(ISSN 0892-1016) The Journal OF Raptor Research Volume 33 September 1999 Number 3 Contents Natal and Breeding Dispersal in Barn Owls. Carl d. Marti 181 Diets of Breeding Peregrine and Lanner Falcons in South Africa. Andrew r. Jenkins and Graham M. Avery 190 Nest Dispersion, Diet, and Breeding Success of Black Kites (Milvus MIGRANS) IN THE ITALIAN PrE-AlPS. Fabrizio Sergio and Alberto Boto 207 Status of Nesting Bald Eagles in Arizona. Daniel e. Driscoll, Ronald e. Jackman, w. Grainger Hunt, Greg L. Beatty, James T. Driscoll, Richard L. Glinski, Thomas A. Gatz and Robert I. Mesta 218 The Influence of Weapons-Testing Noise on Bald Eagle Behavior. Bryan T. Brown, G. Scott Mills, Cornelius Powels, William A. Russell, Glenn D. Therres and James J. Pottie 227 Effects of Recreational Trails on Wintering Diurnal Raptors Along Riparian Corridors in a Colorado Grassland. Robert j. Fletcher, jr., Shawn t. McKinney and Carl E. Bock 233 Short Communications Northern Goshawks Nesting on a Private Industrial Forest in Western Washington. Thomas Bosakowski, Bruce McCullough, Frank J. Lapsansky and Martin E. Vaughn 240 Notes on Eggs of the Bicolored Hawk Accipiter bicolor. Russell Thorstrom and Lloyd F. Kiff 244 Nesting Ecology of Red-shouldered and Red-tailed Hawks in Georgia. Christopher E. Moorman, Doug L. Howell and Brian R. Chapman 248 The Red Kite (Milvus milvus) Reintroduction Project: Modeung the Impact of Translocating Kite Young Within England. Ian Carter, Mick McQuaid, Nigel Snell and Peter Stevens 251 First Recorded Polygynous Mating in the Red Kite (Milvus milvus) . Heiii Van Kleef and Javier Bustamante 254 Mites in Greater Spotted Eagle Nests. Dariusz J. Gwiazdowicz, Tadeusz Mizera and Maciej Skorupski 257 Food Habits oe the Rufous-legged Owl (Strix rufipes) in the Mediterranean Sclerophyllous Forest of Central Chile. Ivan Diaz 260 Food Habits of the Cinereous Harrier ( Circus cinereus) in the Araucania, Southern Chile. Ricardo A. Figueroa R. and Ema S. Corales S 264 Striated Caracara (Phalcoboenus austraus) Population at Staten and Ano Nuevo Islands. Esteban Frere, Alejandro Travaini, Anibal Parera and Adrian Shciavini 268 Letters 270 Book Review. Edited by Jeffrey S. Marks 272 Erratum 273 The Raptor Research Foundation, Inc. gratefully acknowledges a grant and logistical support provided by Boise State University to assist in the publication of the journal. THE JOURNAL OF RAPTOR RESEARCH A QUARTERLY PUBLICATION OF THE RAPTOR RESEARCH FOUNDATION, INC. VoL. 33 September 1999 No. 3 J. Raptor Res. 33(3):181-189 © 1999 The Raptor Research Foundation, Inc. NATAL AND BREEDING DISPERSAL IN BARN OWLS Carl D. Marti^ Department of Zoology, Weber State University, Ogden, UT 84408-2505 US. A. Abstract. — I studied dispersal of the Barn Owl {Tyto alba) in northern Utah from 1977—96. Based on 144 recoveries of 2085 banded nestlings, the average dispersal distance was 102.9 km (median = 60 km, range = 0—1267 km), occurred in most compass directions from natal sites, but was not random with mountains, deserts, and the Great Salt Lake altering dispersal routes. Dispersal distance was not corre- lated with severity of winter weather nor population density. Among owls banded as nestlings and re- captured as breeders, females {N = 48) moved significantly farther {ic = 61.4 km, median = 57.5 km, range = 0-160 km) than males {N = 34, x = 35.7 km, median = 14.7 km, range = 0.8-120 km, P = 0.015). Turnover of breeders at nest sites resulted mostly from individuals dispersing into the study area. Only 19 (of at least 500) breeders moved from one breeding site to another. The mean distance moved between breeding sites of 2.3 km (median = 2.25 km) was not significantly different between males and females {P = 0.9), but more females (16) than males (3) made these moves. Eight of the adults that shifted breeding sites did so in the same year either after a failed first attempt (2) or to produce a second brood (6). The remainder changed nest sites in subsequent years. Key Words: Bam Owh, Tyto alba; breeding dispersal, long-term study; natal dispersal; Utah. Dispersibn natal y reproductiva de Tyto alba Resumen. — Estudie la dispersion de Tyto alba en el norte de Utah desde 1977—96. Con base en 144 recapturas de 2085 pichones anillados, encontre que la distancia de dispersion fue de 102.9 km (media = 60 km, range = 0-1267 km), ocurridas en todas las direcciones desde el sitio de nacimiento. Esta situacion no ocurrio al azar en montanas, desiertos y el Great Salt Lake los cuales alteraron las rutas de dispersion. La distancia de dispersion no estuvo correlacionada con la severidad del clima invernal, ni con la densidad poblacional. Entre las lechuzas anilladas como pichones y recapturadas como re- productores, las hembras (N = 48) se movilizaron significativamente mas lejos {x = 61.4 km, media = 57.5 km, range = 0-160 km) que los machos (N = 34, x = 35.7 km, media 14.7 km, range = 0.8-120 km, P = 0.015). El regreso de los reproductores a los sitios de los nidos, fue el resultado de individuos dispersados dentro del area de estudio. Solo 19 (de por lo menos 500) reproductores se movilizaron de un sitio de reproduccion a otro. La distancia promedio recorrida entre los sitios de reproduccion fue de 2.3 km (media = 2.25 km). Esta distancia no fue significativamente diferente entre machos y hembras (P = 0.9). Mas hembras (16) que machos (3) hicieron estos movimientos. Ocho de los adultos que cambiaron sus sitios de reproduccion lo hicieron en el mismo ano despues de fracasar en un primer intento (2) o para producir una segunda nidada (6). El resto cambio el sitio del nido en los anos subsecuentes. [Traduccion de Cesar Marquez] The Barn Owl (Tyto alba) is among the most widespread of land birds, and although some as- ^ Present address: Boise State University, Raptor Research Center, Boise, ID 83725 U.S.A. pects of its biology closely resemble other owls (e.g., trophic biology), other attributes are striking- ly different. Among the important disparities are aspects of the Barn Owl’s reproductive biology and life-history (Marti 1997). Here, I show that dis- 181 182 Marti VoL. 33, No. 3 persal in Barn Owls conforms with the species’ r- selected life-history strategy (reproduction at an early age, short reproductive life, high reproduc- tive output, and an ability to find new resources — sometimes at great distances — through natal dis- persal) , but also fits some patterns of dispersal that are widespread in other birds. Dispersal is a very important but poorly under- stood element of population biology (Begon et al. 1990). Dispersal may be either natal — the one-way movement by an individual from its birthplace to a breeding (or potential breeding) site, or breed- ing — the movement by adults between breeding sites. Natal dispersal usually covers greater distanc- es than breeding dispersal (Greenwood and Har- vey 1982). Advantages attributed to natal dispersal include reducing the chance of inbreeding, reduc- ing competition, and extending the range (Green- wood 1983, Swingland 1983). In many bird species, dispersal patterns differ between adults and juve- niles and between males and females (Greenwood 1983, Greenwood and Harvey 1982). Relatively few studies of dispersal have been con- ducted on raptors. See for example, Newton (1979) and references within, Newton (1986) and Ferrer (1993) for European diurnal raptors, and Korpimaki et al. (1987), Korpimaki (1988), Kor- pimaki and Lagerstrom (1988), and Coles and Pet- ty (1997) for European owls. In North America, see Jacobs (1995), Woodbridge et al.(1995), Steenhof et al. 1984, and Miller and Smallwood (1997) for diurnal raptors, and VanCamp and Henny (1975), Adamcik and Keith (1978), Marks (1985), Bull et al. (1988), Belthoff and Ritchison (1989), Ganey et al. (1998), Gehlbach (1994), and Arsenault et al. (1997) for owls. Dispersal in Barn Owls has been studied in North America (Stewart 1952), Europe (Frylestam 1972, Schonfeld 1974, Glutz von Blotzheim 1979, Bairlein 1985, Baudvin 1986, Chanson et al. 1988, Taylor 1994, Martinez and Lopez 1995), and to a very minor extent in Australia (Purchase 1972). Only Taylor (1994) presented data on both the dis- persal of nestlings to breeding sites and move- ments of adults between nest sites. Previously, I documented the reproductive pat- tern (Marti 1994) and lifetime reproductive suc- cess (Marti 1997) in a Barn Owl population breed- ing close to the northern limit of its range. Here, I present dispersal patterns in the same popula- tion, test whether sex and age differences in dis- Figure 1. Location and topographic features of Barn Owl study area in northern Utah. persal occurred and look for support that dispersal reduces inbreeding. Study Area and Methods The study area was a narrow (12-25 km wide, 500 km^) valley lying between the Wasatch Mountains and the Great Salt Lake in Box Elder, Weber, and Davis counties of northcentral Utah (Fig. 1) that is close to the Barn Owl’s northern range limit in the Intermountain Region (Marti 1992). The area was shrubsteppe desert but that community has been entirely supplanted by irrigated ag- riculture and urban development. Hot dry summers and cold winters characterize the region; mean temperatures for July and January are 23.9°C and — 3.5°C, respectively. Barn Owl nesting habitat is limited and disjunct in this area; most Barn Owls nest in lower elevation valleys where irrigated agriculture occurs. Rugged mountains and high elevation valleys immediately east of the study area were unsuitable Barn Owl habitat, and, likewise, the Great Salt Lake and alkali deserts to the west of the study area offered little habitat for Barn Owls. See Marti (1994) for more details on the study area and owl nest sites. Most of the Barn Owls on my study area nested in nest boxes (Marti et al. 1979). From 1977-96, I visited these nest boxes year-round at least once per month. I made additional visits as needed to band and color mark nest- lings and adults with a standard USGS aluminum band and a combination of colored plastic bands unique to each bird (two bands per leg) permitting identification of individuals without having to recapture them. Few oth- er suitable nest sites existed on the study area, but owls occasionally nested in buildings and hay stacks. These were often reported to me by farmers or by owners of buildings having various problems caused by the nesting owls. Thus, I was able to document nesting in these sites as well. I attempted to capture all breeding owls each year to determine their identity, age, and movements. Most fe- males and some males were caught by hand in nest boxes but, because males were less often found in nest boxes, I sometimes used nest-box traps to capture them (Sau- September 1999 Barn Owl Dispersal 183 60 ■ 0-40 81-120 161-200 241-260 41-80 121-160 201-240 261 + Dispersal distance, km Figure 2. Dispersal distances in Barn Owls banded as nestlings in northern Utah. rola 1987). For breeding owls not banded as nestlings, age was determined by wing-molt pattern. Barn Owls do not molt any primaries until 13 months of age (P. Bloom pers. comm., Lenton 1984, Taylor 1993). Thus, in the spring, breeding owls with one generation of primaries are in their first year of life, and those with two genera- tions of primary feathers are at least 2-yr old. I also in- cluded data from some nestling owls that I banded on a site similar to my study area located in Cache County, Utah. Similarly, I used data from several Barn Owls band- ed as nestlings by the Utah Division of Wildlife Resources m Utah County, Utah and recaptured in my study area. Barn Owls were nonmigrator y in northern Utah as they appear to be in most if not all other parts of the species’ range (Schneider 1937, Cramp 1985, Taylor 1994). Statistical analyses (t-tests and linear correlation) were performed using the Statistical Analysis System (SAS Inst. 1988), Rayleigh’s test was used to check for uniformity in direction of owl dispersal after the data were transformed into unimodal data (Zar 1984). Alphas for all tests were 0.05 and all tests were two-tailed. Results Natal Dispersal. I banded 2085 nestlings (locals in USGS Bird Banding Laboratory terminology), 384 breeding adults (adults) and 161 fledglings (hatch year) from 451 nesting attempts by at least individual Barn Owls. To exclude birds that may have died before completing their dispersal, only those that were recovered >6 mo after fledging or after they began breeding were included in the fol- lowing analyses. Of those banded as nestlings, 144 (6.9%) were 0-40 81-120 161-200 41-80 121-160 Dispersal distance, km Figure 3. Comparison of dispersal distances between breeding male and female Barn Owls banded as nesdings in northern Utah. recovered (either found dead or identified alive) at an average of 102.9 ± 162.03 (±SD) km from their natal sites (median = 60 km, range = 0—1267 km; Fig. 2). Among owls banded as nestlings and recaptured as breeders, females {N — 48) moved significantly farther {x — 61.4 ± 52.04 km, median = 57.5 km, range = 0-160 km) than males (N = 34, X = 35.7 ± 36.61 km, median — 14.7 km, range = 0.8-120 km; t = 2.48, df = 80, P = 0.015, power = 0.66; Fig. 3) . One female nested in her natal site and two siblings that dispersed only 8 km from their natal site paired and raised young. Sixty-two owls banded as