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The State of Natural and Cultural Resources in the Colorado River Ecosystem:

JUNE 30, 1999 DRAFT REPORT
Grand Canyon Monitoring and Research Center
Flagstaff, AZ 86001
Updated: 30 June 1999

Table of Contents

Endangered Species and Species of Concern

SPECIES OF SPECIAL CONCERN

ENDANGERED KANAB AMBERSNAIL

SUCCINEIDAE: Oxyloma haydeni kanabensis Pilsbry

Figure ES1.1:

Figure ES1.1: Kanab ambersnail habitat changes, March 1995-May 1999 at three stage elevations (Stevens et al. 1997b; draft data courtesy of V.J. Meretsky). Updated 23 June 1999.

Figure ES1.2:

Figure ES1.2: Kanab ambersnail estimated population size, March 1995-September 1998 (elevations (Stevens et al. 1997b; draft data courtesy of V.J. Meretsky). Updated 23 June 1999.

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Kanab Ambersnail Species Account

Distribution and Abundance

Kanab ambersnail (KAS; Succineidae: Oxyloma haydeni kanabensis Pilsbry 1948), is a federally endangered landsnail that was proposed for emergency listing (U.S. Fish and Wildlife Service 1991a, 1991b) and listed in 1992 (U.S. Fish and Wildlife Service 1992). Fossil Oxyloma shells have been recovered from sediments in Grand Gulch (lower San Juan River) that date to 9,200 years ago (Kerns 1993). Living KAS were first collected by J.H. Ferriss in 1909 near Kanab, Utah in seep vegetation (Ferriss 1910, Pilsbry and Ferriss 1911, Pilsbry 1948). This genus has a broad distribution (North America, Europe and South Africa), but the taxonomy has been based on internal and shell morphology, and is being revisited through molecular genetic techniques. Extant populations of KAS are described from: (1) Three Lakes, near Kanab Utah; (2) in seeps and springs along Kanab Creek, near Kanab, Utah; (3) at Vaseys Paradise, a spring at Colorado River Mile 31.5R, in Grand Canyon, Arizona (Spamer and Bogan 1993a, 1993b); and 4) populations on the east slope of the Canadian Rocky Mountains in Alberta (Harris and Hubricht 1982). The Three Lakes population occurs at several, small spring-fed ponds on cattail (Typha sp.; Clarke 1991). The site is privately-owned and the land owner is commercially developing the property. A recently discovered large metapopulation in Kanab Creek is believed to overlap the vicinity of the type locality (V. Meretsky, personal communication).

Vaseys Paradise is a fast-flowing, cool, dolomitic-type spring, with abundant wetland and phreatophyte vegetation, especially native crimson monkeyflower, sedge, smartweed, and poison ivy, and non-native watercress. Monkeyflower, sedge, smartweed and watercress are persistent aquatic wetland or hydrophytes (Kearney and Peebles 1960). KAS were first collected at Vaseys Paradise in 1991 (Blinn et al. 1992, Spamer and Bogan 1993), and an interagency team lead by Reclamation examined KAS ecology there from 1995 through 1997 (Kanab Ambersnail Interagency Work Group 1997a). Within Grand Canyon, KAS is apparently restricted to Vaseys Paradise: no KAS have been detected at more than 150 other Grand Canyon springs surveyed from 1991 through 1998. This suggests that the Vaseys Paradise KAS population, like many southwestern spring species, is a Pleistocene relict which has become restricted in distribution as Holocene climate dried out. Genetic dissimilarity with other O. haydeni populations in the Colorado River drainage further supports this contention (Miller et al. in press).

Vaseys Paradise is a popular water source and attraction site for Colorado River rafters; however access is limited by the dense cover of poison ivy (Toxicodendron rydbergii) and the nearly vertical terrain (Stevens et al. 1997b). Rematched historical photographs of Vaseys Paradise (e.g. Turner and Karpiscak 1980:58-59) reveal that vegetative cover has increased by about 40% at lower stage elevations since completion of Glen Canyon Dam. All vegetation below the approximate 90,000 cfs stage was scoured by annual pre-dam floods in normal years.

Stevens et al. (1997b) defined primary KAS habitat at Vaseys Paradise as that dominated by crimson monkeyflower (Mimulus cardinalis), non-native watercress (Nasturtium officinale), sedge (Carex aquatilis) and smartweed (Polygonum amphibium). Secondary, or marginal, habitat has been defined as patches of other riparian vegetation that are not dominated by these species. KAS is generally restricted to those species at Vaseys Paradise (Stevens et al. 1997b, Kanab Ambersnail Interagency Work Group 1997a,b, L. Stevens personal communication). KAS are rare or absent on other plant species and bare bedrock substrata at Vaseys Paradise.

Land surveys from 1995 through 1997 revealed rapid changes in vegetation cover over the growing season, with 11.2% to 16.1% of the estimated total primary habitat occurring below the 45,000 cfs stage in 1995, and 7.0-12.5% of the estimated total primary habitat occurring downslope from the 45,000 cfs stage from April 1996 through May 1999. The total estimated area of primary habitat was estimated to be 905.7 m2 (0.22 acres) in September 1994, equivalent to the area of secondary habitat, and the total vegetated area was 1811.4 m2 (0.44 acres) in June, 1995. Photogrammetric analyses suggest that the upper primary habitat area in November 1997 had decreased to approximately 720 m2 (L.E. Stevens, personal communication).

The accidental introduction of watercress at Vaseys Paradise provided KAS with an alternate host plant. KAS densities are generally higher on watercress than on the native host plants during the growing season (Kanab Ambersnail Interagency Work Group 1997b). Although watercress is an annual species, its life cycle at Vaseys Paradise is unpredictable. In part, this irregularity is due to the unithermal warm flows of the spring (ca. 16oC), which keep the microenvironment warm enough to prevent freezing during moderately cold winter months. Also, warm winters, such as 1995-1996, do not freeze watercress back, while cold winters (e.g., 1990) freeze and kill the plants. Warm spring flow and warm winters decouple the watercress life cycle from climate, and limit predictability of habitat conditions.

