonies are at risk because of the loss of foraging habitat (Johnston and Stokes 2007). As of 2006, Johnston and Stokes (2007) reported that only two out of the seven remaining colonies in Santa Clara County had not declined greatly in colony size and were not currently under threat.

Populations in the North and East San Francisco Bay Areas appear to be well distributed. However, many colonies are small, and are routinely excluded and evicted from anthropogenic roosts (Greg Tatarian, personal communication). Additionally, loss of oak woodland and savannah to housing development in many areas has resulted in permanent loss of known roosts (Greg Tatarian, unpublished data, CNDDB 2010).

In the NRIS database, the Inyo NF has 18 records, Sequoia NF has no records, and the Sierra NF has 2 records. Figure 33 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Pallid bat occurs in a number of habitats ranging from rocky arid deserts to grasslands into mid-elevation mixed deciduous/coniferous forests. In California, they are most commonly found in low elevation desert washes, western sycamore (Plantanus racemosa) open riparian habitat, coast live oak (Quercus agrifolia) and valley oak (Q. lobata) savannah, mid-elevation black oak (Quercus kelloggii) and mixed deciduous/coniferous forest (black oak, incense cedar (Libocedrus decurrens) and ponderosa pine (Pinus ponderosa) habitat (Barbour and Davis 1969, Central Coast Bat Research Group 2003, H. T. Harvey & Associates 2006, Johnston et al. 2006, Orr 1954, Pierson et al. 2001, Pierson et al. 2002, Rainey and Pierson 1996). It is also associated with both coast redwood and giant sequoia forests (Pierson and Heady 1996, Orr 1954, Rainey et al. 1992). Pallid bat also occurs in mixed agricultural areas (orchards, vineyards) where they are interspersed or adjacent to one or more native habitat types, such as in Napa, Marin, Sonoma and Mendocino Counties (Tatarian, unpublished data, CNDDB 2010). In the coastal plains and inland valleys of southern California, A. pallidus is associated with broad, flat riparian terraces dominated by oaks and sycamores (Stokes, unpublished data), and is also associated with oak woodlands (coast live oak (Quercus agrifolia) and Englemann oak (Quercus englemannii)) in the western foothills of southern California (Stokes, unpublished data). The species appears absent from the highest elevations of southern California. On the eastern slopes and deserts of southern California most commonly found in sagebrush habitats, desert washes, vegetated dune systems, and palm groves (Stokes, unpublished data).

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Figure 33. Map of pallid bat locations from the NRIS Wildlife Database, 2016

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Roosting Habitat Pallid bats are quite eclectic in their roosting habits (Barbour and Davis 1969, Hermanson and O'Shea 1983, Lewis 1994 & 1996, Orr 1954, Pierson et al. 1996). They roost in rock crevices (Orr 1954, Hermanson and O'Shea 1983, Pierson et al. 2002), under rock slabs (Vaughan and O’Shea 1976, Lewis 1996), in tree hollows (Orr 1954, Rainey and Pierson 1996, Rabe et. al. 1998, Pierson et al. 2004), caves, abandoned mines, and a variety of other anthropogenic structures, including buildings (vacant and occupied), porches and garages (van Zyll de Jong 1985), properly designed bat houses (Tatarian 2001a), and bridges (Barbour and Davis 1969, Beck and Rudd 1960, Johnston et al. 2004, Lewis 1996, Orr 1954, Pierson et al. 1996, Pierson et al. 2001, Pierson et al. 2002, Vaughan and O’Shea 1976). Tree roosting appears to be preferred in the forested regions of northern California, and has been documented in large conifer snags (e.g., incense cedar, ponderosa pine, sugar pine) (Baker et al. 2008, Johnston and Gworek 2006), inside basal hollows of redwoods (P.A. Heady pers. comm., Orr 1954, Rainey et al. 1992) and giant sequoias (Pierson and Heady 1996), and bole cavities in oaks and other trees (e.g. cottonwood, cypress) (Hall 1946, Orr 1954, Pierson et al. 2004, Rainey and Pierson 1996).

A radio-tracking study in the central coastal region of California documented winter roosting in the attic of an unheated building, with satellite roosts in trees (Quercus lobata, Q. agrifolia, Umbellularia californica, and Platanus racemosa) on or in the ground (under a large rock, under a dry mop in a shed, and under a concrete outhouse foundation) (Johnston et al. 2006). They have also been reported roosting in stacks of burlap sacks (Beck and Rudd 1960) and stone piles, particularly in the winter.

Though the pallid bat can be found roosting in a variety of situations, it appears to have a strong preference for anthropogenic structures in southern California; of the 14 pallid bat roosts documented in San Diego County over the years, only one was in a natural situation (a rock crevice on a near-vertical surface) (Stokes, ,unpublished data).

Pallid bats typically roost in maternity groups of 20 to 200 during summer (Hermanson and O'Shea 1983, Vaughan and O’Shea 1976) however, colonies of 300 individuals have been recorded using a bat house in Napa (Tatarian 2001a) and 327 bats were recorded in an historic building in Novato (Tatarian 2001b), but this species will also roost singly during pregnancy (Lewis 1996). In fall, maternity colonies disperse into smaller groups, which may be found in many sites where they do not occur in summer (Orr 1954, Barbour and Davis 1969).

Pallid bats in Oregon showed a higher fidelity towards night roosts than day roosts (Lewis 1994). Although Davis (1966) recorded pallid bats traveling 30 kilometers between night roosts, night roosts are more typically located within 1 to 2 kilometers of the day roost (Lewis 1994, Ball 2002, Johnston et al. 2006, Johnston and Gworek 2006, Baker et al. 2008). Roost switching by females is variable; in Arizona, pallid bats were reported to switch roosts in spring and autumn, but not during late pregnancy and lactation (O'Shea and Vaughan 1977), while in Oregon, females switch roosts throughout the summer, perhaps in an effort to benefit from lower ectoparasite loads (Lewis 1993). When using anthropogenic roosts in northern California, reproductive female A. pallidus show high roost fidelity (Tatarian, unpublished data), generally occupying maternity roosts in April or May, and moving to winter roosts in September, October, or even later if weather is moderate. When occupying physically large roosts, such as large building attics, females and their young will often utilize areas within the roost differentially, moving to warmer or cooler locations within and between both days and months in response to interior temperatures (Tatarian 2001b).

Compared to some other California bat species, pallid bats are relatively intolerant of disturbance (O'Shea and Vaughan 1977, Lewis 1996, Johnston et al. 2004) and may abandon a roost when disturbed. Anthropogenic day roosts are usually in more protected sites compared to night roosts (Tatarian 1999,

225 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests unpublished data); however, males and non-reproductive females of various age classes using well- established day or night roosts over several years can become habituated to human activity in and around their roost, and may day-roost on building exterior walls covered with open shed-roofs where they are exposed to humans moving beneath and visible to them (Tatarian 1999, unpublished data). Lewis (1996) noted that distances between day and nighttime roosts were usually less than 200 meters, but ranged from 40 to 1,850 meters, and during a radio-tracking study in 1997 to 1999 (Tatarian, unpublished data), a small, bridge night roost was located 4,986 meters from the large building day roost.

This is one of the species most likely to be found night-roosting under bridges (Barbour and Davis 1969, Johnston et al. 2004, Pierson et al. 1996, 2001), but it can also be found in shallow caves, cliff overhangs, and other man-made structures (Hermanson and O'Shea 1983, Lewis 1994). Lewis (1994) also noted that bridges utilized by pallid bats as night roosts were wooden, or concrete girder. Pallid bats show a higher fidelity towards night roosts than day roosts (Lewis 1994). Night roosts are typically located within 1 to 2 kilometers of the day roost.

Grossinger (2007) suggested that the restoration of the valley oak and valley oak savanna habitats, along with stewardship of adjacent grassland and riparian habitats, could have significant benefits for the Pacific pallid bat by providing reliable roosts and foraging areas.

Foraging Habitat Pallid bats forage close to the ground and vegetation in desert washes, open grassland, oak savannah, and/or forest with limited understory (e.g., ponderosa pine parkland or granite slabs with sparse vegetation) (Hermanson and O'Shea 1983). Pallid bats appear to regularly forage over leaf-litter deposited under oak canopies in southern California (Stokes, unpublished data). Johnston et al. (2006) found that male and female A. pallidus pacificus foraged intermittently through the winter months along and in riparian corridors with western sycamore (Plantanus racemosa) California bay (Umbellularia californica) and coast live oak (Quercus agrifolia) within canyon bottoms in central California; and during summer months, females and males foraged along ridges with grasslands, high open meadows and oak savannah habitats. Central Coast Bat Research Group (2001), Johnston and Gworek (2006), H. T. Harvey & Associates (2006), and Baker et al. (2008) determined that pallid bats frequently foraged on logging roads and in open and semi-open short grass meadows in the northern Sierra Nevada. Foraging appears to be concentrated in two periods: one just after emergence; and one prior to returning to the roost (Hermanson and O'Shea 1983). In southern California pallid bats also appear to prey regularly, or at least seasonally on white-lined sphinx moths and dragonflies; the latter presumably are gleaned from vegetation while resting at night (Stokes, unpublished data).

Lewis (1996) recorded distances of between 0.6 to 2.5 miles traveled between roost sites and foraging areas and Johnston et al. (2006) found similar distances of 0.2 to 4.0 kilometers for males and females during winter months. Tatarian (unpublished data) tracked both males and females foraging to areas between 1,640 to 4,986 meters from their roosts in the Napa-Sonoma region during April, May and October, but total distances travelled during foraging were often much greater, particularly by individuals that specialized in aerial insect prey. Johnston and Gworek (2006), found that radio-tagged bats in the northern Sierra Nevada foraged a mean distance of 1.1 miles from day roosts during summer months in the northern Sierra Nevada. Baker, et al. (2008) noted that the size of foraging areas for this species varied among sex and reproductive classes, with lactating females exhibiting the smallest foraging areas (1.56 square kilometers with a standard error of 0.88) and post-lactating females the largest foraging areas (5.97 square kilometers with a standard error of 2.69).

Pallid bat feeds primarily on medium to large, ground-dwelling prey, such as flightless arthropods (scorpions, Jerusalem crickets, cicadas, wolf spiders and centipedes), (Hatt 1923, Ross 1961, Easterla and

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Whitaker 1972, Hermanson and O'Shea 1983) and typically between 0.8 to 2.7 inches in length (Bell and Fenton 1984). Large cerambycid beetles, particularly Prionus californicus, and ten-lined June beetles (Polyphylla decemlineata) are also major prey items (Barbour and Davis 1969, Johnston and Fenton 2001, Orr 1954, Pierson et al. 2004) during the early part of summer. Johnston and Fenton (2001) found that a colony of A. p. pacificus had specialized individual dietary preferences within the same colony whereas individuals within a colony of pallid bats generally all ate the same prey items on any given night. Antrozous also gleans prey from vegetation (Hermanson and O'Shea 1983, and take prey in flight (Johnston and Fenton 2001). Bell (1982) stated that pallid bats used passive listening, and not echolocation, to detect and capture arthropods. However, pallid bats foraged primarily on a 0.4 inch scarab beetle in flight during mid-summer in Death Valley when the prey species was abundant (Johnston and Fenton 2001). In a radio-tracking and bioacoustic study conducted in Napa between 1997 to 1999 in April, May, and October (Tatarian unpublished data) 2 adult males, 1 subadult male, 1 juvenile male, 1 adult, non-reproductive female, and 1 subadult female, and observed that 3 male pallid bats foraged almost exclusively on flying insects, suggesting that some individuals in a colony specialize in pursuing aerial prey, using search and approach phase calls, as well as occasional feeding buzzes. Using carbon staple-isotope analysis, Herrara et al. (1993) documented that pallid bats obtained substantial amounts of carbon from plant sources during the blooming periods of bat-visited cacti and agaves, and facultative nectar-feeding by pallid bats on the flowers of cardon cacti (Frick et al. 2009).

Reproduction Pallid bats are gregarious, and often roost in colonies of between 20 and several hundred individuals. Males and females congregate in a central winter roost often associated with smaller satellite roosts in late fall and winter months (Johnston et al. 2006) when breeding occurs (Hermanson and O'Shea 1983). During spring months, pregnant females leave the winter roost and gather in summer maternity colonies (Johnston et al 2006), with parturition generally occurring between May and July depending on local climate (Barbour and Davis 1969). Males often leave the winter roost and use a variety of solitary roosts but sometimes form a bachelor colony (Johnston et al 2006). Females can give birth to a single pup, twins and sometimes triplets, with twins being most common (Barbour and Davis 1969). Young are generally weaned in mid to late August. Maternity colonies generally form in early April (Barbour and Davis 1969) and disband between August and October (Hermanson and O'Shea 1983, Lewis 1993).

Migration and Hibernation Pallid bats are relatively inactive during the winter; however, Johnston et al. (2006) found that males and females foraged intermittently throughout the winter months, in central California.

They are not known to migrate long distances (Barbour and Davis 1969), and Johnston et al. (2004) determined that the primary female/male winter roost of a large colony in central California was approximately 1 mile from the primary maternity colony roost. During January and February, pallid bats foraged about once every 6 nights, at temperatures down to 39 degrees Fahrenheit and on rainy nights, and winter prey at a central California coast site included darkling ground beetles (Carabidae), moths (Lepidotera) and other prey types often taken during warmer parts of the year (Johnston et al. 2006). Occasional winter activity has been reported in southern portions of its range and has been observed in Nevada flying during winter when temperatures were as low as 36 degrees Fahrenheit (O’Farrell et al. 1967, O’Farrell and Bradley 1970). Barbour and Davis (1969) reported hibernating or mildly torpid bats in buildings, a hollow post, limestone cliffs (Orr 1954), caves, mines (Hall 1946). Local microclimates may result in variability of fall and winter flight and roosting activity; several torpid individuals were monitored in a partially exposed day roost in the south Napa Valley between October and November, 2000, during periods of heavy rain (Tatarian unpublished data), moving to different locations within the roost, until leaving the roost entirely during a relatively warm, dry period in December.