nestlings were found dead off the study area at distances of 7-1267 km (x = 171.98 ± 223.63 km, median = 109 km) from their natal sites. Sex was determined for only 17 of these and dispersal distances were not significantly different between sexes in this small sample (fe- male, N = H, X = 93.5 ± 63.5 km, median = 110.3, range = 7-167 km; male, N — 9, x = 94.2 ± 90.44 km, median = 52, range = 7-221 km; t = 0.02, df — 15, P = 0.98, power = 0.98). Owls dispersed in all compass directions from their natal sites (Fig. 4), but the pattern of dis- persal direction was not random (Rayleigh’s z — 38.43, P < 0.0005, N = 82). The local topography (Fig. 1) caused many owls to move either to the north, northwest or to the south, southeast. Those 184 Marti VoL. 33, No. 3 Figure 4. Direction and distance of natal dispersal in Barn Owls in northern Utah. that moved beyond the local topographic features dispersed in all directions (Fig. 5) . No relationship was found between the year of fledging and the distance of dispersal (r = —0.08, P = 0.37, N = 135, power = 0.54). Likewise, the severity of a win- ter (based on ambient temperature and depth and persistence of snow cover) was not significantly cor- related with dispersal distance (r = 0.07, P = 0.42, N — 137, power = 0.47). Population density on the study area did not appear to be a factor either; even though numbers of fledglings varied greatly among years (Marti 1994), the number fledged in a year was not correlated with the distance of dis- persal (r — —0.01, P - 0.89, N = 18, power = 0.89), The distance moved from natal site to breed- ing site was not significantly correlated with life- time breeding success in a 19-yr interval (success = number of young fledged in lifetimes [Marti 1997]; r = 0.11, P = 0.32, N = 82, power = 0.70). Unbanded birds that became breeders on my study area provided a measure of dispersal into the area. On average, turnover of breeders at nest sites was 48.1% (range = 21.4— 75.0%/yr), mostly indi- viduals dispersing into the study area. Only 23.3% of first-time breeders had been banded as nestlings on the study area (range = 0-93.8%/yr). The re- maining 76.7% (range = 6.2-100%/yr) were un- handed, apparently having been raised outside the study area (Fig. 6). The nearest known breeding Breeding Dispersal. Nineteen of at least 500 breeders dispersed from one breeding site to an- other. The mean distance moved between breed- ing sites, 2.28 ± 1.77 km (median = 2.25, range = 0.1-6.2 km), did not differ significantly between males and females, but >5 times as many females made those moves (female, N = 16, x = 2.3 ± 1.63 km, median = 2.3, range — 0.1-6. 2 km; male, N = 3, X = 2.17 ± 2.87 km, median — 0.5, range = 0.5- 5.5 km; t = 0.12, df = 17, P = 0.90). Eight of the adults shifted breeding sites in the same year either after a failed first attempt (N = 2) or to produce a second brood following a successful first one (N — 6) . The others changed nest sites in subsequent years. Discussion The natal dispersal that I observed followed a pattern similar to that seen in other Barn Owl pop- ulations (Taylor 1994) with young dispersing soon after fledging and making one-way movements in any direction from the natal site subject to geo- graphic constraints. Distances were usually about 60 km but the longest exceeded 1000 km. Adults, in contrast, tended to be sedentary, rarely moving far from their breeding sites. Stewart (1952) analyzed all banded Barn Owls recovered to 1950 in the U.S. Nestlings banded south of 35°N were all recovered within 144 km of the banding site. Those banded north of 35'’N moved farther: 61% moved >80 km, 28% >320 km and 1% >1600 km. Dispersal, even in the north, was in all directions. Other Barn Owls have been September 1999 Barn Owl Dispersal 185 Total annual turnover of breeders Year Figure 6. Annual turnover rates of breeding Barn Owls in northern Utah. recovered in the U.S. after the 1950s over >1000 km from their banding sites (Broun 1954, Mueller and Berger 1959, Bolen 1978, Soucy 1985). In the United Kingdom, natal dispersal was rel- atively short; only one individual banded as a nest- ling in Scotland moved >20 km to a breeding site (Taylor 1994) and the longest dispersals were <200 km (Bunn et al. 1982). In continental Europe, dis- persals >1000 km were reported from Barn Owl populations in France (Baudvin 1986) and Switzer- land (Glutz von Blotzheim 1979). Over 50% of nestlings banded in Germany bred at distances <50 km from their site of birth, but 24% were re- covered at distances >100 km (Bairlein 1985); movements were shorter in high vole years than in low vole years (Schonfeld 1974). In Spain, natal dispersal covered significantly greater distances than did breeding dispersal (Martinez and Lopez 1995). Natal dispersal occurred in all compass di- rections in Europe, even in Scandinavia (Frylestam 1972) which, like Utah, is at the northern edge of the Barn Owl’s range. In Australia, two nestlings were recovered 250 and 840 km from their nests (Purchase 1972). Dispersal studies of other raptors reveal many similarities. Newton (1979) noted that numerous diurnal raptors in Europe rarely dispersed >50 km and that females dispersed farther than males. The most comprehensive study of dispersal in a diurnal raptor was Newton’s (1986) study on the Sparrow- hawk {Accipiter nisus). Female Sparrowhawks dis- persed significantly farther from their natal areas than did males and both sexes moved in all direc- tions. Most of the natal dispersal occurred in late summer, and population density did not seem to affect dispersal. Dispersal distances were shorter than in Barn Owls (<1-265 km) and 75% settled within 20 km of their natal site. Newton did not record any inbreeding in Sparrowhawks. Breeding dispersal by Cooper’s Hawks {Accipiter cooperii) also resembled the pattern I found in Barn Owls. Male Cooper’s Hawks did not change breeding sites, but a few females moved short distances to new sites (Rosenfield and Bielefeldt 1996). American kestrels {Falco sparverius) in Florida dispersed out of their natal territories but distances were short (71% were <8 km) and the sexes did not differ significantly in distance (Miller and Smallwood 1997). In Wisconsin, natal dispersal by kestrels was much greater and males dispersed far- 186 Marti VoL, 33, No. 3 ther than females (Jacobs 1995). Natal dispersal was not sex-biased in Lesser Kestrels {Falco nau- manni) and 57% settled to breed in their natal col- onies. Those that dispersed moved on average only 18.5 km (Negro et al. 1997). Swainson’s Hawks {Buteo swainsoni) moved on average just 8.2 km (0- 18.1 km) between natal and breeding sites and dis- tances were not significant between the sexes (Woodbridge et al. 1995) . Natal dispersal in a small sample of Ospreys {Pandion haliaetus) averaged 441 km (Johnson and Melquist 1991). Dispersal in owls has not been well documented, but most other species appear to move shorter dis- tances in natal dispersal than do Barn Owls. Me- dian distance moved by radio-tagged Eastern Screech-Owls {Otus asio) from natal sites was only 4.4 km (0.4-16.9) (Belthoff and Ritchison 1989). Also in Eastern Screech-Owls, Gehlbach (1994) re- corded a mean natal dispersal of 3.2 km, but VanCamp and Henny (1975) gave 32 km as the mean natal dispersal distance. However, about half of their birds dispersed <16 km. Mean dispersal by Great Gray Owls {Strix nebulosd) was 18.5 km (7.5- 32; Bull et al. 1988), but Tengmalm’s Owls {Aego- lius funereus) in Finland dispersed on average 55- 70 km (0—320 km) depending on the stage of the vole population cycle (Korpimaki and Lagerstrom 1988). Great Horned Owl {Bubo virginianus) fledg- lings from northern populations moved up to 1305 km from their nests but 53% were recovered within 25 km (Adamcik and Keith 1978). A few male Long-eared Owls {Asia otus) were known to nest within 2.0 km of their natal nest, but females ap- parendy dispersed farther than males before nest- ing (Marks et al. 1994). Dispersal in Spotted Owls {Strix occidmtalis) , despite recent intense study of the species’ biology, is poorly known. Arsenault et al. (1997) and Ganey et al. (1998) radio tracked fledgling Mexican Spotted Owls to distances of 2.1-73.5 km, but only one individual was tracked to a breeding territory at 5.8 km from its natal site. Dispersing juvenile Northern Spotted Owls were tracked from 20-98 km, but none were traced to a breeding territory (Gutierrez et al. 1985). Distance and direction of the natal dispersal I found in northern Utah were effective in reducing inbreeding. The only known inbreeding in my population occurred when dispersal distance was short and siblings from the same brood paired and raised young. Another female bred in her natal site but her mate was not identified. Shaw and Dowell (1989) found one instance of pairing between sib- lings that moved only 5.4 km from their natal site, and another between a mother and son. Incest and close inbreeding have been reported only rarely in other raptors (VanCamp and Henny 1975, Bow- man et al. 1987, James et al. 1987, Postupalsky 1989, Millsap and Bear 1990, Rosenfield and Bie- lefeldt 1992, Taylor 1994, Gutierrez et al. 1995, and Carlson et al. 1998) . It is not clear whether this low level of reported inbreeding is due to the difficulty of detecting it or to a truly low level of occurrence. Natal dispersal may aid in range expansion and repopulation of areas where extinction has oc- curred. The Barn Owls’ ability for long-distance dispersal coupled with their versatility in nest-site and foraging habitat have permitted them to ex- pand their range particularly in response to hu- man-caused habitat changes (Brown 1971, Reese 1972, Stewart 1980, Lenton 1985, McLarty 1995). Barn Owls probably did not nest on my study area until humans provided nesting places (buildings) and increased food availability through irrigated agriculture. Even though I document long-distance move- ments by Barn Owls out of my study area, I do not have any data on the reproductive success of those individuals. However, several individuals that were banded as nestlings made long-distance move- ments out of my study area and were recovered several years later (one 12-yr old), making it likely that they did reproduce. Martinez and Lopez (1995) considered long-distance dispersal by Barn Owls in Spain to be a disadvantage, and Newton and Marquiss (1983) showed that reproductive suc- cess for Sparrowhawks that dispersed furthest was less than those that moved shorter distances. I have considerable data on reproduction by individuals dispersing into my area, but no knowledge of the origin of most of them. Dispersal may also speed the flow of genes among breeding populations, but almost nothing is known about this in Barn Owls or other raptor species. McLarty (1995) compared genetic similar- ity among three Barn Owl populations in British Columbia, northern Utah and southern California, and found sufficient genetic differences to suggest that little gene flow occurs between these popula- tions, Utah and California populations were more similar to each other than the British Columbia population was to either suggesting more east-west than south-north movement by dispersing owls. The British Columbia population is at the north- September 1999 Barn Owl Dispersal 187 ern limit of the species’ range and is of relatively recent origin. My results offer little evidence that natal dis- persal relieves competition because I have repro- ductive data only on birds that moved relatively short (for Barn Owls) distances between natal and breeding sites. I was able to show that lifetime re- productive success was not related to distance of dispersal up to the dispersal distances I was ahle to track, and that dispersal distance was not related to population density. Dispersal in Barn Owls in northern Utah con- formed to the patterns seen in many birds with natal dispersal covering much greater distances than breeding dispersal, and females dispersing farther than males. Natal dispersal apparently was effective in reducing inbreeding, because the dis- tance and randomness of the direction of natal dis- persal made pairings by close relatives unlikely. Acknowledgments Weber State University supported this research through faculty research grants and an endowed scholar program. Phil Wagner helped initiate the study and as- sisted with data collection for the first seven years. I am also grateful to the many landowners who allowed me to use their property for so many years. R.J. Gutierrez, Fred Gehlbach and Dwight Smith made helpful suggestions on an earlier draft. Literature Cited Adamcik, R.S. AND L.B. Keith. 1978. Regional movement and mortality of Great Horned Owls in relation to snowshoe hare fluctuations. Can. Field-Nat. 92:228- 234. Arsenault, D.P., A. Hodgson and P.B. Stacey. 1997. Dis- persal movements of juvenile Mexican Spotted Owls (Strix ocddentalis lucida) in New Mexico. Pages 47-57 in J.R. DuncMi, D.H. Johnson and T.H. Nicholls [Eds.], Biology and conservation of owls on the north- ern hemisphere. USDA For. Serv. Gen. Tech. Rep. NC-190, St. Paul, MN U.S.A. Bairlein, F. 1985. Dismigration und Sterblichkeit in Sud- deutschland beringter Schleiereulen {Tyto alba). Vo- gelwarte 33:81-108. Baudvin, H. 1986. La reproduction de la Chouette ef- fraie {Tyto alba). Le Jean-le-Blanc 25:\-\25. Begon, M., J.L. Harper and C.R. Towsend. 