Demographic analyses based on size class distribution indicate that KAS is essentially an annual species, with much of the population maturing and reproducing in mid-summer (July and August), and most snails over-wintering as small size classes (Kanab Ambersnail Interagency Work Group 1997b). Loose, gelatinous egg masses were observed on the undersides of moist to wet live stems, on the roots of water-cress, and on dead or decadent stems of crimson monkey-flower in mid-summer of all years of study. No data on egg development or emergence success are presently available. In warm winters, such as that of 1995-96, KAS may emerge from dormancy early, and produce a double generation within one year (Kanab Ambersnail Interagency Work Group 1997a).

The other three populations of O. haydeni in northern Arizona and southern Utah appear to be genetically distinct from the Vaseys Paradise population (Miller et al., in press).

KAS at Vaseys Paradise are parasitized by the trematode flatworm, Leucochloridium cyanocittae, with 0.0% to 9.5% of the mature snails expressing sporocysts in mid-summer from 1995 through 1998 (Stevens et al. 1997b, Kanab Ambersnail Interagency Work Group 1997 a,b; V. Meretsky personal communication; Ka[plinski personal communication). Parasitized KAS are apparently able to continue to reproduce. Potential vertebrate predators of KAS at Vaseys Paradise include deer mice (Peromyscus crinitus and P. maniculatus), as well as rainbow trout (Oncorhynchus mykiss) in the stream mouth), resident common raven (Corvus corax) and canyon wren (Catherpes mexicanus), summer breeding Says and black phoebe (Sayornis sayi and S. nigricans), and winter resident American dipper (Cinclus mexicanus).

The total estimated Vaseys Paradise KAS population rose from 18,476 snails in March 1995 to more than 100,000 snails in September, 1995 as reproduction took place in middle to late summer (Stevens et al. 1997b; Table 1). This latter figure has been questioned on the basis of topographic survey accuracy: more recent population peaks have been 20,000-40,000 KAS. The proportion of the total estimated KAS population occurring below the 45,000 cfs stage was 3.3% in March, 11.3% in June, and 16.4% in September in 1995. Three years of population data and laboratory analyses indicate that the KAS population undergoes a substantial reduction through over-wintering mortality (Kanab Ambersnail Interagency Work Group 1997b).

The KAS population and habitat lying downslope from the 45,000 cfs stage was scoured in the BHBF in 1996 (Kanab Ambersnail Interagency Work Group 1997a; Table ES1.1). Habitat recovery was delayed in 1996 and 1997 because of high flows (20,000 to 28,000 cfs) that resulted from high reservoir forecasts and large summer monsoon floods on the Paria River, particularly in 1997. Recovery of habitat continued during the high flows of 1998 and 1999.

Analyses of oblique photographs taken in November of 1994-1997 indicate that no major compositional changes have occurred in the vegetation cover lying upslope from the 70,000 cfs stage; however, area changes may occur as climate and springflow change within and between years (Kanab Ambershail Interagency Work Group 1997). In October 1997, 101.22 m2 (12% of the estimated total primary habitat at Vaseys Paradise existed downslope from the 45,000 cfs stage. October 1997 population data indicate that an estimated 2,187 KAS exist downslope from the 45,000 cfs stage, 6.4% of the estimated total KAS existing at Vaseys Paradise (Kanab Ambersnail Interagency Work Group 1997b). Also, these data indicate that recovery of the Vaseys Paradise KAS habitat and population to pre-1996 BHBF conditions has required more than 2 full years. By late September 1998, preliminary estimates revealed 68.8 m2 of KAS habitat and 3,170 KAS downslope from the 45,000 cfs stage, and approximately 162.3 m2 of habitat with 9,405 KAS downslope from the approximate 60,000 cfs stage elevation.

The GCMRC estimates that approximately 81.47 m2 (11.8%) of the estimated total habitat will be inundated during a 45,000 cfs BHBF in the summer of 1999. If the above habitat assumptions are accurate, if KAS densities are equivalent across stage elevation (as suggested by the Kanab Ambersnail Interagency Work Group, 1997b), and if winter mortality is negligible and not different among stage zones, the April population survey data estimate that 3,170 KAS (7.3% of the 28 September 1998 estimated total population of 43,443 KAS) may be lost during a 1999 BHBF of 45,000 cfs. For reference, the 1996 BHBF removed 119.4 m2 of habitat and would have eliminated an estimated 2,126 KAS, had not 1,275 KAS been moved to higher stage elevations.

KAS and Planned Flooding

BHBF flows are of sufficient magnitude to alter habitat availability and KAS recolonization rates; however, the KAS population at Vaseys Paradise has survived and recovered from innumerable flows equal to or higher than BHBFs in the pre-dam era. No planned flood will be of sufficient magnitude to threaten the integrity of KAS as a population. The introduction of non-native watercress and the construction of Glen Canyon Dam have increased primary KAS habitat area at Vaseys Paradise by nearly 40%, and has undoubtedly substantially increased the snail population. Since 1963, the KAS population at Vaseys Paradise has survived seven flows of >45,000 cfs (i.e., 1965, 1980, 1983-1986, and 1996). Although incremental take from repeated high flows is a concern, KAS and its habitat require the >2 yr recovery period at Vaseys Paradise (Kanab Ambersnail Interagency Work Group 1997b), indicating that the KAS population has existed in a state of recovery from high flows (>45,000 cfs) for at least 16 of the past 36 years (>45% of post-dam time).

Not withstanding these considerations, the KAS habitat lying in the BHBF flood zone is likely to be adversely affected by flows of 45,000 cfs or more in 1999. Several points are relevant: 1) the Vaseys Paradise KAS population appears to be relatively large and self-sustaining; 2) approximately 40% of the present primary KAS habitat at Vaseys Paradise lies below the pre-dam 10-year flood stage of 125,000 cfs and is new, post-dam habitat; 3) the KAS population has survived numerous larger floods both before and after dam construction; 4) the estimated loss KAS habitat and population may exceed permitted levels of take if all habitat downslope from the 45,000 cfs stage is considered; and 5) the vegetation and the KAS population recovery after a 45,000 cfs flows requires >2 yr. In order for Reclamation to meet its commitments under the three previous biological opinions for operations of GCD, and prior to definition of successful second population establishment criteria, BHBFs in 1999 will be restricted to flows <45,000 cfs, and higher flows are to be avoided. Mitigation options include moving KAS in the BHBF flood zone to the Phoenix Zoo, or restocking the second population sites in Grand Canyon.