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Threats

Agricultural Practices Vast areas of monoculture vineyards in Napa, Sonoma, Monterey and San Benito Counties, especially where oaks have been removed, have likely contributed to the decline of pallid bat populations in these areas. (Concerns with San Joaquin Valley practices, however, less or no concern in other parts of State, especially in desert situations). Habitat conversion to accommodate agriculture creates a loss of roosting (in the case of hardwood removal) and foraging habitat. Pallid bats were historically heavily associated with agricultural areas in southern California, particularly citrus groves. It is possible that initially these groves were attractive to pallid bats but over time become unproductive. There are large (and increasing) numbers of avocado groves in areas where pallid bats are known to currently exist in northern San Diego County. It is likely that this type of habitat conversion is unfavorable to the pallid bat.

Anthropogenic Roosts Due to their propensity for using a wide range of buildings as well as bridges, their highly visible roosting habits, urine stains and odor, as well as visible insect prey remains at night roosts, these bats are highly susceptible to negative human contact. Because pallid bats frequently roost in buildings and bridges display considerable roost loyalty in such roosts, and are often found roosting together with T. brasiliensis and M. yumanensis, two species that form large colonies (several hundreds to thousands), often where they are highly visible (e.g., open rafters) they are frequently subjected to vandalism, exclusion (humane or otherwise), even illegal poisoning. This species is often associated with historic buildings in which their presence is typically viewed as a hazard by property managers. Exclusion, renovation and demolition of buildings and urban expansion likely account for observed declines in Los Angeles, Orange, Santa Clara, and San Diego Counties. Particularly vulnerable are rural structures inhabited by pallid bat colonies that become subjected to renovation or demolition due to a change in land ownership or change in land-use practices. These changes are usually associated with the onset of urban development but can occur many years and miles ahead of such development.

Forest Management The removal of snags and damaged trees (particularly large ponderosa pines and incense cedars) and hardwoods during timber harvesting and the loss of hardwoods through conifer and brush competition (from a lack of fire management) have caused reductions for both roosting structures and foraging habitat. These practices may be severe on both private and public lands. Prescribed burning of leaf-litter likely results in a reduction or loss of foraging habitat.

Mines Pallid bats colonies can be impacted by inappropriate mine closures or disturbance during human visitation. Most pallid bat colonies in mines in southern California appear to be found in the desert.

Oak Woodlands The loss of hardwoods due to firewood cutting, urban expansion, conversion to agriculture, rangeland management, and disease (e.g. Sudden Oak Death Syndrome) has caused a serious reductions for both roosting and foraging habitat. Pallid bats are strongly associated with oaks throughout California. They can be found roosting in both dead and live oaks and are found foraging frequently under or at the edge of the oak canopy (Rainey and Pierson 1996, Johnston and Fenton 2001, Johnston et al 2006). Radio- tracking studies have identified pallid bats roosting in blue oaks in lower elevation oak savannah (E.D. Pierson unpublished data), and black oaks in mixed deciduous forest (Rainey and Pierson 1996). At Vandenberg Air Force Base they were radio-tracked foraging in coast live oak habitat (Pierson et al. 2002).

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Oak roosts (Rainey and Pierson 1996). Pallid bats were also radio-tracked to roosts in blue oak in Carmel Valley. Sudden oak death predisposes woodlands to fire.

Transportation Bridge retrofitting can render bridges unsuitable for both day and night roosting by this species), both during construction and after completion. Bridge replacement can result in complete loss of long-term day and night roost habitat, as many bridges being replaced are 40 to 60 years old. Bridges can support large populations of pallid bats, increasing impacts to this species when bridge roosts are lost. Pallid bats may not return to bridge roosts disturbed by construction activities, even when roost sites have not been modified (Johnston et al. 2004). Riparian habitat used for foraging where bridges occur is frequently partially cleared or temporarily disturbed to accommodate construction activities.

Urban Development In an acoustic monitoring study conducted in northern Monterey County, Whitford (2009) found that bat species richness was correlated inversely to road density, an index often used to measure the degree of urbanization. Species richness also increased as the ratio of native plant species to non-native plant species and as the diversity of land cover types increased. Species richness was lower in urban and agricultural land covers than rural land covers. The primary range contractions for Antrozous pallidus documented in Santa Clara and San Diego counties are attributed to the urbanization and conversion of natural habitats (Johnston and Stokes 2007). The loss of hardwoods due to urban expansion has caused serious reductions for both roosting and foraging habitat. House cats (Felis catus) have been observed to take bats (e.g., pallid bats) at roost sites (pers. comm. Monterey County Animal Control),

Pallid bats may be the species most vulnerable to loss of foraging habitat associated with urban development in southern California. While watercourses, wetlands, treed and riparian zones are somewhat maintained under current development practices, the peripheral low gradient terraces and mesas where pallid bats likely depend on for foraging are not preserved but rather become office lots, mini malls, apartment complexes, industrial lots, roads, parking lots, golf courses, and urban parks with asphalt and a monotypic grass species at best (even when the oaks are preserved, the understory is usually converted to a non-native well groomed grass where once would have been a lush layer of oak leaf-litter teeming with arthropods.

Caves Pallid bats colonies can be impacted by inappropriate cave closures or disturbance during human visitation.

Rangeland Management The presence of livestock can severely reduce hardwood recruitment and thus future roost trees. They can also greatly reduce ground cover (when not managed properly) which can lead to a reduction in prey species abundance. Many species of bats do benefit from properly designed water impoundments as a drinking source.

Solar Pallid bats are known to forage over flat creosote-scrub desert terrain. Because solar fields will result in large expanses of low-gradient desert scrub areas to be completely covered and therefore unavailable as foraging habitat, there should be some concern about this practice and its effects on the pallid bat. Additionally, most solar energy facilities require the use of water which will draw down water tables and make surface water in desert regions scarcer.

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Water Management There is a loss of foraging habitat due to inundation since roosts tend to be lower on slopes and close to water. Pallid bats are known to forage over flat creosote-scrub desert terrain. Because solar fields will result in large expanses of low-gradient desert scrub areas to be completely covered and therefore unavailable as foraging habitat, there should be some concern about this practice and its effects on the pallid bat.

Wind Farms Pallid bats are known to fly very high when commuting to and from roost sites to reach foraging habitat. An increase in wind facilities could cause a high degree of mortality.

San Gorgonio Pass, Tehachapi’s, southern California has been identified by computer models as the region most suitable for green energy projects due to the terrain and wind patterns.

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will provide for the habitats used by this species.

Determination Statement Inyo, Sequoia, and Sierra NFs: Implementation of the Forest Plan may impact pallid bat but will not lead towards Federal listing or a loss of viability. By reducing uncharacteristic wildfires, providing large trees and large snags, and providing protection to caves and mines, the management framework of the Forest Plans should be beneficial to pallid bats.

Literature Cited - Pallid Bat Baker, M.D., M.J. Lacki, and G.A. Falxa. 2008. Habitat use of pallid bats in coniferous forests of northern California. Northwest Science, 82: 269-275.

Ball, L. C. 2002. A strategy for describing and monitoring bat habitat. Journal of Wildlife Management, Vol. 66, No. 4 (Oct., 2002), pp. 1148-1153

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Barbour, R.W., and W.H. Davis. 1969. Bats of America. University of Kentucky Press, Lexington, KY, 286 pp.

Beck, A.J., and R.L. Rudd. 1960. Nursery colonies in the pallid bat. Journal of Mammalogy, 41:266-267.

Bell, G.P. 1982. Behavioral and ecological aspects of gleaning by desert insectivorous bat, Antrozous pallidus (Chiroptera: Vespertilionidae). Behavioral Ecology and Sociobiology, 10:217-223.

Bell, G.P. and Fenton, M.B., 1984. The use of Doppler-shifted echoes as a flutter detection and clutter rejection system: the echolocation and feeding behavior of Hipposideros ruber (Chiroptera: Hipposideridae). Behavioral Ecology and Sociobiology, 15(2), pp.109-114.

Black, H.L. 1974. A north temperate bat community: structure and prey populations. Journal of Mammalogy, 55(1):138-157.

Central Coast Bat Research Group. 2001. XXX Plumas report??

Davis, R. 1966. Homing performance and homing ability in bats. Ecological Monographs, 36:201-237.

Easterla, D. A., and J. O. Whitaker, Jr. 1972. Food habits of some bats from Big Bend National Park, Texas. Journal of Mammalogy, 53: 887-890.

Frick, W.F., P.A. Heady, and J.P. Hayes. 2009. Facultative nectar-feeding behavior in a gleaning insectivorous bat (Antrozous pallidus). Journal of Mammalogy, 90:1157-1164.

Grossinger, R. 2007. Historical Ecology Final Report Coyote Creek Watershed Historical Ecology Study. San Francisco Bay Estuary Institute. Oakland, Ca.

Hall, E.R. 1946. Mammals of Nevada. Univ. California Press, Berkeley, 710 pp.

Hatt, R.T. 1923. Food habits of the Pacific pallid bat. Journal of Mammalogy, General Notes, 4:260-261.

Hermanson, J.W., and T.J. O’Shea. 1983. Antrozous pallidus. American Society of Mammalogists, Mammalian Species, 213:1-8.

Herrera, L.G., T.H. Fleming, and J.S. Findley. 1993. Geographic variation in carbon composition of the pallid bat, Antrozous pallidus, and its dietary implications. Journal of Mammalogy, 74: 601-606.

H. T. Harvey & Associates. 2006.

Johnston, D.S., and M.B. Fenton. 2001. Individual and population-level variability in diets of pallid bats (Antrozous pallidus). Journal of Mammalogy, 82(2):362-373.

Johnston, D.S. and J.R. Gworek. 2006. Pallid bat (Antrozous pallidus) habitat use in a coniferous forest in northeastern California. Bat Research News, 47:114.

Johnston, D. S. and D. C. Stokes. 2007. [ABS]. Conservation strategies for the pallid bat (Antrozous pallidus). Western Section of the The Wildlife Society. Jan. 31 – February 2, 2007 Conference in Monterey, Calif.

Johnston, D.S., B. Hepburn, J. Krauel, T. Stewart, and D. Rambaldini. 2006. Winter ecology of pallid bats in central coastal California. Bat Research News, 47:115.

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Johnston, D.J., G. Tatarian, and E.D. Pierson. 2004. California bat mitigation: techniques, solutions, and effectiveness. Contract Report #2394-01 for California Department of Transportation, Sacramento, CA, 125 pp.

Krutzsch, P. H. 1948. Ecological study of the bats of San Diego County, California. Master’s Thesis. University of California, Berkeley, 184 pp.

Lewis, S.E. 1993. Effect of climatic variation on reproduction by pallid bats (Antrozous pallidus). Canadian Journal of Zoology 71:1429-1433.

Lewis, S.E. 1994. Night roosting ecology of pallid bats (Antrozous pallidus) in Oregon. American Midland Naturalist, 132(2):219-226.

Lewis, S.E. 1996. Low roost-site fidelity in pallid bats: associated factors and effect on group stability. Behavioral Ecology and Sociobiology, 39(5):335-344.

Orr, R.T. 1954. Natural history of the pallid bat, Antrozous pallidus (LeConte). Proceedings of the California Academy of Sciences, 28:165-246.

O’Farrell, M.J., W.G. Bradley, and G.W. Jones. 1967. Fall and winter bat activity at a desert spring in southern Nevada. Southwestern Naturalist. 12:163-171

O’Farrell. M.J., and W.G. Bradley. 1970. Activity patterns of bats over a desert spring. Journal of Mammalogy, 51(1):18-26.

O'Shea, T. J., and T. A. Vaughan. 1977. Nocturnal and seasonal activities of the pallid bat, Antrozous pallidus. Journal of Mammalogy 58: 269-284.

Pierson, E.D., and P.A. Heady. 1996. Bat surveys of Giant Forest Village and vicinity, Sequoia National Park. Report for National Park Service, Denver Service Center, Denver, CO. 27 pp.

Pierson, E.D., W.E. Rainey, and R.M. Miller. 1996. Night roost sampling: a window on the forest bat community in northern California. Pp. 151-163 in Bats and Forests Symposium, October 19-21, 1995, Victoria, British Columbia, Canada. Research Branch, Ministry of Forests, Victoria, British Columbia, Working Paper 23/1996.

Pierson, E.D., W.E. Rainey, and C. Corben. 2001. Seasonal patterns of bat distribution along an altitudinal gradient in the Sierra Nevada. Report to California State University at Sacramento Foundation, Yosemite Association, and Yosemite Fund, 70 pp.

Pierson, E.D., P.W. Collins, W.E. Rainey, P.A. Heady, and C.J. Corben. 2002. Distribution, status and habitat associations of bat species on Vandenberg Air Force Base, Santa Barbara County, California. Santa Barbara Museum of Natural History Technical Reports No. 1:1-135.

Pierson, E.D., W.E. Rainey, P.A. Heady and W.F. Frick. 2004. Bat surveys for State Route 104 Bridge over Dry Creek, Amador County: replacement project. Contract Report for California Department of Transportation, Stockton, CA. 53 pp.

Pierson, E.D., and A. Chung-MacCoubrey. 2009. A natural resource condition assessment for Sequoia and Kings Canyon National Parks, Appendix 16 – bats. Natural Resource Report NPS/SEKI/NRR – 2013/665.16.

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Rabe, M.J., T.E. Morrell, H. Green, J.C. deVos, Jr., and C.R. Miller. Characteristics of ponderosa pine snag roosts used by reproductive bats in northern Arizona. Journal of Wildlife Management, 62(2):612-621.

Rainey, W.E., and E.D. Pierson. 1996. Cantara spill effects on bat populations of the upper Sacramento River, 1991-1995. Report to California Department of Fish and Game, Redding, CA, (Contract # FG2099R1). 98 pp.