1990. Ecol- ogy, individuals, populations, and communities. Black- well Sci. Publ., Boston, MA U.S.A, Belthoff, J.R. and G. Ritchison. 1989. Natal dispersal of Eastern Screech-Owls. Condor 91:254— 265. Bolen, E.C. 1978. Long-distance displacement of two southern Barn Owls. 5ird-fiandmg 49:78-79. Bowman, R., J.R. Duncan and D.M. Bird. 1987. Dispersal and inbreeding avoidance in the American Kestrel: are they related? Pages 145-150 in D.M Bird and R. Bowman [Eds.], The ancestral kestrel. Raptor Res. Rep. No. 6, Raptor Research Foundation, Hastings, MN U.S.A. Broun, M. 1954. Long-distance recovery of Barn Owl. Bird-Banding 25:149. Brown, L. 1971. African birds of prey. Houghton Mifflin, Boston, MA U.S.A. Bull, E.L., M.G. Henjum and R.S. Rohweder. 1988. Home range and dispersal of Great Gray Owls in northeastern Oregon./. Raptor Res. 22:101—106. Bunn, D.S., A.B. Warburton and R.D.S. Wilson. 1982 The Barn Owl. T. & A.D. Poyser, Calton, U.K. Carlson, P.D., W.S. LaHaye and A.B. Franklin. 1998. In- cestuous behavior in Spotted Owls. Wilson Bull. 110: 562-564. Chanson, J.M., P. Courbet, P. Giraudoux, G. Michaud AND D. Michelat. 1988. Etude sur la reproduction et les deplacements de la Chouette effraie ( Tyto alba) en Franche-Comte: reflexions methodologiques. Alauda 56:197-225. Coles, C.F. and S.J. Petty. 1997. Dispersal behavior and survival of juvenile Tawny Owls {Strix aluco) during the low point in a vole cycle. Pages 111-118 m J.R. Duncan, D.H. Johnson and T.H. Nicholls [Eds.], Bi- ology and conservation of owls on the northern hemi- sphere. USDA For. Serv. Gen. Tech. Rep. NC-190, St. Paul, MN U.S.A. Cramp, S. [Ed.]. 1985. Handbook of the birds of Europe, the Middle East and North Africa, Vol. 4. Oxford Univ. Press, Oxford, U.K. Ferrer, M. 1993. Juvenile dispersal behaviour and natal philopatry of a long-lived raptor, the Spanish Imperial Eagle Aquila adalberti. Ibis 135:132-138. Frylestam, B. 1972. Uber Wanderungen und Sterbli- chkeit beringter skandinavischer Schleiereulen Tyto alba. Ornis Scand. 3:45-54. Ganey, J.L., W.M. Block, J.K. Dwyer, B.E. Strohmeyer AND J.S. Jenness. 1998. Dispersal movements and sur- vival rates of juvenile Mexican Spotted Owls in north- ern Arizona. Wilson Bull. 110:206-217. Gehlbach, F.R. 1994. Eastern Screech-Owl Otus asio: life history, ecology and behavior in suburbia and the countryside. Texas A&M Univ. Press, College Station, TX U.S.A. Glutz von Blotzheim, U.N. 1979. Zur Dismigration jun- ger Schleiereulen Tyto alba. Ornithol. Beob. 76:1-7. Greenwood, PJ- 1983. Mating systems and the evolution- ary consequences of dispersal. Pages 116—131 in I.R. Swingland and P.J. Greenwood [Eds.], The ecology of animal movement. Clarendon Press, Oxford, U.K. AND P.H. Harvey. 1982. The natal and breeding dispersal of birds. Ann. Rev. Ecol. Syst. 13:1-21. Gutierrez, R.J., A.B. Franklin and W. Lahaye. 1995. Spotted Owl {Strix ocddentalis). In A. Poole and F. Gill [Eds.] , The birds of North America, No. 17. Academy 188 Marti VoL. 33, No. 3 of Natural Sciences, Philadelphia, PA and the Amer- ican Ornithologists’ Union, Washington, DC U.S.A. , , , VJ. Meretsky and J.P, Ward. 1985. Juvenile Spotted Owl dispersal in northwestern California: preliminary results. Pages 60-65 in R.J. Gu- tierrez and A.B. Carey [Tech. Eds.], Ecology and management of the Spotted Owl in the Pacific North- west. USDA Eor. Serv. Gen. Tech. Rep. PNW-185, Port- land, OR U.S.A. Jacobs, E.A. 1995. American Kestrel reproduction and dispersal in central Wisconsin./, Raptor Res. 29:135- 137. James, P.C., L.W. Oliphant and I.G. Warkentin. 1987. Close inbreeding in the Merlin {Falco columbarius) . Wilson Bull. 99:718-719. Johnson, D.R. and W.E. Melquist. 1991. Wintering dis- tribution and dispersal of northern Idaho and eastern Washington Ospreys./. Field Ornithol. 62:517—520. KorpimAki, E. 1988. Effects of territory quality on occu- pancy, breeding performance and breeding dispersal in Tengmalm’s Owl./. Anim. Ecol. 57:97-108. AND M. Lagerstrom. 1988. Survival and natal dis- persal of fledglings of Tengmalm’s Owl in relation to fluctuating food conditions and hatching date. /. Anim. Ecol. 57:433-441. , AND P. Saurola. 1987. Field evidence for nomadism in Tengmalm’s Owl Aegolius funereus. Ornis Scand. 18:1-4. Lenton, G.M. 1984. The feeding and breeding ecology of Barn Owls Tyto alba in peninsular Malaysia, /to 126: 551-575. . 1985. History, distribution and origin of Barn Owls Tyto alba in the Malay peninsula. Bull. Br. Orni- thol. Club 105:54-58. Marks, J.S. 1985. Yearling male Long-eared Owls breed near natal nest. / Field Ornithol. 56:181-182. , D.L. Evans and D.W. Holt. 1994. Long-eared Owl {Asio otus). In A. Poole and F. Gill [Eds.], The birds of North America, No. 133. Academy of Natural Sciences, Philadelphia, PA and the American Orni- thologists’ Union, Washington, DC U.S.A. Marti, C.D. 1992. Barn Owl {Tyto alba). In A. Poole, P. Stettenheim and F. Gill [Eds.], The birds of North America, No. 1. Academy of Natural Sciences, Phila- delphia, PA and the American Ornithologists’ Union, Washington, DC U.S.A. . 1994. Barn Owl reproduction: patterns and var- iation near the limit of the species’ distribution. Con- dor 96:468-484. . 1997. Lifetime reproductive success in Barn Owls near the limit of the species’ range. Auk 114:581-592. , P.W. Wagner and K.W. Denne. 1979. Nest boxes for the management of Barn Owls. Wildl. Soc. Bull. 7: 145-148. Martinez, J.A. and G. Lopez. 1995. Dispersal and causes of mortality of the Barn Owl ( Tyto alba) in Spain. Ar- deola 42:29-37. McLarty, J.R. 1995. Genetic variation in three North American Barn Owl (Tyto alba) populations using DNA fingerprinting. M.S. thesis, Univ. British Colum- bia, Vancouver, B.C. Canada. Miller, K.E. and J.A. Smallwood. 1997. Natal dispersal and philopatry of southeastern American Kestrels in Florida. Wilson Bull. 109:226-232. Millsap, B.A. and C. Bear. 1990. Double-brooding by Florida Burrowing Owls. Wilson Bull. 102:313-317. Mueller, H.C. and D.D. Berger. 1959. Some long dis- tance Barn Owl recoveries. Bird-Banding 30-.182. Negro, J.J., F. Hiraldo andJ.A. Donazar. 1997. Causes of natal dispersal in the Lesser Kestrel: inbreeding avoidance or resource competition? /. Anim. Ecol. 53: 901-912. Newton, I. 1979. Population ecology of raptors. Buteo Books, Vermillion, SD U.S.A. . 1986. The Sparrowhawk. T. & A.D. Poyser, Cal- ton, U.K. AND M. Marquiss. 1983. Dispersal of Sparrow- hawks between birthplace and breeding place. /. Anim. Ecol. 52:463—477. POSTUPALSKY, S. 1989. Inbreeding in Ospreys. Wilson Bull. 101:124-126. Purchase, D. 1972. A report on the handing of Barn Owls. Aust. Bird Bander 10:74—75. Reese, J.G. 1972. A Chesapeake Barn Owl population. Auk 89:106-114. Rosenfield, R.N. and J. Bielefeldt. 1992. Natal dispersal and inbreeding in the Cooper’s Hawk. Wilson Bull. 104:182-184. AND . 1996. Lifetime nesting area fidelity in male Cooper’s Hawks in Wisconsin. Condor 167. SAS Institute, Inc. 1988. SAS/STAT user’s guide: statis- tics. SAS Institute, Inc., Cary, NC U.S.A. Saurola, P. 1987. Mate and nest-site fidelity in Ural and Tawny Owls. Pages 81-86 in R.W. Nero, R.J. Clark, R.J. Knapton and R.H. Hamre [Eds.], Biology and con- servation of northern forest owls. USDA For. Serv. Gen. Tech. Rep. RM-142, Fort Collins, CO U.S.A. Schneider, W. 1937. Berigungs-Ergebnisse an der Mitte- leurpaischen Schleiereulen ( Tyto alba guttata Brehm) . Vogelzug 8:1 59-1 70. ScHONFELD, M. 1974. Ringfundauswertung der 1964- 1972 in der DDR beringten Schleiereulen {Tyto alba guttata Brehm) . Jahrb. Vogelwarte Hiddensee 4:90-122. Shaw, G. and A. Dowell. 1989. Breeding by closely re- lated Barn Owls. Ringing Migr. 10:98. SouCY, L.J. 1985. Bermuda recovery of a Common Barn- Owl banded in New Jersey. /. Field Ornithol. 56:274. Steenhof, K., M.N. Kochert and M.Q. Moritsch. 1984. Dispersal and migration of southwestern Idaho rap- tors./. Field Ornithol. 55:357-368. Stewart, RA. 1952. Dispersal, breeding behavior and longevity of banded Barn Owls in North America. Auk 69:227-245. September 1999 Barn Owl Dispersal 189 , 1980. Population trends of Barn Owls in North America. Am. Birds 34:698-700. SwiNGLAND, I.R. 1983. Intraspecific differences in move- ment. Pages 102—115 in I.R. Swingland and P.J. Green- wood [Eds.], The ecology of animal movement. Clar- endon Press, Oxford, U.K. Taylor, I. 1993. Age and sex determination of Barn Owls Tyto alba alba. Ringing Migr. 14:94—102. . 1994. Barn Owls: predator-prey relationships and conservation. Cambridge Univ. Press, Cambridge, U.K. VanCamp, L.F. and C.J. Henny. 1975. The Screech Owl: its life history and population ecology in northern Ohio. N. Am. Fauna 71:1-65. WooDBRiDGE, B., K.K. FiNLEY AND P.H. Bloom. 1995. Re- productive performance, age structure, and natal dis- persal of Swainson’s Hawks in the Butte Valley, Cali- fornia. /. Raptor Res. 29:187-192. Zar, J.H. 1984. Biostatistical analysis. Prentice-Hall, En- glewood Cliffs, NJ U.S.A. Received 5 December 1998; accepted 9 May 1999 J Raptor Res. 33(3):190-206 © 1999 The Raptor Research Foundation, Inc. DIETS OF BREEDING PEREGRINE AND LANNER FALCONS IN SOUTH AFRICA Andrew R. Jenkins Percy FitzPatrick Institute of African Ornithology, University of Cape Town, Rondebosch 7701, South Africa Graham M. Avery South African Museum, P.O. Box 61, Cape Town 8000, South Africa Abstract. — ^The diets of breeding Peregrine {Falco peregrinus) and Tanner {F. biarmicus) Falcons in South Africa were determined from the analysis of prey remains collected at nest sites and through direct observations to determine the regional variation in Peregrine Falcon prey and to measure diet overlap, and the potential for competition between Peregrine Falcons and sympatric congeners. Direct obser- vations suggested that remains under-sampled small prey by about 10% and over-sampled large prey by about 8%. Peregrine and Tanner Falcons preyed mostly on birds. Pigeons and Streptopelia doves com- prised the bulk (38-66% by frequency; 68—85% by mass) of the Peregrine Falcon prey in each of three study areas. Columbids were supplemented by starlings (mostly European Starling \^Sturnus vulgaris^) on the Cape Peninsula, sandgrouse (Pterocles spp.) and swifts (Apus spp.) on the Orange River, and mousebirds (Colius spp.) in the Soutpansberg. Cape Peninsula Peregrine Falcons had the least diverse diet, the narrowest feeding niche and they took the largest proportion of juvenile birds. Peregrine Falcons on the Orange River had the broadest feeding niche and preyed mainly on ‘commuter’ species rather than sedentary residents. Tanner Falcons in the Soutpansberg took mainly terrestrial or cursorial species, particularly young chickens (Callus gallus, 40%; 37%) and charadriids, but columbids were also important. The diets of sympatric Peregrine and Tanner Falcons overlapped by about 35%. Peregrines Falcons concentrated their foraging on woodland and cliff-dwelling prey, while Tanner Falcons took mainly open-country species. Close-neighboring pairs of congeners did not obviously affect the food- niche parameters of either species suggesting that they were not actively competing for food. Key Words: Peregrine Falcon; Falco peregrinus; Fanner Falcon; Falco biarmicus; diet, niche dimensions-, competition. Dieta de Falco peregrinus y Falco biarmicus en reproduccion en Surafrica Resumen. — Tas dietas de Falco peregrinus y Falco biarmicus en reproduccion en Surafrica fueron deter- minadas a partir del analisis de restos de presas recolectados en los sitios de anidacion y a traves de observaciones directas para determinar la variacion regional en las presas de halcon peregrino, medir su coincidencia y la competencia potencial entre halcones peregrinos y sus congeneres simpatricos. Tas observaciones directas sugirieron que las presas pequehas fueron subvaloradas en un 10% y que las presas grandes fueron sobrevaloradas en un 8%. Tos halcones peregrinos y lanarios se alimentaron mas que todo de aves. Palomas y otras aves del genero Streptopelia representaron la mayoria (38—66% de la frecuencia; 68-85% de la biomasa) de las presas de los halcones peregrinos en cada una de las tres areas de estudio. Tas Columbiformes fueron suplementadas por Sturnus vulgaris en la Peninsula del Cabo, Pterocles spp. y Apus spp. en el Rio Orange y Colius spp. en el Soutpansberg. Tos halcones pere- grinos de la Peninsula del Cabo tuvieron la dieta menos diversa y el nicho alimenticio mas restringido cazando aves juveniles. Tos halcones peregrinos del Rio Orange tuvieron el nicho alimenticio mas amplio y depredaron en “especies pas^eras” contrario a las sedentarias. Tos halcones lanarios en el Soutpansberg se alimentaron de especies terrestres particularmente de polios. (Gallus gallus, 40%; 37%) y Charadriformes, las Columbiformes fueron tambien importantes. Tas dietas de los halcones peregrinos y lanarios simpatricos coincidio en un 35%. Tos halcones peregrinos concetraron su forr^eo en presas de bosques y riscos mientras que los halcones lanarios simpatricos coincidio en un 35%. Tos halcones peregrinos concentraron su forrajeo en presas de bosques y riscos, mientras que los halcones lanarios en presas de espacios abiertos. Ta cercania de parejas de congeneres obviamente no afecto los para- metros del nicho alimenticio de ninguna de las dos especies lo que sugiere que estas dos especies no estan compitiendo por comida. 