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HUMPBACK CHUB

CYPRINIDAE: Gila cypha Miller

Humpback Chub

Condition

Fig. ES2.1: Fig. ES2.1: Mean modeled body mass of 300 mm-long male humpback chub in the lower Little Colorado River during the spring spawn in May 1978-1998. Data from all years do not include spawning condition (pre-versus post-spawning condition). No gender-related were available from 1997-1999. Data from AGFD and USFWS, and calculations courtesy of V.J. Meretsky, Indiana University. Updated 23 June 1999.

Humpback Chub Species Account

Distribution and Abundance

The endangered humpback chub (HBC; Cyprinidae: Gila cypha) is an endemic fish species in the Colorado River basin (Valdez and Ryel 1997). The HBC was taxonomically described by Miller (1946), and was listed as an endangered species in 1968. Stream alteration, including flow modification, diversion for irrigation, channelization, and the introduction of non-native fish species, have been suggested as responsible for declining populations of HBC throughout the Colorado River basin (Valdez 1995). Five HBC populations remain in canyon-bound reaches of the upper Colorado River basin: Black Rocks (upper Colorado River), Westwater Canyon (upper Colorado River), Cataract Canyon, Desolation/Gray canyons (Green River) and in the Yampa River (Valdez and Williams 1993, Valdez and Ryel 1997).

The Grand Canyon supports the only successfully reproducing HBC population in the lower Colorado River basin (Kaeding and Zimmerman 1983, Valdez 1995, Valdez and Ryel 1997). Valdez and Ryel (1995) identified nine distinct aggregations in the mainstream Colorado River downstream from Glen Canyon Dam, including: 30-Mile, the Little Colorado River (LCR) confluence area, Lava/Chuar to Hance Rapids, Bright Angel Creek mouth, Shinumo Creek mouth, Stephens Aisle, Middle Granite Gorge, Havasu Creek mouth and Pumpkin Spring. From 3000 to 3500 adult HBC occupy the mainstream Colorado River, and these are largely concentrated within + 4.2 miles of the mouth of the Little Colorado River (Mile 61.5), the largest and only self-sustaining sub-population. The mainstream HBC in the LCR aggregation use the LCR for spawning, while other HBC appear to be resident in the LCR. The distribution of HBC in the mainstream has not changed over the past two decades (Valdez and Ryel 1997); however, HBC density may have declined in the LCR during the past decade (Douglas and Marsh 1996).

The life history and ecology of HBC in Grand Canyon has been intensively studied (Suttkus and Clemmer 1977, Kaeding and Zimmerman 1983, Carothers and Minckley 1981, Maddux et al., 1987, Gorman 1994, Valdez 1995, Arizona Game and Fish Department 1996, Douglas and Marsh 1996, Valdez and Ryel 1997, Stone 1999). Habitat use by HBC varied between age classes and by time of day in the Little Colorado River (Stone 1999). Young HBC in the Colorado River mainstream commonly use return current channels and other backwater habitats (Maddux et al., 1987, Arizona Game and Fish Department 1996, Valdez and Ryel 1997); however, HBC use of backwater habitats in Grand Canyon has been compromised by fluctuating flows and cold-stenothermic releases which reduce warming and create unstable conditions. In addition, backwater habitat area has been reduced and backwaters have aggraded through siltation under Interim Operations in Grand Canyon (McGuinn-Robbins 1997).

Young-of-year and subadult HBC in the Colorado River mainstream often use irregular shorelines as habitat, and adult HBC often occur in or near eddies (Valdez and Ryel 1995, 1997). Adult radio-tagged HBC demonstrated a consistent pattern of greater near-surface activity during the spawning season and at night, and day-night differences decreased during turbid flows (Valdez 1997).

Dietary analyses reveal HBC to be opportunistic feeders, selectively feeding on aquatic and terrestrial invertebrates (Valdez and Ryel 1995, 1997) and some cannibalism occurs (Stone 1999). HBC diet changes over the course of the year in response to food availability and turbidity-related decreases in benthic standing biomass over distance downstream from Glen Canyon Dam (Valdez and Ryel 1997). Non-native Gammarus lacustris occasionally comprise a large proportion of HBC diet, especially after high mainstream flow events (Valdez and Ryel 1995, 1997; Arizona Game and Fish Department 1996). Gammarus selectively feeds on epiphytes (i.e., diatoms) associated with Cladophora glomerata, the dominant alga particularly in the upper reaches, where clearwater conditions often prevail (Shannon et al. 1994).

Impact of High Flows

High flows, such as the 1996 BHBF, had little detectable effect on the movement patterns or distribution of adult HBC, and the 1996 BHBF did not appear to serve as a spawning cue (Valdez 1997). The increased drift associated with that BHBF resulted in an increase in Gammarus in HBC gut contents, an effect which is not surprising given the opportunistic foraging behavior of this species (e.g., Tyus and Minckley 1988). Given that this species evolved under the highly variable flow regimes that characterized the pre-dam Colorado River, it is unlikely that short-duration BHBF flows of 45,000 cfs would affect subadult or mature HBC.

BHBF's from May through July may affect the HBC population. The timing of high flow events may adversely affect larval and young HBC, through stress and displacement of young fish, depending on the spawning peak in any given year. High flows from January through March are believed to be unlikely to affect young HBC because high winter mortality apparently results in low populations of young fish during winter and spring. High flows that occur during the spawning and drift phase of the HBC life history cycle in the LCR may reduce annual survivorship and recruitment in the mainstream, and may flush refugial backwater habitats along the mainstream. Spring spawning activity from 1995 through 1998 appears to have resulted in rather normal levels of recruitment (Tuegel 1995; Grand Canyon Monitoring and Research Center 1997; T. Hoffnagle, personal communication), and the cold spring conditions in 1998 appeared to have depressed or extended spawning activity (Hoffnagle, personal communication). The extent of HBC spawning in 1999 is under analysis.

A <45,000 cfs BHBF in 1999 may affect, but is not likely to adversely affect subadult or adult HBC during any month between January and July, because HBC appear to be well-adapted to high flow events. Mid-winter or early spring planned floods may be preferable to late spring or summer high flow events for HBC. Reclamation continues to support the Service's recommendations regarding research and recovery efforts on this species, including analyses of winter mortality and establishment of a self-sustaining second population. For a proposed May through July BHBF, the Service suggests monitoring and research into HBC mortality in relation to ponding and predator responses at the LCR mouth, stress and displacement from mainstream shoreline and backwater habitats, and drift in the mainstream.