Rainey, W.E., E.D. Pierson, M. Colberg, and J.H. Barclay. 1992. Bats in hollow redwoods: seasonal use and role in nutrient transfer into old growth communities. Bat Research News, 33(4):71.

Ross, A. 1961. Notes on food habits of bats. Journal of Mammalogy 42(1): 66-71.

Simmons, N.B. 2005. Chiroptera. Pp. 312-529, in Mammal Species of the World: a taxonomic and geographic reference. D.E. Wilson and D.M. Reeder, Editors. Volume I. Johns Hopkins University Press, Baltimore, 743 pp.

Storer, T. I. 1931. A colony of Pacific pallid bats. Journal of Mammalogy 12: 244-247.

Tatarian, G. 1999. Use of buildings and tolerance of disturbance by pallid bats Antrozous pallidus. Bat Research News 40:11-12.

Tatarian, G. 2001a. Successful pallid bat house design in California. Bat House Researcher, 9(2):2-4.

Tatarian, G. 2001b. Bat Habitat Assessment. Olompali State Historic Park. Novato, California. Frame House Restoration. Prepared for: State of California Department of Parks and Recreation. 41 pp. van Zyll de Jong, C. G. 1985. Handbook of Canadian mammals, volume 2: bats. National Museum of Natural Sciences, National Museums of Canada, Ottawa, Ontario, Canada.

Vaughan, T.A., and T.J. O’Shea. 1976. Roosting ecology of the pallid bat, Antrozous pallidus. Journal of Mammalogy, 57(1):19-42.

Whitford, S.K. 2009. Patterns of Bat Species Richness and Activity Levles: A Monterey County Case Study, San Diego state University.

Wilkinson, J. E., J. B. Lack, and R. A Van Den Bussche. 2010. [ABS]. Nuclear DNA phylogeography of the pallid bat (Antrozous pallidus). American Society of Mammalogists 90th Annual Meeting. Lamramie, Wyoming. June 11-15, 2010.

Personal Communications:

Alan Launer. Director, Center for Conservation Biology. Stanford University. Conversation between Dave Johnston and Alan Launer February 2006.

Greg Tatarian

Paul Heady

Monterey County Animal Control

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Pygmy Rabbit (Brachylagus idahoensis) USDA Forest Service: Sensitive Species

Species Account Pygmy rabbit are the only species in Brachylagus and are restricted to the Great Basin of the western United States. They are known to exist in isolated populations of northeastern California, southern Idaho, southwestern Montana, northern Nevada, eastern Oregon, western Utah, western Wyoming, and southeastern Washington. The population in Washington State is confined to the Columbia Basin and thought to have been geographically isolated from other populations of the species for thousands of years (Duszynski 2005). The pygmy rabbit’s current geographic range, excluding the Columbia Basin population, includes most of the Great Basin and some of the adjacent intermountain areas of the western United States (Green and Flinders 1980, USDI 2005). The northern boundary extends into southeastern Oregon and southern Idaho. The eastern boundary extends into southwestern Montana and southwestern Wyoming. The southeastern boundary extends into southwestern Utah. Central Nevada and eastern California provide the southern and western boundaries (USDI 2005).

The historic range (Figure 34) of the pygmy rabbit encompassed 100 million acres or more of sagebrush habitat in the Great Basin and Intermountain West. Currently populations exist in portions of 7 to 8 million acres (petition for federal listing, USDI 2005), of their historic 100 million. The elevational range of pygmy rabbits’ current distribution is narrow. In Nevada they are found from 4,500 to about 7,000 feet and in California a much narrower range of, 5,000 to 5,300 feet (Tesky 1994).

Literature indicates that pygmy rabbits were never evenly distributed across their range (USDI 2005). In California pygmy rabbit has been noted within the Bodie area of Mono County, and in Modoc and Lassen Counties (190 miles to the northwest) (Jones 1957). Pygmy rabbit has also been documented in the Crowley Lake area of Inyo County (Jones 1957). The California Natural Diversity Database (CNDDB) lists occurrences of pygmy rabbit on the Inyo National Forest between highway 167 and Mono Lake. Occurrences are also noted within a mile of the Inyo NF on both sides of highway 395 in Pumice Valley southwest of Mono Lake (Figure 35).No documented sightings are known to occur on any other Forest within the Region (CNDDB 2012, NRIS 2012).

Status Populations have been declining in Washington, Oregon, and California, where sagebrush habitat has been burned, converted to agriculture, or cleared from large areas and replaced with bunch grasses to improve livestock forage (USDI 2005).

The Washington Department of Fish and Wildlife in 2001 determined that the Columbia Basin population of pygmy rabbits (Douglas County, Washington) is genetically distinct and has been isolated from the Idaho and Oregon populations for at least 7,000 years. The population has declined swiftly from approximately 150 pygmy rabbits in 1995 to less than 30 in 2001. The Columbia Basin population of pygmy rabbits (Brachylagus idahoensis) was listed under the Endangered Species Act in March 2003 (Duszynski 2005). A 90 day petition for federal listing (Federal Register Vol 70 No. 96 (USDI 2005) for the remaining populations of pygmy rabbit was submitted, and was ultimately not federally listed.

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Figure 34. Pygmy rabbit distribution map

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Figure 35. Pygmy rabbit occurrences from CNDDB on or adjacent to the Inyo National Forest

Life History Pygmy rabbits display several traits that set them apart from species of cottontails (Sylvilagus), jackrabbits (Lepus), and Old World rabbits (Oryctolagus). They are the only leporids in the continental United States that dig their own extensive and interconnected burrows (Green and Flinders 1980). Burrows typically have 4 or 5 entrances, but may have as few as 2 or as many as 10. Tunnels widen below the surface, forming chambers, and extend to a maximum depth of about 3.3 feet. When alarmed they give alarm calls and other vocalizations (Green and Flinders 1980). They move by a low, scampering gait (not leaping) (Green and Flinders 1980). They are almost totally dependent on sagebrush (Artemisia tridentata) for their diet, especially during winter months, when it may comprise 98 to 99 percent of their food intake (Duszynski 2005, Wilde 1978; Green and Flinders 1980).

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The 2005 petition for federal listing (USDI 2005) was denied in 2010, but noted a wide range of pygmy rabbit population densities across their range:

Janson (1946) reported an estimated pygmy rabbit density of 0.75 to 1.75 per acre in Utah. In another area in Utah, he estimated 3.5 pygmy rabbits per acre. Green (1978) reported an estimate of 18.2 pygmy rabbits per acre in Idaho. Gahr (1993) estimated that 0.09 pygmy rabbits per acre in a grazed area and 0.11 per acre in an ungrazed area in Sagebrush Flat, Washington. In Montana, Rauscher (1997) estimated pygmy rabbit density as 1.2 per acre.

The wide range of population densities is likely tied to a number of factors. The mortality rates of adult and juvenile pygmy rabbits also vary considerably between years, and even between juvenile cohorts within years (Wilde 1978). Juvenile mortality was highest from birth to 5 weeks of age. The annual mortality rate of adult pygmy rabbits may be as high as 88 percent, and more than 50 percent of juveniles can die within roughly 5 weeks of their emergence (Green and Flinders 1980, Wilde 1978). The mortality of adults is highest in late winter and early spring. However, the mortality rates of adult and juvenile pygmy rabbits can vary considerably between years, and even between juvenile cohorts within years (USDI 2005). Population cycles are not known in pygmy rabbits, although local, relatively rapid population declines have been noted in some States (USDI 2005).

Predation is the main natural cause of pygmy rabbit mortality and include badgers (Taxidea taxus), long- tailed weasels (Mustela frenata), coyotes (Canis latrans), bobcats (Felis rufus), great horned owls (Bubo virginianus), longeared owls (Asio otus), ferruginous hawks (Buteo regalis), northern harriers (Circus cyaneus), and common ravens (Corvus corax) (Green and Flinders 1980, Wilde 1978).

Pygmy rabbits are capable of breeding when they are about 1 year old. The breeding season of pygmy rabbits lasts from March through May in Idaho; in Utah, from February through March (Tesky 1994). The gestation period of pygmy rabbits is unknown. It is between 27 and 30 days in various species of cottontails (Sylvilagus spp.). Averages of six young are born per litter and a maximum of three litters are produced per year (Tesky 1994). In Idaho the third litter is generally produced in June (Tesky 1994). It is unlikely that litters are produced in the fall. The growth rates of juveniles are dependent on the date of birth. Young from early litters grow larger due to a longer developmental period prior to their first winter.

Habitat Status Pygmy rabbits typically occur in areas of tall, dense sagebrush (Artemisia spp.) cover, and are highly dependent on sagebrush to provide both food and shelter throughout the year (Katzner 1997, USDI 2005). Pygmy rabbits inhabit dense vegetation along perennial and intermittent stream corridors, alluvial fans, and sagebrush plains probably provided travel corridors and dispersal habitat for pygmy rabbits between habitat areas (USDI 2005). Individual sagebrush plants in areas inhabited by pygmy rabbits are often 6 feet or more in height. Extensive, well-used runways interlace the sage thickets and provide travel and escape routes. Dense stands of big sagebrush along streams, roads, and fencerows provide dispersal corridors for pygmy rabbits (Tesky 1994).

The last remaining population of Columbia Basin pygmy rabbits residing at Sagebrush Flat in central Washington selected ungrazed areas, especially a unit that was ungrazed by livestock since 1957, preferentially over grazed areas when constructing burrows (Thines 2004). Pygmy rabbits are seldom found in areas of sparse vegetative cover and seem to be reluctant to cross open space. In southeastern Idaho, woody cover and shrub heights were significantly greater on sites occupied by pygmy rabbits than on random sites in the same area. Some researchers have found that pygmy rabbits never venture further than 60 feet from their burrows (Tesky 1994). However one observer noted pygmy rabbits range up to 328 feet from their burrows (Tesky 1994). Pygmy rabbits may be active at any time of day. However, they

237 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests are generally most active at dusk and dawn (Tesky 1994). They usually rest near or inside their burrows during midday (Tesky 1994) and may be vulnerable to predators.

Since European settlement of the western United States, dense vegetation associated with human activities (e.g., fence rows, roadway shoulders, crop margins, abandoned fields) may have also acted as avenues of dispersal between local populations of pygmy rabbits (Green and Flinders 1980, USDI 2005).

The winter diet of pygmy rabbits is comprised of up to 99 percent sagebrush (Duszynski 2005, Wilde 1978, Green and Flinders 1980). Some areas inhabited by pygmy rabbits are covered with several feet of snow for up to 2 or more months during the winter. During periods when the snow has covered most of the sagebrush, pygmy rabbits tunnel beneath the snow to find food. Snow tunnels are approximately the same height and width as underground burrows. They are quite extensive and extend from one sagebrush plant to another. Aboveground movement during the winter months is restricted to these tunnel systems (Tesky 1994).

During spring and summer in Idaho, their diet consists of roughly 51 percent sagebrush, 39 percent grasses (particularly native bunch-grasses, such as Agropyron spp. and Poa spp.), and 10 percent forbs (Green and Flinders 1980). In addition, total grass cover relative to forbs and shrubs may be reduced within the immediate areas occupied by pygmy rabbits as a result of its use during spring and summer (Green and Flinders 1980).

Grasses and forbs are also eaten from mid- to late summer. In Idaho, shrubs contributed 85.2 percent (unweighted mean) of pygmy rabbit diets from July to December. Shrub use was lowest in August (73.1 percent) and highest in December (97.9 percent). Big sagebrush was the most important shrub in the July to December diet (54.2 percent), followed by rubber rabbitbrush (Chrysothamnus nauseosus) (25.8 percent) and winterfat (Krascheninnikovia lananta) (4.6 percent). Grasses comprised 10 percent of the July to December diet and were consumed mostly during July and August. Indian ricegrass (Oryzopsis hymenoides) and needlegrass (Stipa spp.) were the most important grasses consumed. Forbs contributed 4.9 percent of the July to December diet (Tesky 1994).

In southeastern Idaho, Green and Flinders (1980) found that pygmy rabbits ate big sagebrush throughout the year but in lesser amounts in summer (51 percent of diet) than in winter (99 percent of diet). Other shrubs in the area were consumed infrequently. Grass and forb consumption was relatively constant throughout the summer (39 percent and 10 percent of diet respectively) and decreased to a trace amount through fall and winter. Thickspike wheatgrass, bluebunch wheatgrass (Pseudoroegneria spicata), and Sandberg bluegrass were preferred foods in the summer.

In the Upper Sonoran Desert pygmy rabbits occur in desert sagebrush associations dominated by big sagebrush and rabbitbrush with bitterbrush and sulphur-flowered buckwheat (Eriogonum umbellatum var. stellatum) (Tesky 1994).

In the Lemhi Valley in east-central Idaho pygmy rabbis are found in relatively continuous sagebrush habitat, although much of the area is characterized by mounded microtopography (mima mounds) with short, sparsely distributed sagebrush between mounds and dense, tall sagebrush on mounds. Burrow systems are primarily on the mounds with deeper soil (pers. comm. Estes-Zumpf 2007).

Juveniles use burrows more than other age groups. Early reproductive activities of adults may be concentrated at burrows. When pygmy rabbits can utilize sagebrush cover, burrow use is decreased. Pygmy rabbits use burrows more in the winter for thermal cover than at other times of the year. Burrows are usually located on slopes at the base of sagebrush plants, and face north, north-east, or east.

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In Oregon, pygmy rabbits inhabited areas where soils were significantly deeper and looser than soils at adjacent sites. Site selection was probably related to ease of excavation of burrows. In areas where soil is shallow pygmy rabbits live in holes among volcanic rocks, in stone walls, around abandoned buildings, and in burrows made by badgers (Taxidea taxus) and marmots (Marmota flaviventris) (Tesky 1994).

In the Mono Lake area, the rabbits are found in soils that have a higher sand content than populations found in Nevada. They are often found in "loamier" inclusions in stabilized sand dunes. Also, the rabbits in the Bodie area live at very high elevations of at least 8,400 feet. These loamy soils support relatively low sagebrush of about 2 to 3 feet tall (Beauvais et al. 2012).