190 September 1999 Peregrine and Lanner Diets in South Africa 191 [Traduccion de Cesar Marquez] Food habits of the Peregrine Falcon {Falco pere- grinus) have been described for arctic and temper- ate regions (Mearns 1983, Hunter et al. 1988, Rat- cliffe 1993, Rosenfield et al. 1995), and for southwestern Australia (Pruettjones et al. 1981, Marchant and Higgins 1993) . However, in the trop- ics, the diet is poorly known and data for African populations are limited (Mendelsohn 1988). The only studies of Lanner Falcon (F. hiarmicus) diets have been in the northern periphery of its range in southern Europe, eastern Sahara, and Israel (Massa et al. 1991, Yosef 1991, Goodman and Haynes 1992). Although Peregrine Falcons share many parts of their distribution with other large falcons having similar resource needs (Cade 1982), few studies have compared Peregrine Falcon diets with those of sympatric congeners (Cade 1960, Por- ter and White 1973) . This paper describes the diets of breeding Peregrine and Lanner Falcons in South Africa, the type and diversity of prey taken by Peregrine Falcons in different environments, defines the feeding niches of the two species in an area of sympatry and assesses the potential for competition between these two congeners in terms of diet overlap. Methods We conducted our study at sites in three areas of South Africa. Sixteen pairs of Peregrine Falcons were studied on the Cape Peninsula, Western Cape Province (34°10'S, 18°25'E) from 1989-95. This area had both inland and coastal nest sites dispersed in a mosaic of urban, forest, and heathland habitats. Altitude varied from 0-1100 m and the climate was temperate with locally variable winter rainfall (400-2000 mm per year). Seven pairs of Pere- grine Falcons were also monitored annually from 1989- 95 on the lower Orange River, Northern Cape Province (28°30'S, 17°00'-20°40'E). This is an arid, hilly to moun- tainous area, sparsely vegetated except for narrow strips of riparian bush or irrigated croplands along the river banks. Altitude varied from 30-700 m and the climate was hot and dry (mean annual rainfall 60-130 mm). Prey remains were collected at seven Peregrine Falcon nest sites in the Soutpansberg range. Northern Province (23°00'S, 29°40'E) from 1988—95, and the diet of these birds was compared with that of a sympatric population of nine pairs of Lanner Falcons over a 3-yr period from 1991-93. Evergreen forest and moist woodland occurred along the eastern foot of the Soutpansberg escarpment, with dry thornveld in the west and grassland with patches of scrub forest and protea woodland along the summit of the mountains (Tarboton 1990). Altitude ranged from 900-1700 m. There was low to moderate summer rainfall (400-1000 mm per year) , increasing on a gradient from northwest to southeast, and temperatures are mild to warm. Prey remains were collected from falcon nest ledges and from below roost sites and feeding perches. Uneaten remains, regurgitated pellets and plucked feathers were used in combination to minimize bias in diet analyses (Simmons et al. 1991, Oro and Telia 1995). Collections were made from just after egg laying to soon after fledg- ing. The frequency of collections varied between sites, years, and areas (Table 1 ) . Care was taken to remove all material on each collection to prevent duplication in sub- sequent samples. Pellets were broken up with tweezers and individual bone remnants were separated from the feather matrix. All avian osteological material was identified using comparative skeletons in the South African Museum and additional material loaned from the Transvaal Museum, Pretoria and the National Museum, Bloemfontein. Indi- vidual body parts were separated according to taxon and recorded as the number of identified specimens (NISP) . The minimum number of individuals (MNI) was calcu- lated from the most common body part among the iden- tified specimens, after accounting for paired elements (Klein and Cruz-Uribe 1984). Incompletely ossified bones were considered to be those of juveniles. Mammal remains were identified according to cranio-dental char- acteristics. Bird and mammal remains were identified to the low- est possible taxonomic level. All prey were assigned to a size class based on bone size for unidentified prey and on body mass data from the literature (Brown et al. 1982, Smithers 1983, Maclean 1993) when prey were identified to at least the family level. Size classes were small (up to sparrow size, average about 20 g), small to medium (star- ling size, average about 60 g), medium (dove size, aver- age about 130 g), medium-large (large dove size, average about 220 g) , large (pigeon size and larger, average about 350 g) , and very large (francolin size and larger, average about 600 g). In biomass calculations, unidentified prey were given these average mass values. Mean body mass values from the literature were used for prey identified to species. Prey identified only to higher taxonomic levels were assigned mass estimates based on published weights of similar or related forms. Arthropod, amphibian and reptile remains were iden- tified at a gross level only. A small number of nonavian remains were considered to be unlikely prey of falcons, particularly Peregrine Falcons, on the basis of size or hab- it. For example, observations suggest that African Pere- grine Falcons hunt only flying prey (Hustler 1983, Tar- boton 1984) and, therefore, are not likely to take terrestrial mammals and reptiles. These taxa (Appendix 1) were excluded from the analyses. Some relatively com- plete, easily identifiable prey were identified in the field and discarded. These were not included in the data and, to avoid duplication, only contributed to the total num- ber of identified prey where fewer individuals of the rel- evant taxon were subsequently identified from other re- mains in the corresponding collection. Pluckings were identified using study skins in the 192 Jenkins and Avery VoL. 33, No. 3 Table 1. Samples of prey remains collected each year at falcon nest sites in three areas of South Africa. NISP denotes the number of identified specimens in the sample and MNI denotes the number of prey individuals identified. The MNI totals are the sum of prey individuals identified from skeletal remains, pluckings and remains identified in situ and discarded. Remains collected from consecutive years were pooled where sample sizes were small. Number Number Year Collections Sites NISP Skeletal Pluckings Discarded Total Cape Peninsula Peregrine Falcons 1989 8 1 318 37 3 22 62 1990 11 4 608 83 6 8 97 1991 7 2 187 33 8 0 41 1992 11 5 412 65 6 1 72 1993 10 5 46 63 10 1 74 1994 22 7 417 88 30 0 118 1995 7 4 161 44 4 1 49 Overall 76 10 2649 413 67 33 513 Orange River Peregrine Falcons 1989-90 7 5 680 102 10 0 112 1991-92 4 4 410 72 8 1 81 1993-95 4 2 333 48 7 0 55 Overall 15 7 1423 222 25 1 248 Soutpansberg Peregrine Falcons 1988-90 5 4 245 36 6 4 46 1991 13 6 719 113 8 0 121 1992 13 5 378 89 5 0 94 1993 14 4 585 93 10 0 103 1994-95 4 3 225 41 2 0 43 Overall 49 7 2152 372 31 4 407 Soutpansberg Tanner Falcons 1991 14 9 00 65 5 0 70 1992 15 7 302 62 3 0 65 1993 10 7 188 38 0 1 39 Overall 39 9 668 165 8 1 174 South African Museum. No attempt was made to deter- mine the number of individuals represented by the sum of pluckings of a particular species collected in each sam- ple. To prevent duplication in these samples, species identified from feather remains contributed one individ- ual to the total for a collection, but only when the rele- vant taxon was not recorded in the skeletal remains. Ju- veniles were identified from plumage characteristics where these differed from adult birds, and from the pre- dominance of sheathed or incompletely grown tail or flight feathers, which indicated that nestling or recently fledged individuals had been taken. Some dietary data were obtained from field observa- tions. These were used to determine bias in the analyses of prey remains (Collopy 1983, Rosenberg and Cooper 1990). Nearly 1000 hr of observations were made at Per- egrine Falcon nest sites on the Cape Peninsula, and over 200 hr were obtained each at nest sites on the Orange River and in the Soutpansberg and at Tanner Falcon sites in the Soutpansberg. Observations were made using 10 X 40 binoculars or a 20— 60X spotting scope from dis- tances of 200-400 m. Whenever a falcon was seen with food, an effort was made to identify or at least estimate the size of its prey. Size classes used were the same as those used for prey remains. Only largely intact prey could be identified or size classified. Indices of diet breadth and overlap were calculated based on the relative frequency of taxa identified in prey remains at the species level wherever possible to refine the quality of these estimates (Greene and Jaksic 1983, Sherry 1990). Diet breadth {B^ was calculated using Lev- ins’ (1968) standardized formula: = — 1/w — 1, where B = \/'% and pi is the proportion of the diet contributed by the ith taxon. Values of B^ range from 0- 1, with larger values indicating a broader diet. Also, the number of frequently used taxa (those comprising 3% or more of the total number of identified prey; Krebs 1989) was tallied for each falcon population as an additional September 1999 Peregrine and Lanner Diets in South Africa 193 estimate of diet breadth. Diet overlap was measured us- ing Morisita’s (1959) index of similarity: C = 2 S pi^ik /2 p,j - 1 ) (A5 - 1) + ^pik (Nk 1 ) 1 ) where p^ and pi)^ are the proportions that taxon i makes up of the diets of species j and k, respectively, and n,* are the number of individuals of taxon i in the diets of species j and k, respectively, and Nj and are the total number of individuals in the diets of species j and k, respectively. Values of C range from 0-1, with larger val- ues indicating a greater dietary overlap. This index is considered the least prone to biases associated with sam- ple size and the number of resources used (Smith and Zaret 1982). Multivariate cluster analyses were conducted to examine qualitative differences in the diet of sympatric Peregrine and Lanner Falcons, using the PRIMER soft- ware package (Plymouth Marine Laboratory, U.K.). Diet composition data were compared using the Bray-Curtis similarity coefficient (using group average linking) to generate a dendrogram of hierarchical clusters and a process of nonmetric multidimensional scaling to gen- erate an ordination plot. To examine the possible competitive influence of near- by Lanner Falcon pairs on the diet of Soutpansberg Per- egrine Falcons and vice versa, the food niche parameters of falcon pairs with congeners as particularly close neigh- bors (pairs <1 km apart, Peregrine Falcons N = 5, av- erage distance to nearest Lanner Falcon pair = 0.6 km, Lanner Falcons N = 4, average distance to nearest Per- egrine Falcon pair = 0.7 km) were compared with those of relatively isolated pairs (>2 km apart, Peregrine Fal- cons N = 4, average distance to nearest Lanner Falcon pair = 6.5 km, Lanner Falcons N = 5, average distance to nearest Peregrine Falcon pair = 4.7 km). The mean distance between conspecific pairs was 9.7 km for Pere- grine and 5.2 km for Lanner Falcons. Two sampling techniques were used to estimate the abundance and distribution of potential avian prey in the vicinity of falcon nest sites in an attempt to make quali- tative assessments of prey selection and habitat use. An estimate of local prey availability was made at nest cliffs by counting the number of potential prey per hour which flew across a sampling area prescribed by a 1 m^ frame. This frame was positioned at the top of the cliff, looking out and down, or on the scree slope below the face look- ing up at the crest. An observer sat 2 m behind the frame, looking through it, to make the count. Prey counts were made at different times of the day at a cross- section of nest sites on the Cape Peninsula and in the Soutpansberg only, mostly from 1991-93. Line-transect counts were walked in the Soutpansberg to estimate bird densities in seven habitats identified in terms of a broad-scale classification of the character and structure of vegetation present {sensu Edwards 1983). Plains to the south of the mountain range comprised ei- ther short, semi-open savanna woodland or short, sparse denuded woodland, often found adjacent to rural setde- ments and heavily impacted by clearing and cultivation. Forest and plantation included tall, closed stands of both indigenous evergreen and exotic eucalyptus or pine for- est. Areas of montane grassland and scrub forest were typical of the upper slopes of the mountains and featured low, open grassland with patches of short deciduous for- est, Deciduous woodland comprised short, moderately closed woodland and occurred on the northern backslo- pes and on the middle and upper slopes at the western end of the range. The lower scree slopes of the escarp- ment were thickly vegetated with low, closed woodland or moist thornscrub in the east and short, moderately closed woodland or dry thornscrub further west. One or two sites were selected as typical of each habitat type and at least three 1 km line transects were walked at these sites each breeding season from 1991-93. Transect sites were located between 150-3000 m from the main