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ENDANGERED RAZORBACK SUCKER

CATOSTOMIDAE: Xyrauchen texanus

Fig. ES3.1:

Fig. ES3.1: Data not available, 23 June 1999.

Razorback Sucker Species Account

Distribution and Abundance

Razorback sucker (RBS; Catostomidae: Xyrauchen texanus) is a widely distributed, endemic, warm water Colorado River fish. RBS formerly occurred throughout the Colorado River, but has declined since 1930 with the regulation of the Colorado River (Dill 1944, Minckley 1991). The decline of RBS has been attributed to thermal regime changes, altered spawning habitat, blockage of migration routes, and introduction of non-native fish species, which have cumulatively resulted in wide-scale recruitment failure (Bestgen 1990, Minckley 1991). This species was listed as an endangered species by the U.S. Fish and Wildlife Service in 1991 (U.S. Fish and Wildlife Service 1991).

The largest RBS population in the Lower Colorado River Basin exists in Lake Mohave, where it was estimated to be approximately 60,000 fish in 1989 (Marsh and Minckley 1989). Other, smaller lower basin RBS populations occur in Lake Mead, downstream from Hoover Dam, and in Senator Wash Reservoir. In the Upper Colorado River Basin, RBS occur regularly in the upper Green and lower Yampa rivers, and individual RBS have been collected at rare intervals in the Colorado River near Grand Junction, Colorado, and in the major tributary arms of Lake Powell. RBS are long-lived (20 to 50 yr), but most wild-caught RBS are old individuals, and recruitment failure may lead to the rapid demise of this species (McCarthy and Minckley 1987, Minckley 1991). Experimental releases in the Upper Basin, and attempts to propagate RBS in Lower Basin reservoirs are encouraging, but the mainstream Colorado River populations continue to decline.

RBS are extremely rare in Grand Canyon. Recent observations include those of Carothers and Minckley (1981) who reported four RBS from the Paria River in 1978-1979; Maddux et al. (1988) reported one blind female RBS at Upper Bass (Colorado River Mile 107.5) in 1984; and Minckley (1991) reported records of three additional RBS captured in the lower Little Colorado River from 1989-1990. Putative hybrids between flannelmouth sucker (Catostomus latipinnis) and RBS have been reported from the Little Colorado River (Suttkus and Clemmer 1979, Carothers and Minckley 1981, Minckley 1991). RBS have not been observed since 1991 in this system.

Life Requisites

RBS are generally associated with calm river reaches, particularly man-made lakes (Tyus 1987); however, river spawning typically occurs in riffle habitats over gravel and cobble substrata (Mueller 1989). Larval RBS drift downstream from the spawning habitat, and concentrate in warm, low-velocity areas (e.g., flooded bottoms). These areas also support post-larval RBS, and channel and mid-stream river habitats floored by fine-grained alluvium are important to subsequent RBS life stages (Minckley 1983, Tyus and Karp 1989, Minckley 1991). Springtime concentrations of adult RBS have been noted in side-channels, off-channel impoundments, and in tributaries (Bestgen 1990, Minckley 1991), in temperatures of 22 to 25oC (Bulkley and Pimentel 1983); however, RBS occur in widely varying temperatures. RBS habitats in the Upper Colorado River Basin are ice-covered during winter, while the temperatures of mainstream habitats in the Lower Colorado River exceed 90oF (Dill 1944).

RBS diet varies by age class and habitat type, but few data are available on the diet of larval and juvenile RBS (Bestgen 1990). Larval RBS are known to feed on phytoplankton and zooplankton, and (in fluvial habitats) on chironomid larvae. Adult RBS in lentic habitats feed on benthic and planktonic algae and macroinvertebrates, while adult RBS in rivers feed primarily on benthic algae and invertebrates.

RBS spawn earlier in the season than do most other native, warm water Colorado River fish (Minckley 1973, 1991). Lake Mohave RBS spawn from November into May, with the peak of spawning activity between January and March when water temperatures were stable (50 to 54oF) or rising from 50 to 59oF (Bozek et al., 1984). In riverine situations in the Upper Basin, RBS begin spawning on the rising limb of the spring (April-May) hydrograph, and spawn for an extended period through the spring runoff. Although it occurred throughout the day, spawning activity is most intense at dusk.

RBS are susceptible to parasitic bacteria, protozoa and copepods. Minckley (1983) and others have reported a high incidence of blindness in one or both eyes; however the reasons for this condition are not clear (Bestgen 1990).

Impacts of BHBF(s) and Conclusions

The 1996 BHBF had no detectable effect on the remaining RBS population in the Colorado River in Grand Canyon. If RBS remain in this portion of the Colorado River, they are likely to be mature or senile fish, which survived comparable or higher mainstream flows in 1965, 1973, 1980, and 1983-1986, and possibly those of the late pre-dam era. These older fish are capable of finding suitable refugia, and the lack of recruitment of this species indicates that no young razorback sucker are likely to be in the system or at risk during any 1999 BHBF's. Because RBS spawn somewhat earlier than HBC, earlier (February-April) BHBF's may stimulate some additional RBS spawning activity; however, the rarity of this species precludes testing of such hypotheses.

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THREATENED BALD EAGLE

BUTEONIDAE: Haliaeetus leucocephalus

Fig. ES4.1: Data not presently available, 23 June 1999

Bald Eagle Species Account

Distribution and Abundance

The bald eagle (Accipitridae: Haliaeetus leucocephalus) has suffered population declines from habitat loss, mortality from shooting and poisoning, and reduced reproductive success from ingestion of contaminants (U.S. Fish and Wildlife Service 1983), and was recognized as a threatened and declining species in 1967. This species occurs throughout North America from Alaska to northern Mexico, and commonly breeds in the northern portion of its range (Stahlmaster 1987). Although bald eagles face numerous threats throughout the 48 states, they have recovered from dramatic population declines over the past several decades. Consequently, the U.S. Fish and Wildlife Service downlisted the bald eagle from endangered to threatened status (U.S. Fish and Wildlife Service 1995).