Potential suitable burrow habitats may be modeled (Himes and Drohan 2006). Based on the three habitat variables that enabled the distinction between most survey sites where pygmy rabbits were present and absent, the rabbits in this study were found in areas that are closer to perennial streams, have greater soil depths, and have more northerly aspects, relative to the availability of other areas. Over time, soils conducive to rabbit burrowing tend to accumulate nearer perennial streams than in surrounding areas due to alluvial deposition following flood events.

Threats Populations have been declining in Washington, Oregon, and California, where sagebrush habitat has been burned, converted to agriculture, or cleared from large areas and replaced with bunch grasses to improve livestock forage (Duszynski, 2005). Pygmy rabbit is dependent on big sagebrush (Artemisia tridentata) and is particularly sensitive to habitat loss. Fires that are more frequent due to the increase of invasive cheatgrass, conversion to cropland and rangeland are contributors to loss of sagebrush habitat. Population isolation due to fragmentation of suitable habitat can cause bottlenecks that threaten subpopulations (Beauvais et al. 2012). The main threat to pygmy rabbits in Nevada and California is fragmentation resulting from vast fires. These fires have extirpated populations from entire valleys and without leaving any cover behind, may not allow rabbits to repopulate the area. This is a severe problem that will most likely become greater with increasing cheatgrass invasion. Agricultural fields are also a threat. Not only are they usually placed on the same loamy soils the rabbits prefer, but they also create a potentially impassable barrier and may increase the habitat suitability for cottontails. Road systems may also pose a barrier to dispersal (Beauvais et al. 2012).

Additional changes to sagebrush habitats include: large-scale irrigation project (including canals), effects from livestock grazing (overstocking, timing of use, trampling of burrows, loss of forage and cover), weed introduction and loss of sagebrush plant community especially from cheatgrass (from livestock, recreation, vehicles, and other sources – and compounded by burning), mining, energy development, military developments, recreation (primarily from off-highway vehicles), hunting, habitat management (such as for sage grouse), non-native diseases (such as West Nile virus), small and disjunct populations that may be vulnerable to impacts, civil infrastructure, and urbanization in sagebrush habitats (USDI 2005). Pygmy rabbit diet restriction to largely sagebrush most of the year means a decimation of the food supply in one year may result in decimation of pygmy rabbits in that area (Wilde 1978) from overuse.

After initial declines, pygmy rabbit populations may not have the same capacity for rapid increases in numbers in response to favorable environmental conditions as compared to other rabbit species. This may be due to their close association with specific components of sagebrush ecosystems and the relatively limited availability of their preferred habitats (Wilde 1978, Green and Flinders 1980, USDI 2005).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and the Inyo National Forest Plan, as amended for sagebrush and sage grouse protection, and it allows for the continued protection of sagebrush areas.

Alternative B, C, and D: Restoration and or maintenance of sagebrush will occur as has a management framework for the Inyo National Forest as shown in Chapter 2 and Chapter 3. Removal of invasive species will reduce the impacts and or competition from other species.

The population on the Inyo National Forest are not compounded my multiple factors such as urbanization that is occurring elsewhere, hence the population is not thought to have decreased as compared to other areas. No specific population trend is known for the Inyo National Forest. Due to the lack of population trends, no Forest specific threats to the pygmy rabbit, and the widespread location of the rabbit, the species was not proposed as a Species of Conservation Concern.

Determination Statement Inyo NF: Implementation of the Forest Plan may impact pygmy rabbit but will not lead towards Federal listing or a loss of viability. Impacts to pygmy rabbit are beneficial impacts such as restoration of sagebrush habitat and removal of invasive species.

Sequoia and Sierra NF: Implementation of the Forest plan will have no impact to pygmy rabbit since they are not known to occur on these forest.

Literature Cited - Pygmy Rabbit Beauvais, G.P., Sequin, E., Rachlow, J., Dixon, R., Bosworth, B., Kozlowski, A., Carey, C., Bartels, P., Obradovitch, M., Forbes, T. & Hays, D. 2008. Brachylagus idahoensis. In: IUCN 2012. IUCN Red List of Threatened Species. Version 2012.2.

Duszynski, Donald W. et. al . 2005. A pathogenic new species of Eimeria from the Pygmy Rabbit Brachylagus idahoensis, in Washington and Oregon, with description of the sporulated oocyst and intestinal endogenous stages. University of New Mexico, Albuquerque, NM

Estes-Zumpf, Wendy. Personal communication about pygmy rabbits to Ginger Bolen from North State Resources (while under contract to Forest Service Pacific Southwest Region).

Green, Jeffrey S., and Jerran T. Flinders. 1980. Mammalian Species, No. 125, Brachylagus idahoensis. (Apr. 15, 1980), pp. 1-4.

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Himes, J.G., and Drohan, P. J. 2006. Distribution and habitat selection of the pygmy rabbit, Brachylagus idahoensis, in Nevada (USA). Journal of Arid Environments, September, 2006.

Jones, 1957. Southern Extention of the Range of the Pygmy Rabbit in California. Journal of Mammalogy, Vol. 38, No. 2, May 1957.

Katzner, Todd E., Parker, Katherine L. 1978. Vegetative Characteristics and Size of Home Ranges Used (Brachylagus idahoensis) during Winter by Pygmy Rabbits. Journal of Mammalogy, Vol. 78, No. 4. (Nov., 1997), pp. 1063-1072.

Montana field guide 2012. http://fieldguide.mt.gov/detail_AMAEB04010.aspx

Tesky, Julie L. 1994. Brachylagus idahoensis. In: Fire Effects Information System, [Online]. U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station, Fire Sciences Laboratory.

Thines, Nicole J. Siegel. Shipley, Lisa A., Sayler, Rodney D. 2004. Effects of cattle grazing on ecology and habitat of Columbia Basin pygmy rabbits (Brachylagus idahoensis). Department of Natural Resource Sciences, Washington State University, Pullman, WA. In Biological Conservation, Vol. 119, January 2004.

USDI 2005. Federal Register. Endangered and Threatened Wildlife and Plants; 90-Day Finding on a Petition To List the Pygmy Rabbit as Threatened or Endangered. Vol. 70, No. 97/Friday, May 230, 2005.

Wilde, Douglas Brian 1978, A population analysis of the pygmy rabbit (Sylvilagus idahoensis) on the INEL site. Idaho State University, PhD., 1978

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Wolverine (Gulo gulo) Regional Forester’s Sensitive Species

Species Account Wolverines are distributed across the circumpolar northern hemisphere (Banci 1994; USFWS 2011) and their range extends south along connected mountain ranges (Pasitschniak-Arts and Larivière 1995). Historically in North America, they were found from Alaska to eastern Canada (Banci 1994; Pasitschniak- Arts and Larivière 1995) and south across the Canada-US border provinces and states, and extended south along the Rocky Mountains to Arizona and New Mexico and along the west coast mountains and Sierra Nevada Mountains (Banci 1994, Pasitschniak-Arts and Larivière 1995).

The southern extent of the wolverine’s range is discontinuous and is found in mountainous terrain. In the contiguous United States, wolverines are found in, Montana, Wyoming, Idaho, Oregon, Washington, and, recently, California (Moriarty et al. 2009, USFWS 2011) and Colorado (USFWS 2010). On the west coast, wolverine population centers probably existed in two centers – the Northern Cascades and the Sierra Nevada Mountains (USFWS 2010).

In Region 5, the wolverine distribution is much reduced from the past. Historical Forest Service and state records indicate it was found from the southern Sierra Nevada north to the Warner Mountains and west to the Klamath and Coast Mountains. Observations or specimens from outside the Sierra Nevada population center were probably transients or dispersing males. Verifiable records were practically absent since the early twentieth century (Moriarty et al. 2009) until a male wolverine was photographed at baited camera stations on the Tahoe NF and adjacent Sierra Pacific Industries land in 2008 through 2010 (Moriarty et al. 2009, USFWS 2010, USFS NRIS records database 2012). These records are north of Interstate 80 in Nevada and Sierra counties, west and south of Sierraville California (see Figure 36). The USFWS considers the Sierra Nevada Mountains to be part of the wolverine’s current range, but a population has not been reestablished (the single male identified in 2008 does not make a population) (USFWS 2010).

The historic range of the wolverine in California included much of the Sierra Nevada ecoprovince (Banci 1994). Wolverines were believed near extinction in the early 1920s and it was concluded that the species was still rare and declining. By 1973 it was believed that wolverines were becoming established in the mountainous areas of north-western California, from "surviving nuclei" to the north (Ibid.). The current range includes a broad arc from Del Norte and Trinity counties through Siskiyou and Shasta counties, and south through the Sierra Nevada to Tulare County (Ibid.). Another researcher expanded this range to include the White Mountains (Ibid.).

Wolverines in the southern part of the Pacific Northwest Coast and Mountains ecoprovince are becoming isolated from the northern portion of the ecoprovince by heavy development in British Columbia. However, occasional reports within the Thomson-Okanogan Highlands ecoprovince of British Columbia and Washington suggest that this may be a dispersal corridor. It is also possible that wolverines have become isolated within the Sierra Nevada ecoprovince of California because of human activities. (Banci 1994)

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Figure 36. Historic and current range of the wolverine in the National Forests of California

There are probably 250 to 300 wolverines in the contiguous United States and these are distributed among five subpopulations: northern Cascades (Washington), Idaho, the Greater Yellowstone Ecosystem, northern Montana, and the Crazy and Belt population (west-central Montana) (USFWS 2011). Furthermore, the distance between these subpopulations is far enough such that dispersal, and therefore genetic exchange, is infrequent (USFWS 2011). The USFWS (2011) indicates at least 400 breeding pairs would be necessary to sustain genetic variability over the long term in the contiguous United States, absent immigration. Genetic structure in the southern and eastern extremes of the distribution of wolverines is relatively high, possibly reflecting the fragmented nature of these populations at the periphery of their historical range (Kyle 2002)

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There is one known wolverine in Region 5. The two source population centers (the North Cascades and western Rocky Mountains) are the nearest to the area, and the Idaho population is the closest. However, the Idaho population is still at least 400 miles away, which makes for a long dispersal distance. Hair and scat samples were acquired and it was genetically determined that this animal most likely from the western Rocky Mountains, and perhaps more accurately, central Idaho (Moriarty et al. 2009).

In the contiguous US, wolverine populations appear to have increased during the last 50 years, as indicated by recolonization of areas from which they were previously extirpated (Aubry et al. 2007). Recently, dispersing males have been documented in California (Moriarty et al. 2009) and Colorado (Inman et al. 2010). These are the first verifiable wolverine records in these states since 1924 and 1919 ,respectively (Aubry et al. 2007), and are likely associated with current population expansions such as the recent observed recolonization of the northern Cascade Range in Washington (K. Aubry, unpublished data).

In the NRIS database, the Inyo NF has 1 records, Sequoia NF has 1 records, and the Sierra NF has 14 records. Figure 37 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot. Do note that these are considered historical locations.

Life History In the contiguous United States, wolverines are found in boreal habitats associated with mountains (Copeland et al. 2010; USFWS 2011). They are usually solitary and found in very low densities. A significant portion of their diet comes from scavenging carcasses. Male home ranges are usually larger than those of females and often overlap that of more than one female.

Females do not breed their first summer (Rausch and Pearson 1972, Banci and Harestad 1988, Banci 1994) and authors have reported varying proportions of the subadult age class (1 to 2 years) that breed. Banci and Harestad (1988) reported 7 percent in the Yukon, contrasting with the 50 percent reported by Rausch and Pearson (1972) in Alaska and the Yukon, and 85 percent reported by other researchers for British Columbia (Banci 1994). Differences in how wolverine ages were classed make comparisons among studies difficult; the subadult age class in the latter two studies may have included adults. Most males are sexually immature until at least 2 years of age (Rausch and Pearson 1972, Banci and Harestad 1988). Testis weights increase throughout the winter (Rausch and Pearson 1972, Banci and Harestad 1988, Banci 1994) and by March, all adult males are in breeding condition (Banci 1994). Rausch and Pearson (1972) reported a peak in testis weights in June, presumably indicating the peak in breeding activity.

Males reach sexual maturity at two and a half years; females may breed at age one (Banci 1994). Mating occurs during the summer (May through August), followed by delayed implantation (Pasitschniak-Arts and Larivière 1995). Because implantation can occur from November to as late as March, kits can be born as early as January or as late as April (Banci 1994), but are usually born in February or March (Pasitschniak-Arts and Larivière 1995). One to four kits make up the typical litter (Banci 1994, Pasitschniak-Arts and Larivière 1995). Kits develop rapidly, are weaned in seven to ten weeks, and are full grown in about seven months (Banci 1994, Pasitschniak-Arts and Larivière 1995).

Dispersal by young is different for each sex. Females typically stay near where they were born, whereas males will not. The longest recorded dispersal distance by a male was 378 kilometers (about 230 miles) from Alaska to Yukon (Banci 1994). If the wolverine from California was a dispersing male from Idaho, it traveled about 400 miles, and upwards of 600 miles if from other subpopulation centers.

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Figure 37. Map of wolverine in the NRIS Wildlife Database, 2016

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Habitat Status Broadly, wolverines are restricted to boreal forests, tundra, and western mountains. The vegetation zones occupied by wolverines include the Arctic Tundra, Subarctic-Alpine Tundra, Boreal Forest, Northeast Mixed Forest, Redwood Forest, and Coniferous Forest (Banci 1994). The wolverine is a terrestrial mammal that occurs in a wide variety of alpine, boreal, and arctic habitats (USFWS 2011). They are absent from all other vegetation zones, including the prairie, deciduous, and mixed forests of eastern North America; California grassland-chaparral; and sagebrush and creosote scrublands (Banci 1994).