escarp- ment. Transects were completed at various times of day, but mostly in the morning or in late afternoon. Transects were walked briskly so only the birds which were con- spicuously active in the area, and hence most likely to provide falcons with hunting opportunities, were record- ed. High, overflying birds unlikely to be resident in the habitat being sampled were not counted, and species considered too large for falcons to catch and subdue un- der normal circumstances (>800 g in weight) were also excluded. The species, number of individuals and ap- proximate perpendicular distance from the transect line were recorded for each sighting (Bibby et al. 1992). Results A minimum of 1168 individuals of at least 82 species from 34 families were identified as prey from the remains collected at Peregrine Falcon nest sites, and a minimum of 174 individuals of at least 24 species from 15 families were identified from Lanner Falcon prey remains (Table 1, Ap- pendix 1). Samples of remains collected at Pere- grine Falcon sites in the Soutpansberg were more substantial than those from sympatric Lanner Fal- con sites, yielding more specimens per collection (Peregrine Falcon x = 45.6, range = 2-211, N = 47 collections; Lanner Falcon x = 18.6, range = 1-60, A = 36 collections; Mann-Whitney Z = 3.01, P = 0.003) and more identified individuals per col- lection (Peregrine Falcon x = 8.0, range = 1-31, A = 49 collections; Lanner Falcon x = 4.5, range = 1-10, A = 39 collections; Z = 2.77, P = 0.006). Peregrine Falcons were observed catching or feeding on 296 prey individuals, about half of which could be identified at least to family. Fifteen species from eight families were identified. Only one species, Black-eyed Bulbul {Pycnonotus barba- tus) was observed as Peregrine Falcon prey and not recorded in prey remains. Only 36 prey individuals were identified during observations at Lanner Fal- con sites in the Soutpansberg. These comprised one locust, one 1-3 day-old chicken ( Gallus gallus ) , one Streptopelia dove, one unidentified murid ro- dent, seven unidentified small birds, nine small- 194 Jenkins and Avery VoL. 33, No. 3 medium birds, 12 medium birds, and four medi- um-large birds. Because relatively few prey items were identi- fied in the field, and the sample of observed prey at banner Falcon sites was so small, the quality of the diet information provided by the analysis of prey remains could only properly be assessed in terms of the size of prey observed at Peregrine Falcon sites in each area. These comparisons sug- gested that remains collected on the Cape Pen- insula provided the least accurate diet estimate, while those from sites in the Soutpansberg were the most accurate (Fig. 1). On average, prey re- mains underrepresented smaller prey hy about 10% and overrepresented larger prey by about 8%. Taxonomically the samples were similar with birds, and particularly columbids, predominating in all. The relative importance of key species in the di- ets of both falcons was fairly consistent between years, so data on prey remains for each year were pooled. Dominant species in the diet of Peregrine Falcons on the Cape Peninsula were Laughing Dove {Streptopelia senegalensis, 28% by frequency and 20% by biomass), European Starling {Sturnus vulgaris, 14 and 7%, respectively), and Cape Turtle Dove {Streptopelia capicola). Peregrine Falcons on the Orange River took Rock Pigeons {Columba guinea, 15 and 41%, respectively), Namaqua Sand- grouse {Pterocles namaqua, 10 and 13%, respective- ly), and Laughing Dove (8 and 6%, respectively) . Peregrine Falcons in the Soutpansberg took Laughing Doves (19 and 16%, respectively). Red- faced Mousebirds {Urocolius indicus, 10 and 4%, re- spectively), and Red-eyed Doves {Streptopelia semi- torquata, 8 and 16%, respectively). During 1991-93, Peregrine Falcons in the Soutpansberg preyed mainly on doves (35 and 45%, respectively) and mousebirds (15 and 7%, respectively), while sym- patric banner Falcons took chickens (40 and 30%, respectively), doves (20 and 25%, respectively), and Crowned Plovers {Vanellus coronatus, 9 and 12%, respectively; Appendix 1). The relative importance of juvenile birds in the diets of Peregrine Falcons varied significantly (x^ = 19.5, P < 0.001) among the three study areas. Most of the juveniles recorded were columbids (Appendix 1) and juveniles were more frequent in the diet on the Cape Peninsula (32 of 513 individ- uals or 6.2%) than in the Soutpansberg (3 of 407 or 0.7%; x^i with Yates’ correction = 15.3, P < 0.001), and comprised 3.2% of the diet of Pere- grine Falcons on the Orange River. The relative frequency of juveniles in the diet of Cape Penin- sula Peregrine Falcons increased through the breeding season from 1.7% by frequency in late October to 6.2% in early November, 6.8% in late November, and 8.1% in early December. A high proportion of the remains recovered from banner Falcon nests were young chickens, and juvenile birds comprised 39.7% of identified prey from these sites. Columbids made up the majority of prey taken by Peregrine Falcons in all three study areas (Table 2). Other consistently important taxa were species in the families Apodidae, Sturnidae and Ploceidae. Soutpansberg banner Falcons took mainly phasian- ids, charadriids and columbids. The average size of prey taken by Peregrine Falcons in each of the three study areas varied (Kruskal-Wallis H = 26.8, P< 0.001) . Cape Peninsula Peregrine Falcons took significantly larger prey (average mass = 144.5 g, range = 15-390 g, N = 513 individuals) than Orange River Peregrine Falcons (average mass = 128.3 g, range = 3-347 g, N = 248 individuals; Mann-Whitney Z= 4.33, P< 0.001) and Peregrine Falcons in the Soutpansberg (average mass = 123.3 g, range = 3-600 g, N = 407 individuals; Z — 4.11, P < 0.001). On average, Soutpansberg Peregrine Falcons took significantly smaller prey (average mass = 116.9 g, range ^ 3-600 g, N = 318 individ- uals) than sympatric banner Falcons (average mass = 123.7 g, range = 3-500 g, N = 174 individuals; Z = 2.41, P = 0.02). Peregrine Falcon diet includ- ed a greater proportion of small and large prey individuals, while banner Falcons tended to con- centrate on medium-sized prey. Of the three Peregrine populations studied, those on the Cape Peninsula had the least diverse diet (30 species were identified from prey remains; Appendix 1) and the narrowest diet breadth {B^ = 0.17). Peregrine Falcons in the Soutpansberg fed on a wide diversity of prey (>50 species) but concentrated on a few taxa and had a moderate diet breadth {B^ = 0.21). On the Orange River, only 38 species were taken but there was less em- phasis on particular families so the diet was rela- tively broad-based {B^ = 0.34). Overall, Peregrine Falcons in the Soutpansberg had a broader feeding niche than sympatric ban- ner Falcons and there was some overlap in the di- ets of the two species {C = 0.34). On a per site basis (excluding one peregrine site and one lanner site with insufficient samples), the difference in September 1999 Peregrine and Lanner Diets in South Africa 195 50 Cape Peninsula small small-medium medium medium-large large 60 S' c (U 3 cr (D 50 40 30 20 - 10 Orange River 0 40 small small-medium medium medium-large large small small-medium medium medium-large large prey size classes □ prey remains ■ observations Figure 1. The relative frequency of prey of different size classes in the diets of Peregrine Falcons in three areas of South Africa determined from the analysis of prey remains and direct observations. Prey remains differed significantly from observations on the Cape Peninsula (remains N = 513, observations N = 159, x ^4 ~ 42.0, P < 0.001) and on the Orange River (remains N = 248, observations AT = 81, x ^4 = 14.0, P = 0.007) but not in the Soutpansberg (remains N = 407, observations N = 56, x ^4 = 7.3, P = 0.12). 196 Jenkins and Avery VoL. 33, No. 3 Table 2. Summary of falcon diets from the analysis of prey remains collected at nest sites (see Appendix 1 for raw data). The importance of each taxon is expressed in terms of its relative frequency in the sample of individuals (%F) and in terms of its contribution to the total biomass of prey in each sample (%M). Families comprising 1% or more of the diet in each sample are listed. Peregrine Falcons (all years): Cape Peninsula N = 513 individuals, 74 107 g; Orange River N = 248 individuals, 31 811 g; and Soutpansberg N = 407 individuals, 50 193 g. Soutpansberg (1991- 93): Peregrine Falcons N = 318 individuals, 37 188 g; banner Falcons N = 174 individuals, 21 532 g. Family Peregrine Falcons (1989-95) Soutpansberg (1991- 93) Cape Peninsula Orange River Soutpansberg Peregrines Lanners %F %M %F %M %F %M %F %M %F %M Phasianidae — — — — 1.5 3.6 1.3 3.6 40.2 37.4 Charadriidae 1.0 — — — 3.9 5.3 3.5 4.9 9.2 12.4 Recurvirostridae — — — — — — — — 1.2 1.6 Glareolidae — — — — 1.0 — — — — — Laridae — — 2.4 2.1 — — — — — — Pteroclididae — — 9.7 14.0 1.0 1.9 — 1.3 — — Columbidae 66.1 85.1 37.5 69.9 48.2 68.4 46.5 67.2 28.2 39.1 Psittacidae 1.0 — — — — — — — — — Apodidae 6.2 1.9 10.1 4.1 6.1 2.9 5.4 2.7 1.2 — Coliidae 1.0 — 1.2 — 13.5 6.1 14.8 7.0 — — Alaudidae — — 7.7 1.0 1.0 — 1.3 — — — Hirundinidae — — 6.1 — — — — — — — Sturnidae 15.0 8.2 1.6 1.4 3.2 2.3 3.1 2.1 — — Ploceidae 4.5 1.1 4.4 — 4.2 1.0 4.4 1.2 4.6 1.2 Estrildidae — — 1.2 — — — — — — — Fringillidae — — 1.6 — 1.0 — — — 1.2 — Molossidae — — 6.5 — — — — — — — Pteropodidae — — — — 1.5 1.6 1.3 1.4 — — Muridae — — — — — — — — 1.2 — diet breadth was statistically significant (Peregrine Falcon average = 0.14, range = 0.11—0.18, N = 6; banner Falcon average = 0.04, range = 0.01- 0.14, A = 8; Mann-Whitney Z = 2.52, P = 0.01). Cluster analyses on diet composition data clearly separated Peregrine and banner Falcon sites (Fig. 2) . Food-niche widths of falcon pairs with pairs of congeners as close neighbors did not differ signif- icandy from those of more isolated pairs (Pere- grine Falcon pairs with close neighboring banner Falcon pairs average 5^ = 0.14, N = 3, isolated pairs average = 0.14, N — 3, Z = 1.00, P — 1.00; banner Falcon pairs with close neighboring Pere- grine Falcon pairs average = 0.08, N = 4, iso- lated pairs average = 0.05, N = 4, Z = 0.72, P = 0.47). However, collectively, interspecific diet overlap was greatest between pairs with close neighboring congeners (close pairs C = 0.44, dis- tant pairs C = 0.26). Counts at cliffs on the Cape Peninsula and in the Soutpansberg had similar average prey passage rates in terms of numbers of individuals counted (Cape Peninsula average = 116.3 birds per hour, range = 2-542 birds per hour, A = 64 counts; Soutpansberg average = 119.0 birds per hour, range = 0-929 birds per hour, A = 114 counts). A significantly greater biomass of birds per hour was recorded in counts on the Cape Peninsula (aver- age = 8567 g per hour, range = 405-33 479 g per hour, A = 64 counts) than in the Soutpansberg (average = 7104 g per hour, range = 0-56 669 g per hour, A= 114 counts; Mann-Whitney Z = 3.09, P = 0.002). Mostly cliff-dwelling species were in- cluded in these counts (e.g.. Rock Pigeons, Alpine Swifts [Apus melba\ and Red-winged Starlings [On- ychognathus morio]). barger species were more abundant at cliffs on the Cape Peninsula than in the Soutpansberg (e.g., Rock Pigeons were record- ed in 72 vs. 18% of counts at cliffs on the Cape Peninsula and in Soutpansberg, respectively, with an average of 3.9 vs. 0.6 individuals per count), whereas aerial insectivores were more abundant in September 1999 Peregrine and Lanner Diets in South Africa 197 LC1 (14) LC2(16) LC3(35)* LC4(15)* , LC5 (26) LC6(32) LC7(20)* LC8(12)* PC2(15)* PCI (92) |~ PC3(110)* — \ PC4 (34) PCS (50) PC6(17)* 20 30 40 50 60 70 80 90 100 Bray-Curtis similarity Figure 2. Results of cluster analyses on diet composition data for Peregrine (P) and Lanner (L) Falcon sites in the Soutpansberg. The dendrogram groups sites according to the Bray-Curtis similarity coefficient (with the number of prey individuals identified from each site in parentheses) . Sites with close neighboring pairs of congeners are marked with an asterisk. The ordination plot (inset) illustrates the relatedness of each site to the others in terms of a multidimensional scaling procedure. the Soutpansberg (e.g., Rock Martins [Hirundo fu- liguld\ were recorded in 19 vs. 52%, respectively, with an average of 0.8 vs. 10.2 individuals per count) . Of the habitats identified in the Soutpansberg, the woodlands on the plain below the escarpment supported the highest diversity and the greatest number and biomass of birds (average = 6-7 spe- cies, 20-30 individuals and 800-1300 g per tran- sect, respectively), while the dense thornscrub on the scree slopes of the mountains featured the most depauperate avifauna (average = <2 species, 2-6 individuals and 40-200 g per transect). Tran- sect data extrapolated to avian density estimates ranging from 1441-1690 individuals/km^ in savan- na and denuded woodland to only 189 individu- als/km^ in moist thornscrub. Some species which were recorded regularly in the diet of falcons in the Soutpansberg were not common in the environment. In particular. Red- faced Mousebirds made up about 10% of the Per- egrine Falcon diet hy frequency but were not en- countered on any of the transect counts. Three species (Chestnut-backed Finchlark [Eremopterix leucotis] , Melba