Wintering bald eagles were first observed to congregate in Grand Canyon in the early 1980's and the winter population there increased dramatically after 1985 (Brown et al., 1989, Brown and Stevens 1991, Brown and Stevens 1992). The wintering bald eagle population has been monitored since 1988, and it occurs primarily throughout the upper half of the Grand Canyon (in Marble Canyon) and on both Lakes Powell and Mead. Density of the Grand Canyon bald eagles during the winter peak (in late February and early March) ranged from 13 to 24 birds between Glen Canyon Dam and the LCR confluence from 1993 to 1995 (Sogge et al., 1995). A concentration of wintering bald eagles occurs in late February at the mouth of Nankoweap Creek, where bald eagles forage on spawning rainbow trout (Brown et al., 1989, Brown 1993). Bald eagle density there ranged from 6 in 1987 to 26 in 1990, and 18 bald eagles occurred at Nankoweap Creek in 1995 (Sogge et al., 1995). An estimated 13 bald eagles occurred simultaneously in Grand Canyon in January 1999 (L.E. Stevens, personal communication). Eagle density was correlated with trout density in the lower 0.5 km of Nankoweap Creek, and trout density was correlated with tributary stream water temperature (Leibfried and Montgomery 1993). Apparent territorial behavior, but no breeding activity, has been detected in Grand Canyon.

Life Requisites

Bald eagles are opportunistic feeders, preying on fish, waterfowl, rabbit and road-killed game (Stahlmaster 1987). Wintering bald eagles frequent rivers, reservoirs and lakes, including western reservoirs (Detrich 1987), and their distribution is dependent on prey availability, perch suitability, weather and human disturbance intensity (Ohmart and Sell 1980, Brown and Stevens 1997). Their wintering range extends from northern Mexico throughout the western United States.

At Nankoweap Creek in Grand Canyon, wintering bald eagles preferentially capture rainbow trout in the shallow creek, rather than in the mainstream where foraging success is low (Brown 1993, Sogge et al., 1995). Bald eagles at Nankoweap Creek prefer roosting and feeding areas that are relatively free of vegetation. The eagle population there consists of all age classes, with considerable piracy and other interactions between individuals (Brown and Leibfried 1990, Brown and Stevens 1991). The ease and relative safety of foraging in Nankoweap Creek affords wintering bald eagles at Nankoweap Creek the opportunity to accumulate energy reserves needed for their long, northward migration flights and initiation of nesting.

Bald eagle distribution in Glen and Grand canyons appears to be negatively related to human disturbance (Brown and Stevens 1997). Although bald eagles are widely known as opportunistic foragers, they are rare in the Glen Canyon and uppermost Grand Canyon reaches. This is surprising given that those reaches contain the most abundant aquatic foodbase and trout populations (Stevens et al. 1997c). Those reaches support the highest intensity of recreation and other human uses, and Brown and Stevens (1991) reported that bald eagles in Grand Canyon are extremely sensitive to human disturbance, often abandoning their foraging sites when human came within 0.5 km. For these reasons, Brown and Stevens (1997) concluded that human disturbance is responsible for the general rarity of bald eagles in the upper reaches.

Impacts of BHBF(s) and Conclusions

Wintering and migrant bald eagles have largely left the Grand Canyon region by late March (Stevens et al. 1997b). The few remaining eagles in April forage opportunistically and may continue to catch trout in the mainstream. The rainbow trout conclude their spawning run in Nankoweap Creek in April as water temperatures warm (Leibfried and Mongomery 1993), and remaining bald eagles no longer have access to that food source. Short-duration BHBF's in January through March may have slight adverse effect on bald eagle foraging at Nankoweap Creek, if the trout spawn is robust. High flows from April through July will have no adverse effect on the bald eagle population in the Colorado River in Grand Canyon because this migratory species is unlikely to be present.

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ENDANGERED PEREGRINE FALCON (PROPOSED FOR DELISTING)

FALCONIDAE: Falco peregrinus anatum

Fig. ES5.1: Data not presently available, 23 June 1999.

Peregrine Falcon Species Account

Distribution and Abundance

The peregrine falcon (Falconidae: Falco peregrinus anatum) is a federally listed endangered raptor, which declined dramatically as a result of the biological concentration of pesticide residues in prey species, and resulting eggshell thinning (U.S. Fish and Wildlife Service 1984). The population in the Rocky Mountain/ Southwest region declined from 180 pairs prior to 1975 to 55 pairs in 1983, largely as a result of DDT/DDE thinning of eggshells (U.S. Fish and Wildlife 1984).

The Grand Canyon peregrine population was thought to be low in the mid-1970's (Ellis and Monson 1989), but apparently increased dramatically in the 1980's, following recovery efforts by the U.S. Fish and Wildlife Service (1984; Glinski 1993). At present, the Grand Canyon supports the largest breeding population of peregrine falcons on a single land management unit in the coterminous United States (Brown et al. 1991a, 1992). Surveys for nesting peregrine falcons in 1988 and 1989 revealed 28 and 58 pairs, in 15% and 24% of the park, respectively. Habitat-based estimation of the potential number of peregrine falcons in Grand Canyon suggested that as many as 96 pair existed in Grand Canyon in 1989.

Life Requisites

Peregrine falcons feed on more than 40 species of birds and several small mammals (Porter and White 1973, Stevens et al. in press). Hunting areas included marshes or narrow tongues of streamside vegetation, and peregrine falcons may forage up to 17 miles from nest sites. Peregrine falcon diet at nest sites in national parks in southern Utah included small and medium-sized birds, especially including white-throated swifts, large shorebirds and Clark's nutcracker (Burnham 1987).

In Grand Canyon, peregrine falcons feed on waterfowl, swifts, swallows and bats (Brown 1991a, Stevens et al. in press), many of which feed on invertebrate species (especially Diptera) that emerge out of the Colorado River (Stevens et al. 1997c). Therefore, dam operations that influence aquatic macro-invertebrate populations exert, at most, only indirect impacts on peregrine falcons.

Peregrine falcons breed up to 3,130 m elevation, typically on ledges on steep cliff faces (U.S. Fish and Wildlife Service 1984). The mean distance between nest sites along the South Rim of Grand Canyon varied from 3.5 to 5.0 linear miles, with minimum distances of 1.8 linear miles (Brown 1991a, 1992). The breeding season in Grand Canyon extends from February to July.