Researchers have generally agreed that wolverine "habitat is probably best defined in terms of adequate year-round food supplies in large, sparsely inhabited wilderness areas, rather than in terms of particular types of topography or plant associations" (Kelsall 1981).

Wolverines naturally occur at low densities, and require cold areas that maintain deep, persistent snow cover into the warm season for successful denning, (USFWS 2011). Within the contiguous United States wolverine habitat is restricted to high-elevation areas in the West (Figure 38). Their current distribution includes functioning populations in the North Cascades Mountains and the northern Rocky Mountains, as well as populations that have not yet reestablished in the southern Rocky Mountains and the Sierra Nevada (Ibid.).

Consistent with field observations indicating that all wolverine reproductive dens are located in areas that retain snow in the spring (Magoun and Copeland 1998), Aubry et al. (2007) concluded that the distribution of persistent spring snow cover was congruent with the wolverine’s historical distribution in the contiguous United States. This relationship was further supported by the findings of Schwartz et al. (2007) that showed historical wolverine populations in the southern Sierra Nevada of California, which occupied a geographically isolated area of persistent spring snow cover, were genetically isolated from northern populations.

Information on the use of natal dens in which the kits are born by wolverines in North America is biased to tundra regions where dens are easily located and observed. These natal dens typically consist of snow tunnels up to 60 meters in length (Banci 1994). Bedding does not appear necessary, inasmuch as kits were found in shallow pits dug on the ground (Ibid.). Snow tunnels in northwest Alaska were also used by lone wolverines (Ibid.), suggesting that they dig tunnels or use existing tunnels as resting sites as well.

Natal dens above treeline appear to require snow 1 to 3 meters deep (Banci 1994) that persists into spring. In Finland, one researcher believed that dens that wolverine had dug themselves were preferred, because caves were rarely used, although available (Ibid.). Little is known of the distribution of den sites in the landscape. The proximity of rocky areas, such as talus slopes or boulder fields, for use as dens or rendezvous sites was important for wolverines in Norway (Ibid.), in the Soviet Union (Ibid.), and in Idaho (unpublished data in Copeland 1993). Natal dens may be located near abundant food, such as cached carcasses or live prey (Rausch and Pearson 1972; Banci 1994).

Limited information is available on dens in forested habitat. In northern Lapland, most of the dens in forests were associated with spruce (Picea sp.) trees; five consisted of holes dug under fallen spruces, two were in standing spruces, and one natal den was in-side a decayed, hollow spruce and it was reported that dens in Kamchatka were usually constructed in the "hollows" (cavities) of large trees (Banci 1994). Rarely, kits have been found relatively unprotected, on branches and on the bare ground. If females are disturbed they will move their kits, often to what appear to be unsuitable den sites (Ibid.).

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Figure 38. Modeled wolverine habitat in the western United States

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It has been hypothesized that one of the factors affecting the selection of a natal den site was the ease with which it could be adapted to a den. It was reported dens in abandoned beaver lodges (Banci 1994), old bear dens, creek beds, under fallen logs, under the roots of upturned trees, or among boulders and rock ledges. In Siberia, dens were found in caves, under boulders and tree roots, and in accumulations of woody debris consisting of broken or rotted logs and dry twigs (Ibid.). Natal dens in Montana were most commonly associated with snow-covered tree roots, log jams, or rocks and boulders (Ibid.).

Foraging Wolverines can travel long distances in their daily hunting, 30 to 40 kilometers being "normal" (Banci 1994). Adult males generally cover greater distances than do adult females (Ibid.) and may make longer and more direct movements (Ibid.). During late winter, lactating females with young move less than solitary adult females (Ibid.). In May and June, hunting mothers periodically return to their young that have been left at rendezvous sites (Ibid.). In northwest Alaska, females returned to rendezvous sites at least daily (Ibid.). Kits were moved to new rendezvous sites every 1 to 9 days and more frequently as they grew older (Magoun 1985). By June, kits were moved every 1 to 2 days (Ibid.). When her kits were 4 to 11 weeks old, a female in central Idaho used 18 to 20 den sites, moving her kits a total of about 26 kilometers (unpublished data in Copeland 1993).

In northwest Montana, wolverines of both sexes made frequent long movements out of their home ranges that lasted from a few to 30 days, and they always returned to the same area (Hornocker and Hash 1981). These long-distance movements appear to be temporary and not attempts to expand the home range. Whether these movements are exploratory or whether wolverine are returning to previously known feeding locations is unknown (Banci 1994)

Except for females providing for kits or males seeking mates, movements of wolverine are generally motivated by food. Wolverines restrict their movements to feed on carrion or other high quality and abundant food sources (Gardner 1985; Banci 1988).

Home range Home ranges of adult wolverine in North America range from less than 100 square kilometers to over 900 square kilometers. The variation in home range sizes among studies partly may be related to differences in the abundance and distribution of food. (Banci 1994)

Young females typically establish residency next to or within the natal home range (Banci 1994). Although some immature females disperse, males are more likely to do so. Male wolverines may disperse either as young-of-the-year or as subadults (Banci 1994, Banci 1988). Dispersal can include extensive exploratory movements (Banci 1994, Banci 1988).

The presence of young restricts movements and home range size of females. Yearly home ranges for a female with young was 47 square kilometers (discounting 2 long-distance movements) in southwest Yukon; 100 square kilometers each for 2 females in Montana; a mean of 105 square kilometers in south- central Alaska; and a mean of 70 square kilometers in northwest Alaska (Banci 1994). Male home ranges are typically larger than those of females (Ibid.). Spring and summer home ranges of adult males, but not adult females, in-creased during the breeding season in Alaska and Montana but not in the Yukon (Ibid.). In the latter, localized and abundant food may have been responsible for females being readily available to the adult male, making extensive breeding movements unnecessary (Ibid.).

This pattern of home range use is consistent with a carnivore spatial strategy in which the spacing of females underlies the distribution of males, at least in the breeding season, but food underlies the distribution of females (Banci 1994). Home ranges of females should reflect the minimum size necessary

248 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests to obtain food more than those of males (Banci 1994). Consistent with this prediction, wolverine females typically cover their home ranges uniformly, unless they have kits and concentrate their movements at natal dens or rendezvous sites (Ibid.). Males, instead, typically have one or more foci of activity within the home range (Ibid.).

Winter home ranges typically overlap with those used in the snow-free season but also include different habitats, even if there are no significant differences in the size of seasonal home ranges (Banci 1994, Banci 1998). Differences between seasonal home ranges can be attributed to changes in prey distribution and availability. Wolverines of both sexes appear to maintain their home ranges within the same area between years (Banci 1994, Banci 1988). There may be slight changes in the yearly boundaries of home ranges with the addition of juvenile females adjacent to the natal area, with mortality, and with immigration. For example, when a resident dies, a neighbor may assume part of the vacant home range (Banci 1994; Banci 1988).

Rivers, lakes, mountain ranges, or other topographical features do not seem to block movements of wolverines (Banci 1994, Banci 1988). Considering the wolverine's avoidance of human developments, extensive human settlement and major access routes may function as barriers to dispersal (Banci 1994)

Security cover Predation may influence wolverine habitat use, depending on the predator complement in the environment, including humans. In south-central Alaska, wolverine use of rock outcrops was greater than the availability of those areas during summer (Banci 1994), perhaps because rock outcrops were being used as escape cover from aircraft. However, wolverines may have also been hunting marmots and collared pikas (Ochotona collaris) (Ibid.). In Squaw Valley California a wolverine was observed in a marmot (Marmota flaviventris) colony in July 1953 (Ruth 1954). Scat suspected from that animal also contained indigenous squirrel remains (Ibid.).

Wolverines may climb trees to escape wolves (Banci 1994), although if the trees are not high enough, such attempts may be unsuccessful (Burkholder 1962). Wolverines are found in a variety of habitats and do not appear to shun open areas where wolves are present. Wolverines occur locally with cougars, especially in British Columbia and the northwestern United States. Trees would not be an effective defense because cougars are adept at climbing. It is likely that wolverines use various habitat components, such as rock outcrops or trees, for escape when they feel threatened.

Aside from anecdotal reports, only one report on the use of resting sites by wolverines in forested habitats is published (Banci 1994). Overhead cover may be important for resting sites as well as natal and maternal dens. Resting sites in Montana were often in snow in timber types that afforded cover (Banci 1994).

Daily, seasonal, and breeding season movements Recently one researcher compiled (Inman 2012) most of the extant spatial data on wolverine denning and habitat use to test the hypotheses that wolverines require snow cover for reproductive dens (Magoun 1998), and that their geographic range is limited to areas with persistent spring snow cover (Aubry et al. 2007). Although Aubry et al. (2007) analysis covered only North America and used relatively coarse EASEGrid Weekly Snow Cover and Sea Ice Extent data (Inman 2012), these relationships with finer scale snow data (0.5-kilometer pixels) obtained throughout the Northern Hemisphere and were confirmed from the Moderate-Resolution Imaging Spectroradiometer (MODIS) instrument on the Terra satellite (Ibid.). Specifically, the data was compiled and evaluated the locations of 562 reproductive dens in North America and Scandinavia in relation to spring snow (Ibid.). All dens were located in snow and 97.9 percent were in areas identified as being persistently snow covered through the end of the wolverine’s

249 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests reproductive denning period (15 May; Aubry et al. 2007) based on MODIS imagery. Additionally, it was that areas characterized by persistent spring snow cover contained 89 percent of all telemetry locations from throughout the year in nine study areas at the southern extent of current wolverine range (Inman 2012). Excluding areas where wolverines were known to have been extirpated recently, persistent spring snow cover provided a good fit to current understandings of the wolverine’s circumboreal range (Ibid.). Moreover, it was found that the genetic structure of wolverine populations in the Rocky Mountains was consistent with dispersal within areas identified as being snow covered in spring, and strong avoidance of other areas. Thus, the areas with spring snow cover that supported reproduction (Ibid.) could also be used to predict year-round habitat use, dispersal pathways, and both historical (Aubry et al. 2007) and current ranges (Inman 2012). Security cover for kits may also be enhanced during winter; snow tunnels or snow caves are characteristic natal and maternal dens for wolverine in many areas (Banci 1994).

The reasons that wolverines of both sexes remain in areas with persistent spring snow cover throughout the year is not well understood. Summer use of these areas may be due to avoidance of summer heat (Inman 2012), prey availability in avalanche chutes and at timberline (Ibid.), or perhaps a combination of both. Whatever the cause, evidence suggests that wolverines occurring at the southern periphery of their range remain within a relatively narrow elevation zone throughout the year (Ibid.). There is no evidence, either currently or historically, that wolverine populations can persist in other areas. For these reasons, it was argued that the bioclimatic niche of the wolverine can be defined by the areal extent of persistent spring snow cover (Ibid.).

In one study males moved approximately 2 to 3 times farther than females on average (Inman 2012). Movement rates of dispersers were similar to resident adults with the exception that dispersers moved a greater maximum distance during a 24-hour period. Based on average 2-hour movement rates, adult wolverines traveled a distance equivalent to the diameter of the average home range in less than 2 days or around the circumference in less than 1 week. (Ibid.). Altercations between young males and adult males may be the proximate encouragement for the former to disperse (Banci 1988).

Data indicate that both males and females are capable of dispersing to areas at least 170 kilometers from their mother’s home range; however, this may also underestimate the distances wolverines disperse. One individual moved greater than 225 kilometers between 36 months and 40 months of age (Ibid.). Wolverines estimated to be 2 to 3 years old made several movements of approximately 200 kilometers in Idaho (Moriarty 2009). Wolverines have traveled as far as 300 kilometers and 378 kilometers in Alaska (Banci 1994), and genetic sampling suggests the potential for wolverines to disperse as much as 500 kilometers (Moriarty 2009).

Prey species All studies have shown the paramount importance of large mammal carrion, and the availability of large mammals underlies the distribution, survival, and reproductive success of wolverines. Over most of their range, ungulates provide this carrion, although in coastal areas, marine mammals may be used. Wolverines are too large to survive on only small prey (Banci 1994).

Bone and hide may be important foods. They may be available for several months after an ungulate dies (Banci 1994). Small mammals are primary prey only when carrion of larger mammals is unavailable (Banci 1988).

Porcupines (Erethizon dorsatum) occur in wolverine diets in Alaska, the Yukon, and Montana. Although they represent a large meal, porcupines appear to be limited to those wolverines that have learned to kill them (Banci 1988). The frequency of red squirrels (Tamiasciuris hudsonicus) in wolverine diets in northern forested habitats (Gardner 1985, Banci 1988) is a reflection of their wide distribution and

250 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests availability throughout winter. Arctic ground squirrels (Spermophilus parryi) composed 26 percent of all sciurids in the winter diet of Yukon wolverines (Banci 1988) and the majority of the diet in northwest Alaska, where snowshoe hares were absent (Magoun 1985). Wolverines cache hibernating sciurids such as ground squirrels and hoary marmots (Marmota caligata) in the snow-free months for later use and excavate them from winter burrows (Gardner 1985; Magoun 1985). Birds occur in the diet according to their availability. Wolverines prey on ptarmigan (Lagopus spp.) in winter in the Yukon (Banci 1988), Alaska, and the Northwest Territories (Banci 1994), Prey that occurs sporadically in diets, such as American marten, weasel (Mustela spp.), mink (M. vison), lynx, and beaver (Castor canadensis), likely are mostly scavenged. Vegetation is consumed incidentally although ungulate rumens and may contain nutrients that wolverines cannot obtain from other foods (Banci 1988).

General habitat relationships of primary prey Although wolverines are mostly scavengers, they can prey on ungulates under some conditions. Because of their low foot loads (pressure applied to substrate) of 22 grams per square centimeter (Banci 1994), wolverines can prey on larger mammals in deep snow and when ungulates are vulnerable. Caching of food by wolverines has been described by most studies except that in Montana. The frequency of caching by wolverines may be affected in various ways by the presence of other carnivores (Ibid.).