Finch [Pytilia melba] and Blue Wax- bill {Uraeginthus angolensis\) were common on the plains at the foot of the escarpment (x = 13.6, 1.5, and 8.5 individuals, respectively, of each species were recorded per transect through savanna or de- nuded woodland) but were infrequent or did not 198 Jenkins and Avery VoL. 33, No. 3 Table 3. Foraging habitat use by breeding Peregrine and banner Falcons in the Soutpansberg as inferred from the habitat affinities of prey species identified in food remains. (A) only indicates prey species which were recorded in passage-rate or line-transect counts. (B) combines the species in (A) with a number of important prey species which were not sighted on sample counts (e.g., domestic chicken, Namaqua Dove and Red-faced Mousebird). Habitat preferences in species in (A) were determined from count data. The additional species in (B) were assigned to habitats on the basis of incidental observations in the area. Both comparative distributions were significantly different; (A) x^6 = 12.6, P = 0.049; (B) x^e = 79.7, P < 0.001. Favored Habitat OF Prey Species % Frequency IN Peregrine Diet % Frequency in Tanner Diet (A) (B) (A) (B) Savanna woodland 26.7 39.0 19.0 19.5 Denuded woodland 7.9 12.3 12.1 49.4 Forest and plantation 0.3 0.3 — — Grassland and scrub forest 5.4 5.4 1.4 1.4 Moist thornscrub — — — — Deciduous woodland 0.3 0.3 1.4 1.4 Dry thornscrub 0.3 0.3 — — Cliffs 9.1 9.1 5.2 5.2 occur in falcon diets, perhaps because these small birds were underrepresented in prey remains. Sim- ilarly, Black-eyed Bulbuls were ubiquitous in the area (0.5-1 .6 individuals per transect in all habitats except denuded woodland) , but were recorded as falcon prey only once. There was a significant dif- ference in the foraging habitats used by Peregrine and banner Falcons in the Soutpansberg, as in- ferred from the habitat preferences of their prey (Table 3) . While both species favored the wooded plains below the mountain range. Peregrine Fal- cons took more woodland species from relatively pristine habitats whereas Tanner Falcons took open-country species from denuded woodland and free-range chickens from around human settle- ments. Also, Peregrine Falcons preyed more heavi- ly than Tanner Falcons on cliff-dwelling species (Table 3). Soutpansberg Peregrine Falcons took significantly more ‘commuter species’ (aerial in- sectivores, migrants or species which regularly commute between distant resources, e.g., sand- grouse, large columbids) than sympatric Tanner Falcons (76 of 318 or 23.9% vs. 19 of 174 or 10.9% commuters, respectively; x^i ~ 11.3, P < 0.001), but fewer overall than Cape Peninsula Peregrine Falcons (27.0 vs. 31.6% commuters, respectively) and significantly fewer than Peregrine Falcons on the Orange River (110 of 407 or 27.0% vs. 114 of 248 or 46.0% commuters, respectively; x^i = 23.7, P< 0.001). Discussion Columbids and particularly Streptopelia doves ap- pear to be the staple food of Peregrine Falcons in southern Africa (Hustler 1983, Tarboton 1984, Mendelsohn 1988, this study) . A similar preference for columbiforms has been recorded in many oth- er areas of the Peregrine Falcon’s distribution where pigeons and doves are available prey (Cade 1982, Ratcliffe 1993). Small, aerially dexterous spe- cies such as swifts and bats appear more consis- tently in the diet of African Peregrine Falcons than has generally been reported for other populations (Hustler 1983, Brown 1988, Mendelsohn 1988). Conversely, relatively large, terrestrial species such as anatids, galliforms, and charadriiforms are less frequently taken in Africa than in other areas (Cade 1960, Porter and White 1973, Pruettjones et al. 1981, Mearns 1983, Ratcliffe 1993). These differences may partly reflect differences in prey availability, but may also be due to a greater ten- dency for African Peregrine Falcons (and perhaps other small, tropical forms of the species) to ‘catch and carry’ prey rather than strike it to the ground and retrieve it (Cade 1982). Three Streptopelia doves comprised nearly half (by frequency and mass) of the diet of Peregrine Falcons on the Cape Peninsula. British Peregrine Falcons specialize to a similar degree on larger col- umbids (Mearns 1983, Ratcliffe 1993). Typical of peri-urban peregrine populations around the September 1999 Peregrine and Lanner Diets in South Africa 199 world (Cade and Bird 1990), Cape Peninsula pairs also took substantial numbers of commuting feral or domestic pigeons ( Columba livid) and European Starlings, and occasionally preyed on escaped avi- ary birds. Inexperienced, newly-fledged birds are particularly vulnerable to predation by raptors (Newton and Marquiss 1982, Rosenfield et al. 1995), and young doves and starlings were taken quite frequently by Cape Peninsula Peregrine Fal- cons. As observed in other temperate falcon pop- ulations (Newton et al. 1984, Parr 1985), juvenile birds became more prevalent in the diet as the breeding cycle progressed. This supports the con- tention that breeding by Peregrine Falcons on the Cape Peninsula may in part be timed to exploit the period of maximum productivity of their principal prey species (Jenkins 1991, 1998) Streptopelia doves featured least prominently in the diet of Peregrine Falcons on the Orange River, the most generalized feeders of the three Pere- grine Falcon populations studied. These falcons usually foraged in the river valley (Jenkins 1995), which was the focus of bird movements in the area. Aerial insectivores (swifts, hirundines, and micro- chiropteran bats) and obligate drinkers (sand- grouse, columbids, and granivorous passerines) were the groups most frequently taken and were probably exposed to attack crossing rocky gorges or open stretches of the river, or traveling over arid flats adjacent to the floodplain. Soutpansberg Peregrine Falcons preyed on a wide variety of species, most of which were secured in bird-rich woodlands at least 2 km from the base of the escarpment. Relatively fewer transient or commuter species, and more sedentary residents, were taken by these pairs than by Peregrine Fal- cons in the other two areas. However, the wood- land species recorded as prey were generally those likely to fly furthest from, and highest above, the protective canopy of trees (e.g.. Red-faced Mouse- birds, Lamprotornis starlings). Mousebirds were more common Peregrine Falcon prey in the Sout- pansberg than has been found in studies at other woodland sites, whereas rollers and woodpeckers were not taken (Hustler 1983, Tarboton 1984), perhaps because they were less common in the en- vironment. Francolins are among the largest spe- cies regularly taken by African Peregrine Falcons, and those recorded as prey of Soutpansberg Pere- grine Falcons were probably all caught by females. These essentially cursorial birds may have been caught in flight as they descended from the top of the cliff-line to the scree slopes below, as suggested by Hustler (1983). Diet breadth indices calculated for South Afri- can Peregrine Falcons are comparable with equiv- alent data for populations in other areas (Table 4) . These figures suggest that peregrines in temperate areas (e.g., the Cape Peninsula [this study], south- ern Scotland [Mearns 1983], Victoria, Australia [PruettJ^ones et al. 1981]) are relatively specialized feeders. In contrast. Peregrine Falcons appear to become more generalized in extreme environ- ments such as deserts (e.g., this study), taiga and tundra (Cade 1960, Cade et al. 1968). Relatively small samples of identifiable prey re- mains were recovered from Lanner Falcon nest sites, perhaps because much of the prey they con- sumed was completely digestible. Few data are available on the diet of southern African Lanner Falcons. The only quantitative studies, based on small samples, suggest that phasianids (including domestic fowl) and columbids are important prey groups (Barbour 1971, Tarboton and Allan 1984, Kemp 1993) . Lanner Falcons in the Soutpansberg were relatively specialized feeders, with young chickens comprising nearly 40% of the prey re- corded in remains. Given the greater susceptibility of juvenile bones to damage and acid erosion, chickens may have been underrepresented in these samples. Columbids were also important, and contributed more than phasianids to the biomass of identified prey. Sources of free-ranging poultry were a considerable distance from most Lanner Falcon nest sites, and their frequency in prey re- mains suggests that Lanner Falcon hunting ranges extended well beyond those of nearby Peregrine Falcon pairs. One radio-tracked male Lanner Fal- con foraged over rural settlement areas >10 km from the main escarpment. Terrestrial or cursorial species predominated in the diet of Soutpansberg Lanner Falcons, and most prey were probably tak- en on or close to the ground. Southern African Peregrine and Lanner Falcons are of approximately equal body mass and com- parisons of key food handling and flight perfor- mance measurements yield ratios consistently low- er than values traditionally considered minimum for noncompetitive coexistence (Jenkins 1998). Despite these structural similarities, diet overlap between Soutpansberg Peregrine and Lanner Fal- cons was moderate relative to those calculated us- ing data from other studies of Peregrine Falcons and sympatric congeners, and fell well below the 200 Jenkins and Avery VoL. 33, No. 3 Table 4. Indices of Peregrine Falcon diet breadth for a number of locations around the world, and measures of diet overlap between Peregrine Falcons and sympatric congeners. Study Population Number of Taxa Used Number of Taxa >3% Diet Breadth (Ba) Southern Scotland^ 90 5 0.04 Victoria, Australia^ 77 6 0.12 Cape Peninsula, South Africa^ 39 6 0.17 Soutpansberg, South Africa^ 63 9 0.21 West Greenland^ 11 6 0.24 NW Territories, Canada® 19 7 0.34 Orange River, South Africa^ 45 10 0.34 Taiga zone, Alaska® 61 11 0.81 Sympatric Peregrine Falcons and Gyrfalcons {F. rusticolus), Colville River, Alaska® Peregrine Falcon 55 12 0.33 Gyrfalcon 55 2 0.004 Diet overlap: C = 0.07 Sympatric Peregrine and Lanner Falcons, Soutpansberg, South Africa^® Peregrine Falcon 67 10 0.20 Lanner Falcon 67 6 0.07 Diet overlap: C = 0.34 Sympatric Peregrine and Prairie Falcons (F. mexicanus), Wasatch Mountains, Utah^^ Peregrine Falcon 33 9 0.31 Prairie Falcon 33 6 0.26 Diet overlap: C = 0.58 Food niche parameters were calculated using data from: ^ Mearns 1983, ^ Pruett-Jones et al. 1981, ^ this study, this study (all years), ® Rosenfield et al. 1995, ® Bradley and Oliphant 1991, ’ this study, ® Cade et al. 1968, ® Cade and White 1971, this study (1991-93), Porter and White 1973. suggested critical value for competing species of about 0.62 (Bosakowski and Smith 1992). Also, while competition theory predicts narrower food niches and reduced diet overlap between close pairs of competing species vs. distant pairs (Nilsson 1984, Korpimaki 198'7), the diet of falcon pairs in this study was not obviously affected by the pres- ence of nearby pairs of congeners. Therefore, there was little evidence suggesting active compe- tition for food between Peregrine and Tanner Fal- cons in the Soutpansberg. As concluded in previ- ous studies of Peregrine Falcons and sympatric congeners (e.g.. Cade 1960, Porter and White 1973), relatively subtie differences in morphology, flying performance, and hunting techniques (Jen- kins 1995) are evidently sufficient to segregate for- aging habitat and diet of Peregrine and Tanner Falcons in South Africa. Acknowledgments Dave Allan, Zelda Bate, Tim Wagner, and Anthony van Zyl assisted in the field. Dave Allan kindly contributed prey remains collected at a Peregrine Falcon nest site in the Soutpansberg in 1988. South African National Parks allowed us to conduct research in the Augrabies Falls Na- tional Park. Tom Cade, Rob Davies, James Enderson, Phil Hockey, Rob Simmons, and Reuven Yosef made useful comments on earlier drafts of this manuscript. This re- search was partly funded by the Foundation for Research Development. Literature Cited Barbour, D.Y 1971. Notes on the breeding of the lanner. Bokmakierie 23:2-5. Bibby, C.J., N.D. Burgess and D.A. Hill. 1992. Bird cen- sus techniques. Academic Press, London, U.K. Bosakowski, T. and D.G. Smith. 1992. Comparative diets of sympatric nesting raptors in the eastern deciduous forest biome. Can. J. Zool. 70:984-992. Bradley, M. and L.W. Oliphant. 1991. The diet of Per- egrine Falcons in Rankin Inlet, North-West Territo- ries: an unusually high proportion of mammalian prey. Cowfior 93:193— 197. Brown, C.J. 1988. Diet of Booted Eagles and Peregrine September 1999 Peregrine and Lanner Diets in South Africa 201 Falcons in the Waterberg Plateau Park, Namibia. Ga- bar 3:26-28. Brown, L.H., M.K. Urban and K. Newman. 1982. The birds of Africa. Vol. 1. Academic Press, London, U.K. Cade, T.J. 1960. Ecology of the peregrine and Gyrfalcon populations in Alaska. Univ. California Publ. Zool. 63: 151-290. . 1982. The falcons of the world. Collins, London, U.K. AND C.M. White. 1971. Cliff-nesting raptors and ravens along the Colville River in Arctic Alaska. Living Bird 10:107-150. AND D.M. Bird. 1990. Peregrine Falcons, Fato^er- egrinus, nesting in an urban environment: a review. Can. FieM-Nat. 104:209-218. Cade, T.J., C.M. White and J.R. Haugh. 