The primary reason for the national decline of the peregrine falcon population has been eggshell thinning from DDE and other environmental contaminants, which are biologically concentrated through the food chain (U.S. Fish and Wildlife Service 1984). DDE sources to peregrine falcons are derived from their prey, many of which are migratory insectivores. Burnham (1987) reported that swifts, shorebirds, and other migratory insectivores contained 5.8 ppm DDE (wet weight), while mean DDE levels in granivorous migrants, such as grosbeaks and mourning doves, was only 0.14 ppm DDE. Peregrine eggshells from southern Utah parks from 1985 to 1987 were 21% thinner than those from the pre-DDE era, indicating poor viability of eggs (Burnham 1987). Brown (1991b) reported that peregrine falcon eggshells from the Grand Canyon in 1988 were 11.4% to 12.7% thinner than pre-DDE controls.

In addition to pesticide concentrations, competition with other raptors has been considered as a possible cause of peregrine falcon population declines; however, Porter and White (1973) examined peregrine and prairie falcon interactions and concluded that competition was not important.

Impacts of BHBF(s)

Most wintering waterfowl on which peregrine falcons feed will have migrated from Grand Canyon by late March; however, mallard and late migrating gadwall and American widgeon are still likely to be common (Stevens et al. 1997a, in press). Springtime food sources (swifts, swallows and bats) should be present in large numbers at that time of year (Stevens et al. in press), and are only indirectly influenced by dam operations. Therefore, peregrine falcons will not lack food resources during the proposed high release. BHBFs at any time of the year will have no effect on peregrine falcons in the Colorado River downstream from Glen Canyon Dam.

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ENDANGERED SOUTHWESTERN WILLOW FLYCATCHER

TYRANNIDAE: Empidonax trailli extimus

Figure ES6.1:

Figure ES6.1: The number of breeding southwestern willow flycatcher pairs detected along the Colorado River corridor in Grand Canyon from Mile 46-72 from 1982-1999. No surveys were conducted from 1988-1990 (Sogge 1998; 1998 and 1999 data courtesy of J. Spence, GLCA, unpublished). Updated 23 June 1999.

Figure ES6.2:

Figure ES6.2: The number of willow flycatcher nests detected along the Colorado River corridor in Grand Canyon from Mile 46-72 from 1982-1999. Lighter areas represent known renesting attempts (following failed previous nests) within the same territory. No surveys were conducted from 1988-1990 (Sogge 1998; 1998 data courtesy of J. Spence, GLCA). No 1999 data were available at the time of report preparation. Updated 23 June 1999.

Southwestern Willow Flycatcher Species Account

Distribution and Abundance

The southwestern willow flycatcher (SWWF; Tyrannidae: Empidonax trailii extimus) is a Neotropical migrant. Overall, the willow flycatcher species has a broad breeding range, extending from Nova Scotia to British Columbia and south to Baja California. The SWWF is an obligate riparian insectivore (Hunter et al., 1987), preferring habitat near open water (Gorski 1969; Sogge 1995). The historic breeding range of the SWWF includes Arizona, New Mexico, southern California, and southern portions of Nevada, Utah, and perhaps southwestern Colorado, and extends east into western Texas (U.S. Fish and Wildlife Service 1993). It probably winters from Mexico to Panama, with historical accounts from Colombia (Phillips 1948). The SWWF is distinguished from other subspecies by distribution, morphology and color, nesting ecology, but not by song dialect (Phillips 1948, Aldrich 1953, King 1955, Sogge 1995).

The regional SWWF population has declined over the past 50 years, corresponding with loss and modification of riparian habitats (Phillips 1948). Southwestern riparian ecosystems support a rich avian fauna (Johnson and Haight 1987) and habitat changes have resulted in reduction or extirpation of many avian species (Hunter et al., 1987). Modification and fragmentation of these systems through development and livestock grazing have precipitated devastating changes to SWWF populations. Destruction of native willow/cottonwood vegetation has provided opportunity for invasion by non-native plant species, notably saltcedar. Habitat fragmentation and modification has been beneficial to some southwestern avian species, especially cowbirds (Molothrus spp.), which parasitize SWWF nests, contributing to the precipitous population declines of SWWF (Brown 1994, Johnson and Sogge 1995, Sogge et al. 1995). SWWF habitat loss in Central and South America has also undoubtedly contributed to recent SWWF population declines, although little information is available.

The SWWF has been extirpated from much of its range (Hunter et al. 1987). Population reduction since 1950 was so dramatic that it was proposed (U.S. Fish and Wildlife Service 1992) and listed, with critical habitat, under the Endangered Species Act, on July 23, 1993 (U.S. Fish and Wildlife Service 1993). The SWWF is more rare than most other currently listed avian species (Unitt 1987). An estimated 300-500 breeding pairs remain in the United States, including 115 pairs in California and approximately 100 pairs in New Mexico (U.S. Fish and Wildlife Service 1993). Limited information exists for Colorado, Utah, Nevada, and Texas. It has been given special protection status by the Game and Fish Departments in Arizona, New Mexico and California.

Arizona has experienced the sharpest decline in SWWF numbers. SWWF formerly bred throughout the state at high and low elevations (Paradzick et al. 1999). For example, a 1931 breeding record exists from the south rim of the Grand Canyon (Brown et al., 1984), indicating that this taxon bred at high elevations, even at the northern edge of its range. By 1987, the State population was estimated at less than 25 pairs (Unitt 1987; U.S. Fish and Wildlife Service 1993), but much habitat was not surveyed. At least 52 territories or active nests were reported during extensive surveys in 1993 in Arizona (Muiznieks et al. 1994), and at least 62 active nests were located during a more thorough inventory in 1994 (Sferra et al. 1995). In Arizona, there were approximately 113 SWWF pairs in 1996 (Sferra et al. 1997), and in 1998 250 nesting attempts were detected in 34 drainages, with 53.0% nest success.

From 1974 through 1996 the Grand Canyon population was detected between Colorado River miles 47 and 71 (Unitt 1987, Sogge et al. 1995, 1997). In its recent proposal the Service included the Colorado River from River Mile 39 to River Mile 71.5 as critical habitat U.S. Fish and Wildlife Service 1993), and stipulated in a subsequent final rule that defines such habitat as that within 100 meters of the edge of areas with surface water during the May to September breeding season and within 100 meters of areas where such surface water no longer exists owing to habitat degradation but may be recovered with habitat rehabilitation (U.S. Fish and Wildlife Service 1997) . The boundary of this area in Grand Canyon includes the main Colorado River channel and associated side channels, backwaters, pools and marshes.