Although data are limited, in general, diets during snow-free periods are more varied than in winter. These activities could include wolf predation, excessive harvesting by humans and human-caused losses of ungulate winter ranges. Some ungulate species may be enhanced by the provision of early seral stages through logging or burning. However, these and other land-use activities may exclude wolverines from areas that ungulates still use if these habitats do not provide for the other the wolverine life needs (Banci 1994). Because young wolverines mature rapidly, the availability and distribution of food during the snow-free season may determine the survival of females with kits (Ibid.).

Preferences for some forest cover types, aspects, slopes, or elevations have been primarily attributed to a greater abundance of food (Banci 1994, Banci 1988), but also to avoidance of high temperatures and of humans (Banci 1994). The greater use of subalpine coniferous habitats by males in southwest Yukon in winter was speculated to be due to higher densities of ungulate kills in these habitats (Banci 1988). Similarly, the use of alpine areas in south-central Alaska in summer was attributed to the arctic ground squirrels there (Banci 1994). In Montana, it was believed that wolverines used higher ranges during the snow-free season because they were avoiding high temperatures and human recreational activity (Ibid.).

The shrinking range of wolverines coincided with that of wolves in the late 1800's and the early 1900's. In some areas, predator control was coupled with the decimation of large mammal populations, such as the northern caribou herds (Banci 1994), reducing food available to wolverines.

Threats Wolverines and Sierra Nevada red foxes were vulnerable to historical trapping; however, they are also described as being extremely sensitive to the presence of people (Banci 1994). Most of the native generalist mesocarnivores including gray fox (Urocyon cinereoargentus), striped skunk (Mephitis mephitis), spotted skunk (Spilogale gracilis), and ringtail (Bassariscus astutus), and one large carnivore black bear (Ursus americanus) appear to occupy regions today that were also occupied in the early 1900s (Zielinski 2004). Over the same period of time, the distribution of two exotic and generalist species, the lowland red fox (V. vulpes), (Ibid.) and the opossum (Didelphis virginanus); a non-carnivore that is detected regularly at track-plate stations, have increased. The regions currently occupied by two forest specialists, the fisher (Martes pennanti) and the marten (M. americana), appear to have decreased compared with their historical distributions. Fishers are apparently absent from the region from Mount

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Shasta south to Yosemite, and martens are distributed patchily in the southern Cascades and northern Sierra Nevada (Zielinski 2004)

Injuries believed to have been inflicted by a cougar (Felis concolor) were described (Banci 1994) and it has been suggested that bears and eagles could kill wolverines, especially kits. The importance of predation on wolverine kits has not been documented. Wolverine mothers go to great lengths to find secure dens for their young, suggesting that predation may be important. Although not documented, adult males may kill kits. Some wolverines, especially males, may be killed by conspecifics. Males in northwest Alaska had fresh wounds on their heads when captured in April, suggesting that the approach of the breeding season in-creases aggressive behavior (Ibid.).

Starvation likely is an important mortality factor for young and very old wolverines. Suspected deaths from starvation include two young-of-the-year females in southwest Yukon (Banci 1988) and a young female and an old male in Montana (Banci 1994). These animals relied heavily on baits just before their deaths, suggesting that very young and old age classes may be unsuccessful foragers, even if food is abundant (Banci 1994, Banci 1988). Documenting the fates of young males is difficult be-cause of their extensive movements and it is not possible to predict whether sexes differ in their susceptibility to starvation (Banci 1994).

The age-specific mortality reported in studies of collared wolverines was 57 percent for adults, 7 percent for subadults, and 36 percent for young of the year. However, the mortality rates of juvenile wolverines are underestimated in these studies. The long distances covered by young of the year and sub-adults, especially males, makes it difficult to as-certain their fates unless they are trapped and their deaths reported. Mortality in these young age classes likely is substantial. Transients likely have a higher mortality rate than residents because they do not benefit from hunting in familiar home ranges. So, they likely have a greater chance of starvation, of being killed by conspecifics and of encountering traps (though wolverine trapping no longer occurs in California) (Banci 1994). It is believed that one-third to one-half of subadult wolverines perished during dispersal (Ibid.).

Climate changes are predicted to reduce wolverine habitat and range by 23 percent over the next 30 years, and 63 percent over the next 75 years, rendering remaining habitat significantly smaller and more fragmented. This increased fragmentation and isolation of subpopulations is expected to limit the regular dispersal of wolverines that is necessary to maintain genetic exchange and metapopulation dynamics. Other secondary threats to the wolverine that could work in concert with climate change include harvest, disturbance, infrastructure, transportation corridors, and small effective population sizes. The primary threat of habitat and range loss due to climate change would affect wolverine habitat and, therefore, the magnitude of threats to the wolverine is high (USFWS 2011). If these scenarios are valid, then conservation efforts should focus on maintaining wolverine populations in the largest remaining areas of contiguous habitat and, to the extent possible, facilitating connectivity among habitat patches (McKelvey et al 2011). However climate change has not yet had a detectable effect on the wolverine to this point in time; the threat is nonimminent (USFWS 2011).

Population Management Strategies Refugia, large areas that are not trapped and free from land-use impacts, can serve as sources of dispersing individuals and have been shown to be effective at ensuring the persistence and recovery of fisher and American marten populations (Banci 1994). The persistence of wolverine populations in Montana, despite years of unlimited trapping and hunting, was attributed solely to the presence of designated wilderness and remote, inaccessible habitat (Banci 1994). Wolverines persisted in southwestern Alberta despite their extirpation elsewhere in the province, largely because of the presence of large refugia in the form of national parks (Ibid.).

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Wolverine is only known from one male individual in the Tahoe National Forest. Although historic within the range and shown on the maps above, it is doubtful that wolverine exist within the Inyo, Sequoia, or Sierra National Forests. First, no wolverines have been detected in the camera tracks, track plates, or hair snare traps. Second, the lack of continuous deep snow pack make this area less desirable for wolverine.

However, due to the elusivity of wolverine, the locations that they would exist is in the high mountain area where the snow pack would be deep in winter. Habitat components such as those associated with late seral stage components will be maintained or enhanced. Threats such as uncharacteristic wildfires are being ameliorated through the implementation of the plan. Climate change is being addressed through creation and maintenance of corridors and landscape linkages so that the species can move from one location to another. However, forest plan direction proposed in all alternatives would not modify snowpack.

Recreational activities that may occur could have an impact of noise disturbance. Hunters that leave gut piles or that only wound an animal that would later die, hence becoming food for the wolverine. Although it is expected that recreation will increase over the life of the plan, the locations where wolverine would be, there still would be limited impact due to the lack of trails and easy access to the high mountain areas.

Determination Statement Inyo, Sequoia and Sierra NFs: Implementation of the Forest Plan may affect wolverine but will not lead towards Federal listing or a loss of viability. Impacts to wolverine are beneficial impacts such as reducing the risk of catastrophic wildfires by restoring conifer areas thus protecting potential den sites. Other benefits are maintaining and improving the aquatic zones that could provide food source for wolverines.

Literature Cited - Wolverine Aubry et al. 2007. Distribution and Broadscale Habitat Relations of the Wolverine in the Contiguous United States. The Journal of Wildlife Management 71(7).

Banci, Vivian. And Harestad, Alton. 1988. Reproduction and Natality of Wolverine (Gulo gulo) in Yukon. Ann. Zool. Fennici 25:265-270.

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Banci. V. 1994. Wolverine. In: The scientific basis for conserving forest carnivores: American marten, fisher , lynx, and wolverine in the western United States. U.S. Department of Agriculture Forest Service, Rocky Mountain Forest and Range Experiment Station. General Technical Report RM- 254. Fort Collins, CO.

Inman et al. 2010. Spatial Ecology of Wolverines at the Southern Periphery of Distribution. Journal of Wildlife Management 76(4)

Kyle, Christopher J. and Stroebeck, Curtis. 2002. Connectivity of Peripheral and Core Populations of North American Wolverines. Journal of Mammology, 83 (4).

McKelvey, et al. 2011. Climate change predicted to shift wolverine distributions, connectivity, and dispersal corridors. In: Ecological Applications, 21(8)

Magoun Audrey J. and Copeland, Jeffrey P. 1998; Characteristics of Wolverine Reproductive Den Sites. The Journal of Wildlife Management, Vol 62, No. 4.

Moriarty, K., W. Zielinski, A. Gonzales, T. Dawson, K. Boatner, C. Wilson, F. Schlexer, K. Pilgrim, J. Copeland, and M. Schwartz. 2009. Wolverine confirmation in California after nearly a century: native or long-distance migrant? Northwest Science. 83(2).

Pasitschniak-Arts, M. and S. Larivière. 1995. Gulo gulo. In: Mammalian Species. No. 499. Pp. 1-10.

Rausch, R. A. and Pearson A. M. 1972; Notes on the Wolverine in Alaska and the Yukon Territory, The Journal of Wildlife Management, Vol. 36, No. 2.

Ruth, Ferdinand S. 1954. Wolverine Seen in Squaw Valley, California. Journal of Mammology, Vol. 35, No.4.

Schwartz Michael K. et al. (2007) Inferring Geographic Isolation of Wolverines in California. Journal of Wildlife Management. 71(7).

Schwartz, Michael K et al. 2009. Wolverine Gene Flow across a Narrow Climatic Niche. In: Ecology, Vol. 90, No. 11.

U.S. Department of Interior Fish and Wildlife Service (USFWS). 2011. Endangered and Threatened Wildlife and Plants; 12-Month Finding on a Petition To List the North American Wolverine as Endangered or Threatened. Federal Register. Vol. 75, No. 239, 78030-78061.

Zielinski, William J. 2004. The Status and Conservation of Mesocarnivores in the Sierra Nevada. USDA Forest Service Gen. Tech. Rep. PSW-GTR-193

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Western Pond Turtle (Actinemys marmorata) USDA Forest Service: Sensitive Species

Species account The western pond turtle (Actinemys marmorata and Actinemys pallida) is found on the west coast of North America. Historically it was found from as far north as British Columbia, Canada to as far south as Baja California mostly west of the Cascade-Sierra crest (Lovich and Meyer 2002). Fossil fragments have been found east of the current range indicating that the species was once more widespread (Buskirk 2002). Disjunct populations have been documented in the Truckee, Humboldt and Carson Rivers in Nevada, Puget Sound in Washington, and the Columbia Gorge on the border of Oregon and Washington. It is currently unclear if these are relictual or introduced populations (Lovich and Meyer 2002). Modern distribution is limited to parts of Washington, Oregon, California and northern Baja California (Buskirk 2002). Western pond turtles are the only native aquatic turtle in California and southern Oregon, in the northern part of its range it coexists with only the western painted turtle (Chrysemys picta bellii) (Germano and Rathbun 2008).

On Region 5 lands this turtle can be found on all National Forests, except the Inyo and Lake Tahoe Basin.

Official taxonomy by the Society for the Study of Amphibians and Reptiles has placed the turtle into a new genus, Actinemys. This is confirmed by both NatureServe and Itis as the current genus. However, Emys is still used within the scientific literature. In addition, the Society no longer recognizes subspecies for the western pond turtle. Presumably this is based on recent genetic work that indicates that the recognized subspecies were not geographically or genetically correct, and the currently recognized species likely represents as many as four cryptic species. However, the study that identified the four distinct clades of pond turtle did not elevate any to species status as the authors wanted to wait until further molecular work was undertaken. The two former subspecies were the northwestern pond turtle (Emys marmorata marmorata) and the southwestern pond turtle (Emys marmorata pallida) with a subspecies split along the transverse mountain range in southern California (Spinks and Shaffer 2005).

Further work by Spinks et al. (2014), proposed to have all species north of San Francisco Bay area plus populations from the Great Central Valley north including the apparently introduced Nevada population as Emys marmorata. Emys pallida is restricted to those populations inhabiting the central coast range south of the San Francisco Bay area to the species southern range boundary, including the Mojave River (Spinks et al. 2014). Spinks et al. (2014) further found very limited intergradation between Emys marmorata and Emys pallida in a few populations in the northern central coast range and adjacent Sierra Nevada foothills, although pure individuals of the locally prevalent species were also noted. These individuals may represent human-mediated translocations/pet releases, a phenomenon that regularly occurs in turtles (Carr 1952, Storer 1930 in Spinks et al. 2014).

Abundance has been well studied in this species. In some stream habitats densities can exceed 1,000 turtles per hectare. In Oregon, small ponds can hold over 500 turtles per hectare. These densities represent extremes with typical densities ranging from 23 to 214 turtles per hectare throughout most of the range (Lovich and Meyer 2002). Capture rates at one site in southern California were about 2 to 2.6 turtles per trap night (Germano 2010). These density estimates are likely accurate for populations on National Forest System lands where habitat is suitable.

In the NRIS database, the Inyo NF has no records, Sequoia NF has no records, and the Sierra NF has 1,433 records. Figure 39 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

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Figure 39. Map of western pond turtle from multiple sources, 2010

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Habitat Status The western pond turtle inhabits a Mediterranean climate defined by mild, wet winters and long hot, dry summers. In the northern portion of its range winters are colder with more rainfall than in southern areas (Germano and Rathbun 2008). Aquatic habitats include lakes, natural ponds, rivers, oxbows, permanent streams, ephemeral streams, marshes, freshwater and brackish estuaries and vernal pools. Additionally, these turtles will utilize man-made waterways including drainage ditches, canals, reservoirs, mill ponds, ornamental ponds, stock ponds, abandoned gravel pits, and sewage treatment plants (Buskirk 2002). Turtles captured at waste-water treatment plants grew quickly, had successful recruitment and produced large clutches (Germano 2010). Turtles favor areas with offshore basking sites including floating logs, snags, protruding rocks, emergent vegetation and overhanging tree boughs, but also will utilize steep and/or vegetated shores. Hatchlings additionally require shallow, eutrophic, warm areas which are typically at the margins of natural waterways (Buskirk 2002). Terrestrial habitats are less well understood. In southern California animals spend only one to two months in terrestrial habitats while animals in the northern portions of the range can be terrestrial for up to eight months (Lovich and Meyer 2002). Animals have been documented to overwinter under litter or buried in soil in areas with dense understories consisting of vegetation such as blackberry, poison oak and stinging nettle which reduces the likelihood of predation (Davis 1998).