1968. Peregrines and pesticides in Alaska. Condor 70:170-178. COLLOPY, M.W. 1983. A comparison of direct observation and collections of prey remains in determining the diet of Golden Eagles, y. Wildl. Manage. 47:360—368. Edwards, D. 1983. A broad-scale structural classihcation of vegetation for practical purposes. Bothalia 14:705- 712. Goodman, S.M. and C.V. Haynes. 1992. The diet of the lanner {Falco biarmicus) in a hyper-arid region of the eastern region of the Sahara./. Arid Env. 22:93-98. Greene, H.W. and F.M. Jaksic. 1983. Food niche rela- tionships among sympatric predators: effects of level of prey identification. Oikos 40:151-154. Hunter, R.E., J.A. Crawford and R.E. Ambrose. 1988. Prey selection by Peregrine Falcons during the nest- ling stage. /. Wildl. Manage. 52:730-736. Hustler, K. 1983. Breeding biology of the Peregrine Fal- con in Zimbabwe. Ostrich 54:161-171. Jenkins, A.R. 1991. Latitudinal prey productivity and po- tential density in the Peregrine Falcon. Gabar 6:20-24. . 1995. Morphometries and flight performance of southern African Peregrine and Lanner Falcons. J. Avian Biol. 26:49-58. . 1998. Behavioural ecology of Peregrine and Lan- ner Falcons in South Africa. Ph.D. dissertation, Univ. Cape Town, Cape Town, South Africa. Kemp, A.C. 1993. Breeding biology of Lanner Falcons near Pretoria, South Africa. Ostrich 64:26-31. Klein, R.G. and K. Cruz-Uribe. 1984. The analysis of an- imal bones from archaeological sites. Univ. Chicago Press, Chicago, IL U.S.A. Korpimaki, E. 1987. Dietary shifts, niche relationships and reproductive output of coexisting kestrels and Long-eared Owls. Oecologa 74:277-285. Krebs, C.J. 1989. Ecological methodology. Harper Col- lins, New York, NY U.S.A. Levins, R. 1968. Evolution in changing environments: some theoretical explorations. Princeton Univ. Press, Princeton, NJ U.S.A. Maclean, G.L. 1993. Roberts’ birds of southern Africa. John Voelcker Bird Book Fund, Cape Town, South Africa. Marchant, S. and PJ- Higgins [Eds]. 1993. Handbook of Australian, New Zealand and Antarctic birds. Vol 2. Oxford Univ. Press, Auckland, New Zealand. Massa, B., F. Lo Valvo, M. Siracusa, M. and A. Ciaccio. 1991. II Lanario Falco biarmicus feldeggi (Schlegel) in Italia: status, biologia e tassonomia. Naturalista Sicil. S. 15:27-63. Mearns, R. 1983. The diet of the Peregrine Falcon Falco peregrinus in south Scotland during the breeding sea- son. Bird Study 30:81-90. Mendelsohn, J.M. 1988. The status and biology of the peregrine in the Afrotropical Region. Pages 297-306 in T.J. Cade, J.H. Enderson, C.G. Thelander and C.M. White [Eds.], Peregrine Falcon populations: their management and recovery. The Peregrine Fund, Boi- se, ID U.S.A. Morisita, M. 1959. Measuring of interspecific association and similarity between communities. Mem. Fac. Sa. Kyushu Univ. Ser. E (Biol.) 3:65-80. Newton, I. and M. Marquiss. 1982. Food, predation and breeding season in Sparrowhawks {Accipiter nisus) . J Zool, Lond. 197:221-240. , E.R. Meek and B. Little. 1984. Breeding season foods of Merlins Falco columbarius in Northumbria. Bird Study 31:49—56. Nilsson, I.N. 1984. Prey weight, food overlap, and repro- ductive output of potentially competing Long-eared and Tawny Owls. Ornis Scand. 15:176-182. Oro, D. and J.L. Tella. 1995. A comparison of two meth- ods for studying the diet of the Peregrine Falcon. J. Raptor Lies. 29:207-210. Parr, S.J. 1985. The breeding ecology and diet of the Hobby FaZco subbuteo in southern England. Ibis 127:60- 73. Porter, R.D. and C.M. White. 1973. The Peregrine Fal- con in Utah: emphasizing ecology and competition with the Prairie Falcon. Brigham Young Univ. Sci. Bull. 18:1-74. Pruett-Jones, S.G., C.M. White and W.R. Devine. 1981. Breeding of the Peregrine Falcon in Victoria. Emu 80: 253-269. Ratcliffe, D.A. 1993. The Peregrine Falcon. T. & A.D. Poyser, London, U.K. Rosenberg, K.V. and RJ. Cooper. 1990. Approaches to avian diet analysis. Stud. Avian Biol 13:80—90. Rosenfield, R.N., J.W. Schneider, J.M. Papp and W.S. Seegar. 1995. Prey of Peregrine Falcons breeding in west Greenland. Cowdor 97:763-770. Sherry, T.W. 1990. When are birds dietarily specialized? Distinguishing ecological from evolutionary ap- proaches. Stud. Avian Biol 13:337-352. Simmons, R.E., D.M. Avery and G. Avery. 1991. Biases in diets determined from pellets and remains: correc- tion factors for a mammal and bird-eating raptor J Raptor Res. 25:63—67. 202 Jenkins and Avery VoL. 33, No. 3 Smith, E.P. and T.M. Zaret. 1982. Bias in estimating niche overlap. Ecology 63:1428-1253. Smithers, R.H.N. 1983. The mammals of the southern African subregion. Univ. Pretoria, Pretoria, South Af- rica. Tarboton, W.R. 1984. Behaviour of the African pere- grine during incubation. Raptor Res. 18:131-136. . 1990. The Soutpansberg. Fauna & Flora 47:1-6. AND D.G. Allan. 1984. The status and conserva- tion of birds of prey in the Transvaal. Transvaal Mus. Monogr. 3:1-115. Yosef, R. 1991. Foraging habits, hunting and breeding success of Tanner Falcons {Falco biarmicus) in Israel. J. Raptor Res. 25:77-81. Received 5 December 1998; accepted 26 April 1999 Appendix 1. Invertebrate and vertebrate taxa identified from remains collected at Peregrine and Tanner Falcon nest sites in South Africa. Remains were collected at 10 Peregrine Falcon sites on the Cape Peninsula, seven Peregrine Falcon sites on the Orange River, and seven Peregrine Falcon sites and nine Tanner Falcon sites in the Soutpansberg. Data provided are the average mass of each taxon and the number of individuals identified, with the number of nest sites at which each taxon was collected in parentheses. Items marked with an asterisk were not considered as falcon prey in the diet analyses. Prey Taxon Average Mass (g) Cape Peninsula Orange Rtver Soutpansberg Peregrines Tanners Insects Unidentified orthopterans 3 1 (1) 12 (2) 3 (2) Unidentified coleopterans 3 — 1 (1) 2 (1) — Amphibians Unidentified frogs 20 *2 (1) — Birds Tittle Sparrowhawk — male 80 1 (1) Acdpiter minullus Peregrine Falcon — nestling 250 *2 (2) Tanner Falcon — nestling 250 — — — *1 (1) Coqui Francolin 230 — — 1 (1) — Francolinus coqui Crested Francolin 340 2 (1) 1 (1) F sephaena Grey-wing Francolin 390 1 (1) - F. africanus Cape Francolin — pullus 150 1 (1) F. capensis Natal Francolin — female 425 2 (2) F. natalensis Natal Francolin — unsexed 520 1 (1) Unidentified adult francolin 500 — — — 2 (2) Unidentified pullus francolin 150 — — — 1 (1) Domestic chicken — <3days old 60 — — — 41 (8) Callus gallus Domestic chicken — ±7 days old Domestic chicken — ±14 days old 150 23 (3) Harlequin Quail 250 — — — 1 (1) C. delegorguei White-fronted Plover Charadrius marginatus 49 2 (2) Kittlitz Plover C. pecuarius 43 2 (1) — — — September 1999 Peregrine and Lanner Diets in South Africa 203 Appendix 1. Continued. ^ Soutpansberg Average Cape Orange Prey Taxon Mass (g) Peninsula River Peregrines Lanners Crowned Plover — adult 167 Vanellus coronatus Crowned Plover — immature 1 67 Greenshank 191 Tringa nebularia Curlew Sandpiper 57 Calidris ferruginea Little Stint 24 C. minuta Black-winged Stilt 175 Himantopus himantopus Temminck’s Courser 74 Cursorius temminckii Common Tern 124 Sterna hirundo White-winged Tern 57 Chlidonias leucopterus Unidentified Tern 100 Namaqua Sandgrouse 185 Pterocles namaqua Burchell’s Sandgrouse 235 P. burchelli Double-banded Sandgrouse 235 P bicinctus Feral or domestic pigeon — adult 320 Columba livia Feral or domestic pigeon — ^young 320 Rock Pigeon — adult 347 C. guinea Rock Pigeon — fledgling 347 Rameron Pigeon 415 C. arquatrix Unidentified pigeon — adult 330 Unidentified pigeon — ^young 330 Red-eyed Dove — adult 235 Streptopelia semitorquata Red-eyed Dove — fledgling 235 Cape Turtle Dove 153 S. capicola Laughing Dove — adult 102 S. senegalensis Laughing Dove — fledgling 102 Unidentified dove — fledgling 130 Namaqua Dove 40 Oena capensis Green-spotted Dove 65 Turtur chalcospilos Small unidentified columbid 50 Medium unidentified columbid 160 Cockatiel 90 Nymphicus hollandicus 1 (1) — 16 (5) 14 (6) — — 2 (2) — 1 (1) — 1 (1) 3 (1) — — — — 1 (1) — — — — 2 (1) 2 (2) — — 4 (3) — 2 (1) 3 (1) — — — 1 (1) — — 2 (2) — — — 24 (6) — — — — 1 (1) — — — 3 (2) — 47 (7) 8 (4) 17 (3) — 1 (1) — — 28 (7) 35 (6) 20 (5) 6 (2) 8 (3) 2 (2) 1 (1) 1 (1) — — 2 (2) — 1 (1) — — — 2 (2) — — — 40 (7) 16 (5) 34 (6) 10 (4) 1 (1) — — — 61 (7) 5 (4) 25 (5) 8 (5) 140 (8) 20 (5) 79 (7) 21 (7) 4 (2) — — 2 (2) 3 (3) — — — 3 (3) 7 (5) 15 (5) — — — 1 (1) 1 (1) 1 (1) — — — 1 (1) — 1 (1) — — — 204 Jenkins and Avery VoL. 33, No. 3 Appendix 1. Continued. Prey Taxon Average Mass (g) Cape Peninsula Orange River Soutpansberg Peregrines Lanners Budgerigar Melopsittacus undulatus 28 4 (1) — 1 (1) — Grey Loerie Corythaixoides concolor 269 — 1 (1) — Freckled Nightjar Caprimulgus tristigma 60 — 2 (1) — — Unidentified nigh^ar 60 — — 1 (1) — Black Swift — adult Apus barbatus 45 13 (5) — 12 (3) — Black Swift — fledgling 45 1 (1) — — — Bradfield’s Swift A. bradjieldi 45 — 18 (6) — — Little Swift A. af finis 26 12 (3) "■ ~ 2 (2) — Alpine Swift A. melba 77 6 (4) 6 (5) 11 (3) 2 (2) Unidentified swift 35 — 1 (1) — — Speckled Mousebird Colius striatus 53 — — 14 (6) 1 (1) White-backed Mousebird — adult C. colius 41 1 (1) ' 2 (2) — — White-backed Mousebird — fledgling 41 — 1 (1) — — Red-faced Mousebird Urocolius indicus 56 4 (3) — 41 (7) — Narina Trogon Apaloderma narina 65 — — 1 (1) — Pied Kingfisher Ceryle rudis 82 — 1 (1) — — European Bee-eater Merops apiaster 55 — — ^ 1 (1) — Hoopoe Upupa epops 57 1 (1) 1 (1) — Scimitar-billed Woodhoopoe Phoeniculus cyanomelas 35 — 1 (1) — — Red-billed Hornbill Tockus erythrorhynchus 130 — — — 1 (1) Black-collared Barbet Lybius torquatus 57 — 2 (2) — Red-capped Lark Calandrella cinerea 26 — 1 (1) — Chestnut-backed Finchlark Eremopterix leucotis 14 — — 1 (1) 1 (1) Grey-backed Finchlark E. verticalis 17 — 19 (1) — — Unidentified lark 20 — — 2 (1) — European Swallow Hirundo rustica 18 1 (1) — — — Rock Martin H. fuligula 22 1 (1) 4 (4) — — Brown-throated Martin Riparia paludicola 13 — 11 (3) — September 1999 Peregrine and Lanner Diets in South Africa 205 Appendix 1 . Continued. Average Cape Orange Soutpansberg Prey Taxon Mass (g) Peninsula River Peregrines Lanners Black-headed Oriole Oriolus larvatus 69 — — 1 (1) — Unidentified sylviid 20 — 2 (2) — — Fiscal Shrike Lanius collaris 41 — 1 (1) Unidentified shrike 60 — 1 (1) — — European Starling — adult Sturnus vulgaris 76 68 (9) ■ ' European Starling — fledgling 76 5 (3) — — — Pied Starling Spreo bicolor 107 1 (1) Plum-colored Starling Cinnyricinclus leucogaster 46 1 (1) Cape Glossy Starling Lamprotornis nitens 84 6 (3) Greater Blue-eared Glossy Starling L. chalybaeus 76 1 (1) Unidentified starling 80 — — 2 (1) — Red-winged Starling — adult Onychognathus morio 135 3 (3) 1 (1) Red-winged Starling — fledgling 135 — — 1 (1) — Pale-winged Starling O. nabouroup 107 4 (2) Unidentified starling 80 — — 1 (1) 1 (1) White-browed Sparrow-weaver Plocepasser mahali 48 “““ 3 (3) 1 (1) Sociable Weaver Philetairus socius 27 1 (1) House Sparrow Passer domesticus 24 2 (1) — Cape Sparrow P. melanurus 26 10 (3) 4 (3) 4 (3) 2 (2) Cape Weaver Ploceus capensis 45 7 (4) 3 (2) Masked Weaver P. velatus 27 1 (1) 1 (1) 1 (1) Unidentified weaver 35 — — 1 (1) 1 (1) Red-billed Quelea Quelea quelea 19 — 2 (2) 6 (3) 1 (1) Red Bishop Euplectes orix 23 2 (2) 1 (1) Yellow-rumped Widow E. capensis 45 2 (1) Unidentified ploceid 40 1 (1) 1 (1) — 2 (2) Common Waxbill Estrilda astrild 8 3 (3) Red-headed Einch Amadina erythrocephala 23 3 (3) Yellow-eyed Canary Serinus mozambicus 13 1 (1) 206 Jenkins and Avery VoL. 33, No. 3 Appendix 1 . Continued. Prey Taxon Average Mass (g) Cape Peninsula Orange Rtver SOUTPANSBERG Peregrines Lanners Forest Canary 16 — — 1 (1) — S. scotops Bully Canary 26 — — — 1 (1) S. sulphuratus Yellow Canary 17 1 (1) 1 (1) — — 5. flaviventris White-throated Canary 27 — 2 (1) — — S. albogularis Unidentified canary 15 1 (1) — 1 (1) — Unidentified fringillid 15 — 1 (1) 1 (1) 1 (1) Unidentified small birds — adults 20 4 (3) 2 (2) 9 (3) 2 (2) Unidentified small birds — ^young 20 1 (1) — — Unidentified small-medium birds 60 6 (4) — 11 (4) 8 (6) Unidentified medium birds 130 5 (3) 4 (4) 2 (2) 4 (2) Unidentified very large birds 600 — — 1 (1) — Mammals Unidentified shrew 10 — — — 1 (1) Unidentified fruit bat 130 — — 6 (2) — Flat-headed free-tailed bat 14 — 8 (1) — — Sauromys petrophilus Egyptian free-tailed bat 15 — 8 (1) — — Tadarida aegyptiaca Hildebrandt’s horseshoe bat 30 — — 1 (1) — Rhinolophus hildebrantii Unidentified insectivorous bats 15 1 (1) 6 (4) 3 (3) — Multimammate mouse 65 — — =^1 (1) — Mastomys natalemis Namaqua rock mouse 45 — n (1) — — Aethomys namaquensis Small unidentified murid 25 — — — 1 (1) Medium unidentihed murid 45 *2 (2) *3 (3) *2 (2) 1 (1) Large unidentihed murid 120 — — *2 (2) — Unidentihed squirrel 160 — — *1 (1) — Small unidentihed bovid 5000 — — *1 (1) — TOTAL 515 (10) 252 (7) 418 (7) 175 (10) J. Raptor Res. 33(3):207-2l7 © 1999 The Raptor Research Foundation, Inc. NEST DISPERSION, DIET, AND BREEDING SUCCESS OF BLACK KITES {MILVUS MIGRANS) IN THE ITALIAN PRE-ALPS Fabrizio Sergio Edward Grey Institute of Field Ornithology, Department of Zoology, South Parks Road, Oxford 0X1 3PS, U.K. Alberto Boto Dipartimento di Biologia Animals, Piazza Botta 9, 27100 Pavia, Italy Abstract. — ^We studied a population of Black Kites (Milvus migrans) from 1992-96 in a 100-km^ study area in the Italian pre-Alps