SWWF were common in Glen Canyon and the lower San Juan River prior to impoundment by Glen Canyon Dam (Woodbury and Russell 1945, Behle and Higgins 1959). This area was inundated by Lake Powell and no singing male SWWF were detected from 1991-1998 in the Glen Canyon reach below the dam, however, a singing individual was detected in 1999 (Brown 1991a; J. Spence, GLCA, personal commun.). SWWF were rather commonly reported along the pre-dam Colorado River at Lees Ferry, with records at Lees Ferry in 1909, 1933, 1935, and 1961, and near Lava Canyon in 1931 and near the Little Colorado River confluence in 1953 (reviewed by Sogge et al. 1997); however, the pre-dam distribution of SWWF in Marble Canyon and through Grand Canyon is poorly known. Carothers and Sharber (1976) reported only one pair of SWWF in Grand Canyon in the early 1970's surveys. Brown (1988) noted a brief population increase in the Grand Canyon from two in 1982, to a maximum of 11 (two in Cardenas Marsh), with a subsequent decline to seven in 1987. Brown (1991a) detected two pairs in 1991, with nests located at River Mile 50.7 and at River Mile 71.1 (Cardenas Marsh). Surveys in 1992 detected seven SWWF, three unpaired males and two breeding pairs in Cardenas Marsh (Sogge et al. 1995a). A total of five SWWF were detected in Grand Canyon in 1995: three territorial but non-breeding males and one breeding pair that fledged a single young (Sogge et al. 1995a). The unpaired male SWWF established territories between Colorado River miles 50.5 and 65.3, and the breeding pair nested at mile 50.5. In 1996 Sogge et al (1997) reported three singing SWWF, but only one successfully breeding pair along the Colorado River in upper Grand Canyon. The single pair apparently fledged two young. In 1997, the single nest in upper Grand Canyon was parasitized by brown-headed cowbirds and failed. A single SWWF nest near mile 265 in 1997 produced two young (Grand Canyon Monitoring and Research Center 1997). In both 1997 and 1998 SWWF failed to nest successfully in upper Grand Canyon because of cowbird brood parasitism (M. Sogge, U.S. Geological Survey Biological Resources Division, personal communication). The single nesting pair of SWWF at Mile 50.5L in upper Grand Canyon failed to produce young successfully in 1998 (N. Brown, personal communication). A pair has established a territory there in 1999, but nesting activity has not yet been reported. Other 1996-1998 reports of SWWF breeding in the lower Colorado River basin have stimulated additional research there.

The Service's 1996 B.O. on the BHBF defined several measures to mitigate impacts on the SWWF in Grand Canyon. Stevens et al. (1996) studied habitat changes at four historic SWWF nest sites in Grand Canyon. Fluvial marshes associated with these sites were dominated by common reed, horsetail and cattail. SWWF research activities associated with that BHBF included verifying stage-to-discharge relations, quantifying flow depth and velocity at nest sites, and determining nest site and foraging habitat structure, litter/understory characteristics, and nesting success.

The 1996 BHBF impacts on Grand Canyon SWWF habitat were reported by Stevens et al. (1996). Nest stand vegetation impacts were nominal: two stands were slightly scoured, and three sites sustained a slight reduction in ground cover and/or branch abundance at <0.6 m above the ground; however, no reduction in branch abundance or alteration of stand composition occurred, and the BHBF did not inundate the bases of any historic nest trees. Impacts on marsh foraging habitats were more severe, with decreases in area of 1% to >72%. Two of four SWWF sites regained vegetated area during the summer of 1996, while two other marshes sustained slight additional losses in cover through the 1996 growing season. The 50.05L marsh has not recovered appreciably since the 1996 BHBF (Stevens personal communication).

Life Requisites

SWWF arrive in the Grand Canyon area in mid-May, but may be confused with another subspecies, the more common E.t. brewsteri, which migrates through to more northern breeding grounds (Aldrich 1951; Unitt 1987). E.t. brewsteri sings during migration, making sub-specific distinctions difficult until mid-June (Brown 1991b). Males arrive earlier than females and establish territories. The characteristic territorial song is a fitz-bew, most frequently heard in the morning before 10 AM (Tibbitts et al., 1994).

SWWF are highly territorial. Nest building begins in May after breeding territories are established. The nest is placed in a fork or horizontal branch 1-5 meters above ground (Tibbetts et al. 1994). A clutch of three or four eggs is laid from late May through July (Unitt 1987), but in Grand Canyon two or three eggs (usually three) are usually laid (Sogge 1995). Breeding extends through July and singing ceases at the end of the breeding season.

After a 12-14 day incubation, nestlings spend 12 or 13 days in the nest before fledging (Brown 1988; Tibbetts et al., 1994). The breeding season (eggs or young in nest) in Grand Canyon extends from early June to mid-July, but may extend into August. One clutch is typical, however re-nesting has been known to occur if the initial nest is destroyed or parasitized (Brown 1988).

Riparian modification, destruction and fragmentation provided new foraging habitat for brown-headed cowbirds (Molothrus ater) and populations of brown-headed cowbirds continue to expand (Hanka 1985, Harris 1991). Brood parasitism is currently the greatest threat to SWWF and probably many other Neotropical migrants as well (Bohning-Gaese et al., 1993; Sogge et al., 1995). Over half the nests in Brown's study (1988) contained brown-headed cowbird eggs. Cowbirds may remove prey eggs, their eggs hatch earlier, and the larger nestlings are more competitive in the nest. Cowbirds fledged from Sierra Nevada SWWF nests while SWWF nestlings died shortly after hatching (Flett and Sanders 1987). Brown-headed cowbirds occur extensively around mule corrals on the rim of the canyon and travel down to the Colorado River.

SWWF may remove cowbird eggs or, more commonly, abandon the nest if the parasite's eggs are deposited. The second nesting attempt is energetically expensive, requiring a new nest to be built (Sogge 1995), although Brown (1988) noted that a SWWF pair covered a cowbird egg with fresh nesting material and laid a new clutch. The second nest, already at a temporal disadvantage, is often parasitized as well. Cowbird parasitism could be largely responsible for the absence of SWWF in otherwise suitable habitat in the Grand Canyon (Unitt 1987). Bronzed cowbirds (Molothrus aenus) have recently been reported colonizing the Grand Canyon and represent another threat (Sogge 1995).