Western pond turtles are generalist omnivores and have been documented to eat a wide variety of prey. Prey items include larval insects, midges, beetles, filamentous green algae, tule and cattail roots, water lily pods, and alder catkins (Germano 2010). Filamentous algae is considered to be an important food source for females after egg laying (Buskirk 2002). Additionally, animals will opportunistically feed on other items such as floating duck carcasses, ducklings (pers. obs.) and dog food in backyards while on walkabouts (Buskirk 2002).

Turtles move upland at different times across the range of this species. Animals can move upland as early as September, but typically move following the first winter storm in November or December. Not all animals move upland, some move to nearby ponds for the winter (Davis 1998). Animals have been observed moving underneath ice in ponds and potentially congregate in shallow areas (Buskirk 2002). Upland animals remain somewhat active throughout the winter and can be observed basking on warm winter days (Davis 1998). Upland movements for both overwintering and reproduction typically occur in the afternoon and evenings. Walkabouts to scout for nest sites can be completed within one day or they can last up to four days (Crump 2001). Home ranges differ between males and females with male home ranges averaging 0.976 hectares and females averaging 0.248 hectares. Although western pond turtles are likely not territorial, disputes over basking sites are commonly observed (Buskirk 2002).

Local climatic and water level variations can alter the timing of nesting in this species (Crump 2001). The nesting season is from late April through mid-July at low elevation, and June through August at higher elevations (Scott et al. 2008). Although some females can reproduce with a carapace length as small as 111 millimeters, 120 millimeters is the minimum reproductive size in most areas with most gravid females being at least 140 millimeters (Scott et al. 2008). Animals of this size are often at least seven years old in southern areas and eight to twelve years old in northern areas. Western pond turtles have an average life expectancy of approximately forty years if they survive to adulthood (Buskirk 2002).

Some western pond turtles have shown nest site fidelity. Four of five detected nesting areas in one study area had instances of nest side fidelity. It is likely that nest site fidelity is common, and sites are changed only after a negative encounter during either a walkabout or while forming a nest at a particular site (Crump 2001). Most females nest within 50 meters of water; however some females nest upwards of 400 meters away from water (Lovich and Meyer 2002). It is believed that in coastal populations nesting

257 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests occurs far from water in order to protect overwintering hatchlings from being injured during winter floods (Lovich and Meyer 2002).

Mean clutch size ranges from 4.5 with a measurement error of 0.25 eggs on the Santa Rosa Plateau to 7.3 with a measurement error of 1.18 eggs in southern Oregon. More research is needed to determine if clutch size varies with latitude (Germano and Rathbun 2008). Average annual egg production for 39 animals in southern California was 7.2 with a measurement error of 3.9 eggs. This number did not vary statistically among females of differing carapace length or among different streams and in many cases represented two clutches per female. Clutch size varies significantly among drainages; however it does not differ significantly across years or within individual drainages. When double clutching occurs, the first clutch typically contains more eggs than the second clutch (Scott et al. 2008). Several head-starting programs claim that temperature dependent sex determination is utilized by western pond turtles, but they have not published evidence proving this (Buskirk 2002).

Hatchlings in the Mojave River population overwinter in the nest and emerge as early as March of the following year (Lovich and Meyer 2002). However, most hatchlings in southern California emerge in late fall of the year they were laid. Northern animals typically emerge the following spring. Delayed emergence can be caused by soil structure where sandy soil results in earlier emergence (Crump 2001). Microhabitat use, behavior and diet differ between juvenile and adult western pond turtles (Lovich and Meyer 2002). Little is known about the specific requirements of hatchling turtles as they are cryptic and are rarely represented in population assessments of many species including those with known stable populations (Germano and Rathbun 2008).

Growth and maturation in western pond turtles is heavily influenced by ambient air and water temperatures and basking behaviors which include aerial basking, and cryptic behaviors such as burying in warm sand or lying in warm algal mats (Germano and Rathbun 2008). Sites with cold water require turtles to bask more causing average body size to be smaller compared to sites with warmer water. Areas which have higher invertebrate densities typically classified as having organic mud bottom substrates yield larger turtles (Lubcke and Wilson 2007).

Threats Western pond turtles have significantly declined in number with many populations representing less than 10 percent of the historical population. In California alone there has been a loss of 80 to 85 percent of western pond turtles since the 1850’s. The Puget Sound population in Washington, which encompassed the type location for this species, as well as British Columbia populations has been considered extirpated since at least the 1970s. Ninety-eight percent of the population is gone in Oregon’s Willamette Valley, 95 to 99.9 percent of the population in the San Joaquin Valley is gone and most of the Nevada populations have disappeared. In southern California there are only 12 known viable populations (greater than 25 adult animals) between Los Angeles County and the Mexican border (Buskirk 2002).

The major threat to this species is habitat loss or degradation. Most of the historical habitat for this species has been permanently lost as a result of development for human occupancy. Riparian and wetland habitats are cleared for agriculture use, destroyed by cattle, channelized and stripped of vegetation, or invaded by the saltcedar shrub which destroys water quality, alters stream structure and dries streams. Ground water pumping lowers water tables and further stresses riparian plant communities. Gold and gravel mining can both directly destroy habitat as well as introduce toxins through toxic spills and illegal dumping of chemicals (Buskirk 2002, Lovich and Meyer 2002).

Additional human-caused threats further jeopardize population viability. Cattle grazing can destroy riparian habitat, cattle trample and kill turtles and nests, and cattle waste pollutes waterways. Western

258 Draft Biological Evaluation, Wildlife and Fish Forest Plan Revision Inyo, Sequoia, and Sierra National Forests pond turtles, especially gravid females, are easily killed on roadways by direct impact with vehicles. Historically animals were also collected for the pet trade with hundreds of animals from a single site being exported to Europe in the 1960’s. Although collection and sale of western pond turtles have been banned for many years, animals are still listed for sale in the eastern United States. Animals were collected for food in great numbers from the mid-19th century to the 1930s when animals first started to become scarce. Modern watercourse recreation also impacts these turtles. Recreation which interferes with basking or causes direct injury or mortality include high-speed boating, water skiing, jet skiing and fishing where animals may be directly caught or killed because they are viewed as competition (Buskirk 2002).

Disease poses a notable threat to western pond turtles, as has been seen in Washington. A die-off in 1990 was attributed to a syndrome similar to an upper-respiratory disease. Several years later, as part of a head- starting program, several animals were found dead with no apparent cause of death (Vander Haegen et al. 2009). Animals from a wastewater treatment pond in California were found to be less healthy in both the short and long term compared to animals in a natural habitat despite being larger in size. Although larger, these animals had more chronic stress in the form of more interactions with humans and invasive species, increased water pollution and greater exposure to water-borne diseases (Polo-Cavia et al. 2010). Dehydration also poses a threat to turtles under a year old which likely makes these animals more susceptible to disease (Vander Haegen et al. 2009).

In addition to threats that affect entire populations, many populations are failing as a result of extremely high juvenile mortality. While adults may have annual survival rates of 95 to 97 percent, nests, juveniles and sub-adults have extremely high mortality rates (Vander Haegen et al 2009). Nest destruction by raccoons can approach or reach 100 percent of nests at many Oregon nest sites (Buskirk 2002). Nests are also destroyed when exposed to too much moisture or are crushed by cattle or machines. There are many predators of hatchling turtles, including two very successful non-native predators- large-mouth bass and bullfrogs. Subadult mortality can be as high as 85 to 90 percent annually for animals under 4 years old, however head-started subadults had mortalities as low as 10 percent when carapace length was greater than 90 millimeters. Natural predators that have been documented to take sub-adult turtles include: raccoons, coyotes, black bears and western river otters with most predations occurring while the animal was terrestrial (Vander Haegen et al. 2009). Adults face less predation risk. A study documented one predation of an adult turtle by a loon, and only 3 of 196 turtles had evidence of predation attempts such as shell or limb damage (Davis 1998).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will refer to the analysis in the BE for that plan.

Alternative B, C, and D: Restoration and or maintenance of riparian zones is proposed for each of these alternatives. Reduction in uncharacteristic wildfires is proposed through restoration, thus reduced sedimentation and maintenance of water quality should occur. Removal of invasive species will reduce the impacts and or competition from other species.

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Sequoia and Sierra NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, removing invasive species, and reducing the risk of catastrophic wildfire.

Literature Cited – Western Pond Turtle Buskirk, J. 2002. The western pond turtle, Emys marmorata. Radiata 11(3): 3-30.

Crump, D.E. 2001. Western pond turtle (Clemmys marmorata pallida) nesting behavior and habitat use. Master’s Theses. Paper 2210. http://scholarworks.sjsu.edu/etd_thesis/2210.

Davis, C.J. 1998. Western pond turtle (Clemmys marmorata). Master’s Theses. Paper 1694. http://scholarworks.sjsu.edu/etd_thesis/1694.

Germano, D.J. 2010. Ecology of western pond turtles (Actinemys marmorata) at sewage-treatment facilities in the San Joaquin Valley, California. The Southwestern Naturalist, 55(1): 89-97.

Germano, D.J. and Rathbun, G.B. 2008. Growth, population structure, and reproduction of western pond turtles (Actinemys marmorata) on the central coast of California. Chelonian Conservation and Biology 7(2): 188-194.

Lovich, J. and Meyer, K. 2002. The western pond turtle (Clemmys marmorata) in the Mojave River, California, USA: highly adapted survivor or tenuous relict? Journal of Zoology London 256: 537- 545.

Lubcke, G.M. and Wilson, D.S. 2007. Variation in shell morphology of the western pond turtle (Actinemys marmorata Baird and Girard) from three aquatic habitats in northern California. Journal of Herpetology 41(1): 107-114.

Polo-Cavia, N., Engstrom, T., Lopez, P., and Martin, J. 2010. Body condition does not predict immunocompetence of western pond turtles in altered versus natural habitats. Animal Conservation 13: 256-264.

Scott, N.J., Rathbun, G.B., Murphy, T.G., and Harker, M.B. 2008. Reproduction of pacific pond turtles (Actinemys marmorata) in coastal streams of central California. Herpetological Conservation and Biology 3(2): 143-148.

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Spinks, P.Q. and Shaffer, H.B. 2005. Range-wide molecular analysis of the western pond turtle (Emys marmorata): cryptic variation, isolation by distance, and their conservation implications. Molecular Ecology 14: 2047-2064.

Spinks, P.Q., R.C. Thomson, and H.B. Shaffer. 2014. The advantages of going large: genome-wide SNPs clarify the complex population history and systematics of the threatened western pond turtle. Molecular Ecology 23: 2228-2241.

Vander Haegen, W.M., Clark, S.L., Perillo, K.M., Anderson, D.P., and Allen, H.L. 2009. Survival and causes of mortality of head-started western pond turtles on Pierce National Wildlife Refuge, Washington. The Journal of Wildlife Management 73(8): 1402-1406.

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Panamint Alligator Lizard (Elgaria panamintina) USDA Forest Service: Sensitive Species

Species account The Panamint alligator lizard is endemic to the desert mountains of Inyo and Mono Counties, California. This species is considered to be a relictual species that became isolated from the similar Elgaria multicarinata and Elgaria kingii as the surrounding area transformed from a well-watered area to a dry desert (Stebbins 1958). Its current range is limited to canyons and creeks with a permanent source of water (Jennings and Hayes 1994). The known area of occupancy (Figure 40) is very small represented by 23 known localities, and is presumably less than five square kilometers total (Hammerson 2007, Jennings and Hayes 1994). Additionally, all but two known populations are on private lands (Jennings and Hayes 1994). Its elevational range is from 2,500 to 7,500 feet (Mahrdt and Beaman 2009).

Within Region 5 this species only occurs on the Inyo National Forest. This species has been documented in several locations on this forest in the White and Inyo Mountains from the Nevada and California border south to Conglomerate Mesa.

Although the species has been known since 1958, less than 30 individuals have been collected and deposited in museums, and less than 15 sight records have been submitted (Hammerson 2007). This is likely a direct result of the secretive nature of the species as well as the difficulty of surveying in the dense vegetation and talus slopes where this species has been observed (Stebbins 1958). The total population of this species is unknown, but is probably at least 1,000 individuals assuming each of the 23 documented locations have subpopulations of at least 50 animals. Presumably the population is stable, however abundance has likely declined in areas where habitat has been degraded and further study is needed (Hammerson 2007).

In the NRIS database, the Inyo NF has 22 records, Sequoia and Sierra NFs have no records. Figure 41 below displays the distribution, but note that many of the records are of the same location, just at a different time, and appear stacked on top of each other as one dot.

Habitat Status Like other alligator lizards, the Panamint alligator lizard is associated with mesic microhabitats in the vicinity of riparian areas and springs (Stebbins 1958). In low elevation sites dominated by creosote scrub, these animals are only found in the vicinity of water (Papenfuss 1985). These riparian areas are often extremely limited in size being only a few meters wide and 0.75 to 3.1 kilometers in length and are primarily confined to canyon bottoms (Stebbins 1958; Jennings and Hayes 1994). Higher elevation animals can be found further from water in boulder and talus slopes, amongst desert scrub and in pinon- juniper woodland (Mahrdt and Beaman 2009, Papenfus 1985, Stebbins 2003). Typical vegetation associated with this species includes various willows, Virgin’s bower (Clematis lugusticifolia), and wild grape (Vitis girdiana) (Mahrdt and Beaman 2009).

Life history data on this species is very limited. Pitfall studies indicate that Panamint alligator lizards are most active in May, June and September when temperatures are mild. They are thought to estivate or be nocturnal from July through August (Jennings and Hayes 1994). Panamint alligator lizards feed on invertebrates including insects and spiders. Breeding is thought to occur in April or May followed by oviposition in September. Juveniles are first observed in spring (Mahrdt and Beaman 2009). There is no information available about home range size, territoriality or activity during winter months.

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Figure 40. Map of Panamint alligator lizard from multiple sources, 2010

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Figure 41. Map of Panamint alligator lizard from NRIS Databases, 2016

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Threats The most significant threat to this species is its limited range located primarily on private lands. Current threats include habitat loss as a result of mining, feral and domestic livestock grazing, and off-highway vehicle use (Jennings and Hayes 1994). Over the last decade off-highway vehicle activity has increased significantly in the Panamint-Inyo-White Mountain system and will likely continue to increase (Mahrdt and Beaman 2009). Additionally, the introduction of the invasive plant salt cedar (Tamarix ramosissima) may impact this species (Mahrdt and Beaman 2002). The threats posed by predation and disease are not studied. No predators are documented for this species, however several species known to consume other alligator lizards are present including coachwhips (Coluber flagellum), striped whipsnakes (Coluber meniatus), loggerhead shrikes (Lanius ludovicianus) and red-tailed hawks (Buteo jamaicensis) (Jennings and Hayes 1994). It is currently illegal to pursue or collect this species under current California state regulations (C.C.R. 2012). This species is likely to be significantly impacted by future climate changes. As it is a relictual population of an environment that no longer exists, the odds are great that it will eventually go extinct as the remaining patches of mesic habitat are converted to desert in a warming, drying climate (Stebbins 1958).

Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will provide for the protection of the riparian zone that supports this species.

Alternative B, C, and D: Restoration and or maintenance of riparian areas is proposed for each of these alternatives. Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species. Removing invasive species, particular tamarisk, will be a benefit to this species.

Determination Statement Inyo NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. This species is considered stable across the Inyo NF, but it is in limited distribution within the Forest. Alternative B, C, and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

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Literature Cited – Panamint Alligator Lizard California Code of Regulations. 2012. Title 14, Chapter 2, Article 4, Section 5.60.

Hammerson, G. A. 2007. Charina bottae. In: IUCN Red List of Threatened Species. Version 2011.2 http://www.iucnredlist.org. Downloaded on 21 May 2012.

Jennings, M. R. and Hayes, M. P. 1994. Panamint Alligator Lizard Elgaria panamintina Stebbins 1958. In: Amphibian and Reptile Species of Special Concern in California. California Department of Fish and Game, Sacramento, California.

Mahrdt, C.R. and Beaman, K.R. 2002. Panamint alligator lizard, Elgaria panamintina. Species account, West Mojave Management Plan. Riverside, California.

Mahrdt, C.R. and Beaman, K.R. 2009. Panamint alligator lizard, Elgaria panamintina Stebbins, 1958. Pp. 488-491. In: Jones, L.E. and Lovich, R.E. (eds.) Lizards of the American Southwest A Photographic Field Guide. Rio Nuevo Publishers, Tucson, Arizona.

Papenfuss, T.J. 1985. Amphibian and reptile diversity along elevational transects in the White-Inyo Range. Pp. 129-136. In: Hall Jr., C.A., and Young, D.J. (eds.) Natural history of the White-Inyo Range, eastern California and western Nevada, and high altitude physiology. White Mountain Research Station Symposium, California.

Stebbins, R.C. 1958. A new alligator lizard from the Panamint Mountains, Inyo County, California. American Museum Noviates 1883: 1-27.

Stebbins, R.C. 2003. A Field Guide to Western Reptiles and Amphibians. Third edition. Houghton Mifflin, New York.

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California Legless Lizard (Anniella pulchra) USDA Forest Service: Sensitive Species

Species account The California legless lizard is found in the coastal ranges from Antioch (Contra Costa County), California south to Baja California, Mexico (Jennings and Hayes 1994, Morey 2000). Where habitat is suitable, this species is also found in the San Joaquin Valley, southern California mountains, foothills of the Sierra Nevada Mountains and the Mojave Desert (Figure 42). This species has been documented from sea level to near 6000 feet (Morey 2000). Formerly the dark animals found near Monterey, California were considered to be a separate subspecies but recent genetic studies found that they are paraphyletic with the more common light colored lizards (Pearse and Pogson 2000). Additional studies on the genetics of this species are ongoing and although there are several distinct lineages no subspecies are currently recognized for this species (Parham and Papenfuss 2009).

Within Region 5 the species has been documented on the Angeles, Cleveland, Los Padres, San Bernardino and Sequoia National Forests.

No baseline surveys have been conducted to determine population levels on National Forest System lands. Due to the secretive, fossorial nature of this lizard it is often undetectable by standard herpetofaunal survey techniques. Successful population estimates requires labor intensive surveys which require significant habitat disturbance. Standard survey techniques such as those which utilize cover boards missed a substantial amount of the known population and surveys utilizing these techniques may underestimate or fail to document a population of California legless lizards (Kuhnz et al. 2005).

Habitat Status California legless lizards inhabit a range of habitats including coastal dune, valley-foothill, chaparral and coastal scrub (Morey 2000). Populations are most dense along the coast indicating that sandy habitats are preferred (Klauber 1932). The predominant factors which define the habitat for this species and prevent range expansion are the moisture content of soil, ground temperature, soil structure, and vegetation. Legless lizards are not found in areas which have high clay soils (Miller 1944). Although they prefer loose soil, they can be found in areas of dense soil and among rocks as long as these areas have sufficient leaf litter for burrowing (Klauber 1932). Both the oak-annual grass association and lupine-stable dune association are preferred by this species as they allow for the build-up of leaf litter to burrow through (Kuhnz et al. 2005; Miller 1944). Animals are rarely encountered in the open, and are typically found under cover objects or in the course of soil excavation (Klauber 1932). Cover objects include stones, burlap sacks, boards, leaf litter and other debris (Klauber 1932; Miller 1944; Kuhnz et al. 2005). California legless lizards are typically found between one and three inches under cover, however they are documented to be as deep as a foot below the surface of the substrate (Miller 1944).

California legless lizards primarily consume insect larvae and adult beetles. They will also consume other small invertebrates such as small insects and spiders. Wild lizards have been observed eating large (more than 1.25 inches in length) insect larvae and two species of small ground dwelling beetles (Helops and Platyderma). In laboratory settings larvae of grain beetles (Tenebrio molitor) and termites (Zootermoptsis sp.) are readily eaten (Miller 1944). Foraging occurs at the base of vegetation either on the surface, in leaf litter or in sandy soil (Stebbins 2003).

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Figure 42. Map of California legless lizard for multiple sources, 2010

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Legless lizards are documented to passively drink water straight from moist soils (Fusari 1985). It is unlikely that legless lizards regularly encounter standing water as animals placed in water were unable to swim, attempted to burrow in the water and drowned. Moisture is also important to shedding in this species as animals kept only in dry sand retained shed resulting in reduced foraging success and occasionally death (Miller 1944). Rainfall in some areas where California legless lizards have been collected has been as low as three inches annually (Klauber 1932).

California legless lizards can tolerate a wide range of temperatures. The upper thermal limit is 40 degrees Celcius, above which animals quickly perish. Animals do not appear to have a specific lower limit as animals were able to survive temperatures of 4 degrees Celcius overnight, however it is likely that a realistic lower activity limit is 13 degrees Celcius (Miller 1944). Although these lizards can tolerate a wide variety of temperatures, they appear to avoid temperatures below 20 degrees Celcius and above 28 degrees Celcius in laboratory settings (Bury and Balgooyen 1976).

California legless lizards in southern California are primarily seen and collected in February, but have been collected in July and August. Activity is primarily fossorial, and these animals rarely travel above ground (Klauber 1932). As these animals can achieve extremely high densities (3,582 animals per 0.228 square meters) it is likely that they do not exhibit territoriality and have small home ranges (Kuhnz et al. 2005). Emigration rates are low for this species but individual animals actively create new burrow systems in the laboratory (Miller 1944).

California legless lizards are live bearing with litter size ranging from one to four offspring (Miller 1944). In one population the average litter size was 1.3 offspring with a maximum of two offspring per female (Goldberg and Miller 1985). Ovulation occurs in May, June or July with insemination occurring before, during and after ovulation (Miller 1944). Although the exact breeding time is unknown, it likely begins in early spring and continues into July. In a laboratory setting offspring are born in mid to late October with a gestation period of approximately four months (Goldberg and Miller 1985) Presumably sexual maturity is attained when the animals are two to three years old (Miller 1944). Females likely do not reproduce annually, however the exact cycle is not known (Goldberg and Miller 1985).

Threats Outside of National Forest System lands the primary threat to the California legless lizard is habitat loss due to habitat development. As this species is secretive, not all of its populations are documented, and many populations will likely disappear before being discovered. Parham and Papenfuss (2009) found that of three populations sampled in the Bakersfield region, two were extirpated by the conclusion of their study. California legless lizards are significantly impacted by bulldozing and plowing which alters and compacts the soil structure rendering it unusable to this species. Additional impacts are caused by livestock grazing, off-highway vehicle use, erosion, and the introduction of invasive habitat altering plants (Jennings and Hayes 1994). Predation is likely a major factor in survival of this species as most animals are found to have regenerated tails indicating previous predation attempts. Although these animals can autotomize their tail, they are still preyed upon by numerous other animals including: house cats (Felis domesticus), white-footed mice (Peromyscus maniculatus), loggerhead shrikes (Lanius ludovicianus), and common kingsnakes (Lampropeltis getula). Other predators which likely consume California legless lizards include domestic dogs (Canis domesticus), Norway rats (Rattus norvegicus), California ground squirrels (Spermophilis beecheyi), meadow mice (Microtus californicus), pocket gophers (Thomomys bottae), moles (Scapanus latimanus), crows (Corvus brachyrhychos), alligator lizards (Elgaria cooeruleus and E. multicarinatus), gopher snakes (Pituophis catenifer) and racers (Coluber sp.) (Miller 1944). Wildfire likely impacts this species, albeit not directly, due to the impacts caused by bulldozers used in firefighting efforts and the loss of leaf litter.

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Indirect Effects Alternative A: Implements the 2004 Sierra Nevada Framework and will provide for the protection of the riparian zone that supports this species.

Alternative B, C, and D: Reduction of uncharacteristic wildfires will help reduce the need for fire suppression, thus reducing the threats to this species. Removing invasive species will be a benefit to this species.

Determination Statement Sequoia NF: All alternatives of the Forest Plan may affect this species. The management framework will not trend it towards federal listing nor a loss of viability at the Forest Plan level. Alternative B, C and D will provide a beneficial effect to the species by restoring and maintaining riparian areas and water quality and quantity, as well as reducing the risk of catastrophic wildfire.

Inyo and Sierra NFs: The Forest Plan will have no effect on this species, nor will it trend towards federal listing, nor will there be a loss of viability at the Forest Plan. This species does not occur on these forests.

Literature Cited – California Legless Lizard Bury, R.B. and Balgooyen, T.G. 1976. Temperature selectivity in the legless lizard, Anniella pulchra. Copeia 1976(1): 152-155.

Fusari, M.H. 1985. Drinking of soil water by the California legless lizard, Anniella pulchra. Copeia 1985(4): 981-986.

Goldberg, S.R. and Miller, C.M. 1985. Reproduction of the silvery legless lizard, Anniella pulchra pulchra (Anniellidae), in southern California. The Southwestern Naturalist 30(4): 617-619.

Jennings, M. R. and Hayes, M. P. 1994. California legless lizard Anniella pulchra Gray 1852. In: Amphibian and Reptile Species of Special Concern in California. California Department of Fish and Game, Sacramento, California.

Klauber, L.M. 1932. Notes on the silvery footless lizard, Anniella pulchra. Copeia 1932(1) 4-6.

Kuhnz, L.A., Burton, R.K., Slattery, P.N. and Oakden, J.M. 2005. Microhabitats and population densities of California legless lizards, with comments on effectiveness of various techniques for estimating numbers of fossorial reptiles. Journal of Herpetology 39(3): 395-402.

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Miller, C.M. 1944. Ecologic relations and adaptations of the limbless lizards of the genus Anniella. Ecological Monographs 14(3): 271-289.

Morey, S. 2000. California legless lizard Anniella pulchra. In: California’s Wildlife. Vol. I. Amphibians and Reptiles. D. C. Zeiner, W. F. Laudenslayer Jr., K. E. Mayer, and M. White. California Department of Fish and Game, Sacramento, California.

Parham, J.F. and Papenfuss, T.J. 2009. High genetic diversity among fossorial lizard populations (Anniella pulchra) in a rapidly developing landscape (Central California). Conservation Genetics 10: 169- 176.

Pearse, D.E. and Pogson, G.H. 2000. Parellel evolution of the melanic form of the California legless lizard, Anniella pulchra, inferred from mitochondrial DNA sequence variation. Evolution 54: 1041-1046.

Stebbins, R.C. 2003. A Field Guide to Western Reptiles and Amphibians. Third edition. Houghton Mifflin, New York.

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Contributors  Catherine Yasuda, Formerly Wildlife Technician, Region 5  Dave Pattison, GIS mapping, Regional Office, Region 5  Emilie Lang, Forest Wildlife Biologist, Sequoia NF  Joel Grapentine, Database, Regional Office, Region 5  Joseph Furnish, Formerly Regional Aquatic Ecologist, Region 5  Kim Sorrini, Wildlife Biologist, Sierra NF  Leeann Murphy, Resource Staff Officer, Inyo NF  Linda Angerer, Wildlife Biologist, Mendocino NF  Lisa Sims, Forest Range Program Manager, Inyo NF  Michael Kellett, Formerly Regional Fish Program Leader, Region 5  Nina Hemphill, Forest Fisheries Biologist, Sequoia NF  Robin Galloway, District Biologist, Sequoia NF  Sean Hill, Wildlife Biologist, Above and Beyond Enterprise Unit  Stephanie Coppeto, Wildlife Biologist, Lake Tahoe Basin Management Unit

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