The SWWF in Grand Canyon occupy sites with average vegetation canopy height and density (Brown and Trossett 1989). SWWF commonly breed and forage in dense, often multistoried, riparian vegetation near surface water or moist soil (Whitmore 1977, Sferra et al., 1995), along low gradient streams (Sogge 1995). Nesting in Grand Canyon typically occurs in non-native Tamarix approximately 4-7 m tall (13-23 feet), with a dense volume of foliage 0-4 m from the ground (Tibbetts et al., 1994). SWWF commonly and preferentially nest in saltcedar in upper Grand Canyon (Brown 1988), and nested in saltcedar in Glen Canyon before completion of the Glen Canyon Dam (Behle and Higgins 1959). Arizona SWWF preferentially nest in Tamarix: 194 of 203 nests detected in 1998 were situated in Tamarix (Paradzick et al. 1999). Although habitat is not limiting in Grand Canyon (Brown and Trossett 1989), required patch size is not known. The 1997 and 1998 nesting records from lower Grand Canyon demonstrates that this species can colonize new habitat; however, that habitat is influenced by Lower Basin Lake Mead management and is not within the purview of this report.

Proximity to water is necessary and is correlated with food supplies. Little is known of SWWF food preferences but it is probably a generalist feeder. They typically flycatch (sally) from conspicuous perches, but also hover and glean insects from foliage (Stevens personal communication). SWWF also forage on sandbars, backwaters, and at the waters edge in the Grand Canyon (Tibbetts et al., 1994).

SWWF return to wintering grounds in August and September (Brown 1991b), but neither migration routes nor wintering areas are well known. Birds sing and perhaps defend foraging territories in Central America during winter, and winter movement may be tied to water availability (Gorski 1969). Threats to SWWF on the wintering grounds are poorly documented, but habitat losses in Latin America may be a major factor in the decline of this species.

Impact of BHBF(s)

A BHBF between May and July may affect the southwestern willow flycatcher and its critical habitat along the Colorado River in Glen, Marble and upper Grand Canyons. Nesting sites for SWWF would not be affected because SWWF nest several meters up in tamarisk trees that stand at or above the 45,000 cfs stage. In upper Grand Canyon, SWWF generally nest in Tamarix trees and the bases of the nest trees typically lie at or above the 45,000 cfs stage. The Tamarix stands in which SWWF nest are unlikely to sustain direct damage from BHBF(s). Stevens and Waring (1988) demonstrated that Tamarix is exceptionally tolerant of flooding in the Grand Canyon, persisting through many weeks of inundation. The Tamarix trees in which the SWWF presently nest survived the >92,600 cfs flows of 1983 as well as the 1996 BHBF (Stevens et al. 1996), and are therefore unlikely to be scoured by one or more brief, <45,000 cfs BHBF's in 1999.

Although the Tamarix stands in SWWF nest are not likely to be affected by a ca. 45,000 cfs flow, the wetlands and low-lying areas near SWWF nesting habitats and in which they occasionally forage, may be affected by BHBFs. Impacts to associated wetlands ranged from 1% to >72% from the 1996 BHBF, and impacts on those sites persisted through the 1996 growing season (Stevens et al. 1996). Although those habitats were strongly affected by the 1996 flood, actual impacts on SWWF food resources remain undocumented. It is unlikely that the 1996 BHBF affected SWWF foraging, but impacts are impossible to document with so few birds to study.

Although little is known of SWWF food preferences, it is probably a generalist feeder on invertebrates. It typically hovers and gleans insects from foliage, or flycatches from conspicuous perches (Stevens, personal communication). SWWF also forage on sandbars, backwaters, and at the water's edge in the Grand Canyon (Tibbetts et al., 1994). SWWFs likely forage on both adult aquatic flying invertebrates, and terrestrial (non-aquatic) flying invertebrates. Although aquatic species could be impacted by the BHBF, populations of terrestrial flying invertebrates are unlikely to be affected by the flow, and any that are affected are likely to recover promptly after the event. Stevens (1985) reported that riparian invertebrate populations increased rapidly following a flow comparable to a BHBF in 1980 (U.S. Bureau of Reclamation 1990), a condition which would seem to provide required food needs for any SWWF present during or following a BHBF. Stevens (1985) also documented that tamarisk and willow, the dominant shrub/tree species in the Grand Canyon SWWF nest stands, support abundant invertebrate populations.

SWWF probably primarily forage on adult, terrestrial (non-aquatic) flying invertebrates, populations which are unlikely to be affected by a brief 45,000 cfs BHBF, and which are likely to recover promptly after the event. Stevens (1985) reported that riparian invertebrate populations increased rapidly following a flow comparable to a BHBF in 1980 (U.S. Bureau of Reclamation 1990). Invertebrate populations may require a longer recovery period after higher stage BHBFs.

The SWWF habitat and population in lower Grand Canyon is influenced by Lower Colorado River basin management of Lake Mead, and is not part of this Reclamation Office's purview. Even if that section of the river corridor is considered, present data indicate no impacts of a BHBF on SWWF habitat there. Results of Hualapai Indian Tribal analyses on riparian resources in lower Grand Canyon indicate that the impacts of the 1996 BHBF extended no farther than Mile 255 (Christensen 1997). This point lies approximately 10 miles upstream from recently reported SWWF nesting areas on upper Lake Mead. Therefore, this, and the shorter duration of the proposed 1998 BHBF, suggest that a planned high flow should have no effect on potential SWWF habitat in lower Grand Canyon .

In conclusion, BHBF(s) in 1999 may affect the SWWF. Foraging habitat of SWWF may be adversely affected, but flows >45,000 cfs may have an impact on nest trees. High flows are likely to continue to reduce marsh areas associated with nest site stands, while higher stage BHBF's may eliminate those marshes (Stevens et al. 1995). Impacts on SWWF territories in lower Grand Canyon are primarily a function of reservoir stage, and therefore not the purview of the AMWG. At any rate, SWWF are sufficiently rare that high flows impacts on their foraging success are unlikely to be statistically discernable.

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