around Lake Lugano. Population density increased from 24 territorial pairs per 100 km^ in 1992 to 38 in 1996. Nearest neighbor distances were variable, averaging 1288 m for solitary pairs {N = 24) and 306 m for colonial ones {N = 151). Regular spacing of nest sites was the rule within colonies, but inter-nest distance for solitary breeders increased as new pairs settled in the area. Nests occurred both in trees (58%, N = 84) and on cliffs (42%); 23% and 5% {N = 84) of the nests were originally built by Common Buzzards (Buteo buteo) and Ravens (Corvus corax), respectively. Mean laying date was 25 April (N = 42), mean clutch size was 2.29 eggs {N = 42) and mean number of fledged young was 0.97, 1.11, and 1.78 young per territorial, reproductive, and successful pairs, respectively (N = 143, 95, 78). The percentage of successful territorial pairs was 55% {N = 143). Diet was dominated by fish and birds, which accounted for 62% and 28% of 307 identified prey items, respectively. Compared with other European populations, this population showed an intermediate den- sity and an average clutch size but the lowest fledging and breeding success ever recorded for the species. Reasons for the low success were unclear, but may have been related to low food availability or water pollution and consequent pesticide contamination. Key Words: Black Kite, Milvus migrans; coloniality, density, diet, breeding success; Italy. Dispersion de nidos, dieta, y exito reproductivo de Milvus migrans en los pre-Alpes Italianos Resumen. — Estudiamos una poblacion de Milvus migrans desde 1992-96 en un area de 100 kms en los pre-Alpes italianos alrededor del Lago Lugano. La densidad poblacional aumento de 24 parejas terri- toriales por 100 kms en 1992 a 38 en 1996. La distancia mas proxima entre vecinos fue variable y promedio 1288 m entre parejas solitarias {N = 24) y 306 m entre las parejas coloniales {N = 151). Los espacios regulares de los nidos fueron la regia entre las colonias, pero la distancia entre nidos para los reproductores solitarios se incremento a medida que se asentaron nuevas parejas en el area. Los nidos se encontraron tanto en arboles (58%, N = 84) como en riscos (42%), 23% y 5% {N = 84) de los nidos fueron originalmente construidos por Buteo buteo y Corvus corax, respectivamente. La media del dia de la puesta fue el 25 de abril {N = 42). La media del tamano de la nidada fue 2.29 huevos (AT = 42) y la productividad media fue 0.97, 1.11, y 1.78 juveniles por territorio. Las parejas productivas territori- almente exitosas representaron el 55% {N = 53). La dieta fue principalmente dominada por peces y aves (62% y 28% respectivamente) de los 307 items de presas identificadas. A1 comparar con otras poblaciones europeas, estas poblaciones mostraron una densidad intermedia y una nidada promedio, pero al mismo tiempo el mas bajo exito reproductivo y reclutamiento registrado para la especie. Las causas de esto, son poco claras y pudieron estar relacionadas con la poca disponiblidad de comida o con la polucion del agua y la consecuente contaminacion por pesticidas. [Traduccion de Cesar Marquez] The Black Kite {Milvus migrans) is a medium- sized accipitrid that is distributed throughout the Old World and Australasia. It has been defined as an adaptable and opportunistic feeder (Geroudet 1965, Delibes 1975, Arroyo 1978, Jones and Manez 1990) and as one of the most numerous and suc- cessful birds of prey in the world (Brown and Ama- don 1968). Despite its local abundance, the Black Kite was recently classified as a declining and vul- nerable species in Europe (Vinuela and Sunyer 207 208 Sergio and Boto VoL. 33, No. 3 1994). Although populations were reported to be stable or increasing in western Europe between 1970-90 (e.g., Bustamante and Hiraldo 1993, Dou- meret 1995), pronounced decreases have been ob- served in Portugal, eastern Europe, and Russia during the same period (Vihuela and Sunyer 1994, Bijlsma 1997) . Recently, recorded cases of pesticide contamination (Hernandez et al. 1986, Jenni-Eier- mann et al. 1996), nest robbing (Garcia Ferre and de Juan 1983) and water pollution (e.g., Ceschi et al. 1996) in the species’ core distribution areas point to an urgent need for quantitative monitor- ing of breeding populations (Vinuela and Sunyer 1994). In Italy, both local population increases and de- clines have been reported (Petretti 1992, De Gia- como et al. 1993), but information is fragmentary and mainly nonquantitative (Vihuela and Sunyer 1994). During the 20^*^ century, the species de- clined dramatically in the Po Plain. For example, one colony in Bosco Fontana, near the city of Man- tova, declined from over 100 pairs in the 1930s (Ar- rigoni degli Oddi and Moltoni 1931) to <10 pairs in the 1970s (Micheli 1990). The largest Italian populations are currently concentrated in the pre- Alps (Micheli 1990). No quantitative estimates of density, diet or productivity have been published for these populations. In this paper, we report the results of a 5-yr study on a Black Kite population in a pre-Alpine area of northern Italy. The aim of the research was to pro- vide quantitative data on Black Kite density, nest dispersion, colony size, diet and productivity, and to compare them with estimates from other pop- ulations. Study Area The study area was a 100-km^ plot located along the Italian coast of Lake Lugano. It included two pre-Alpine valleys: the Ceresio Valley along the lake and the Ganna Valley further inland from the lake. Altitude ranged from 275-1125 m. The landscape was characterized by pre-Al- pme mountains, mainly of sedimentary origin, ranging from 526-1125 m in peak elevation. Mountain slopes were covered by continuous deciduous woodland which accounted for 77% of the study area and was locally in- terrupted by a few open areas. These open areas resulted from human activities (regular burning or sheep graz- ing) and were covered by dry grassland, ferns, or various bushes, mainly common hazel {Corylus avellana). Domi- nant tree species included sweet chestnut {Castanea sali- va) , downy oak ( Quercus pubescens) , sessile oak ( Quercus petraea), European hop-hornbeam {Ostrya carpinifolia) , and locust (Robinia pseudoacacia). Forests were managed for timber production by stool shoot regeneration (cop- pice system, Matthews 1989) on a 20-30 yr rotation basis. Single mature trees are often kept into the next rotation as seed bearers (coppice with standards system, Matthews 1989). Thus, even though single mature trees were pres- ent almost everywhere, most woodlots had been thinned or clearfelled in recent times (<40 yr), resulting in a homogeneous cover of second growth forest. Following the recent decline in coppice management throughout Europe (Matthews 1989), some woodlots were being con- verted to mature forest; nevertheless, mature woodland was confined to a few steep and often inaccessible slopes. Cultivated fields, mainly grassland and maize {Zea mays) fields, were located on the valley floors and ac- counted for 3% of the land. Sixteen small villages, all confined to the valley floors, covered 13% of the study area. Thus, except for forestry operations, human activ- ities were mainly confined to lowland and were virtually absent from the mountain slopes. The Lake Lugano wa- ter surface accounted for 6% of the study area which also included two smaller lakes (17 and 25 ha); overall, aquat- ic habitat covered 7% of the study area. Climate was temperate continental, with wet springs and dry summers (Pinna 1978). Annual rainfall ranged from 1600-2100 mm, with two peaks, one in the spring and one in the autumn, and the former more pro- nounced than the latter (Belloni 1975). Methods We censused Black Kite nests from 1992-96. The ear- liest territorial individuals were observed in the study area on 18 March. As soon as kites settled on their breed- ing territories, they started to perch in highly prominent positions, refurbish nests and carry out conspicuous ae- rial displays to signal territory ownership or to attract a mate (Vinuela 1993). Nests were censused by progressive, complete searches as new pairs arrived from migration and settled on territories. No pairs settled on territories after the middle of April, when some pairs were already laying eggs. All nests were found during the pre-incuba- tion period, by watching the resident pair’s territorial dis- plays and nest material transfers. This also allowed us to census the nonbreeding sector of the territorial popula- tion (i.e., pairs that failed to lay eggs). Nest dispersion was analyzed by means of the G-statistic (Brown 1975), calculated as the ratio of the geometric to the arithmetic mean of the squared nearest neighbor dis- tances (NNDs) Values ranged from 0-1 with values rang- ing from 0.65-1 .00 indicating a regular dispersion of nest sites (Brown 1975). Pairs were defined as solitary when they nested >700 m from their nearest neighbor. How- ever, five nests belonging to two territories were defined as solitary despite having a NND <700 m. These were located on opposite sides of a mountain and the birds could not see their nearest neighbor while in their nest area; also, they were not seen to interact with the neigh- boring pairs during each of three 1-hr observation ses- sions during the pre-laying period (see Brown and Brown 1996). Apart from these five cases, all pairs <700 m from their nearest neighbor were defined as colonial, and all interacted regularly with their neighbors, mostly through contests caused by territorial intrusions. Whenever possible, nest contents were checked three times: first during incubation to assess clutch size, second September 1999 Black Kite Breeding Ecology 209 Table 1. Density and extent of coloniality in a Black Kite population in the Italian pre-Alps between 1992-96. Variable 1992 1993 1994 1995 1996 No. of territorial pairs 27 33 35 39 41 Territorial pairs/ 100 km^ 24 30 32 36 38 No. of colonial pairs (%) 26 (96) 29 (88) 28 (80) 35 (90) 33 (80) No. of colonies 6 6 6 7 5 just after hatching to estimate hatching success, brood size, and date of hatching, and third, when the nestlings were 40-45 d old to record the number .of young raised (nestlings fledge at about 48 d old, Bustamante and Hir- aldo 1993). We checked nest contents by climbing the nest tree, going down the nesting cliff with a rope or simply watching the nest cup from a distant vantage point with binoculars or a 20-60 X telescope. To minimize dis- turbance, risk of desertion or nest predation by Ravens ( Corvus corax) , only nests that could be checked very rap- idly were visited during incubation or early hatching. Thus, clutch size, laying date and number of laying pairs were assessed from a subsample of nests. Hatching date was calculated by backdating from the feather develojj- ment of nestlings first observed when <15 d old and by comparison to reference information in Richard (1934), Cramp and Simmons (1980) and Hiraldo et al. (1990). Laying date was estimated by subtracting 29 d, the aver- age incubation period (Vinuela 1993), from hatching date. Additional data were collected from a loose colony of three pairs 3 km from the border of the study area. We collected prey remains found under nests during each nest visit and identified them to the genus or spe- cies level assuming the smallest possible number of in- dividuals. Fish scales were identified following Steinmetz and Muller (1939) and Baligliniere and Le Louarn (1987) ; other vertebrate remains were identified by com- parison to a reference collection. Terminology followed Steenhof (1987) with a repro- ductive pair being one which laid eggs, a successful pair being one which raised at least one young until it was 40 d old and breeding success being the percentage of suc- cessful territorial pairs. Raptors generally nest in tradi- tional nest areas year after year (Newton 1979); a nest area was defined as an area where more than one nest was found in the same or in different years, but where only one pair nested each year. Stepwise logistic regression analysis (Norusis 1993) was employed to analyze among year differences in the prob- ability of territorial pairs laying eggs and of eggs hatch- ing. Logistic regression uses a linear combination of in- dependent variables to explain the variance of a dependent dichotomous variable. To meet the assump- tions of normality, NNDs were always log transformed pri- or to parametric tests. Means are given with ±1 SE. All tests were two-tailed and statistical significance was set at P < 0.05. Results Density and Nest Dispersion. The number of ter- ritorial pairs increased steadily from 27 in 1992 to 41 in 1996. Density correspondingly increased from 24 to 38 territorial pairs per 100 km^ (Table 1). The loose colony outside the study area con- tained 3 pair in each of the five study years. The number of nest areas censused was 48. Of these, six were always occupied by solitary pairs, 37 by colonial pairs and five by pairs that switched from colonial to solitary nesting {N — 3), or vice versa {N = 2), during the study period. Territorial behavior was intense within 50-200 m of nests but was rarely observed at foraging areas. The number of colonies censused was stable and ranged from five to seven. Colonies contained on average 5.0 nests (±0.6, N = 30). The percentage of colonial pairs ranged from 80—96% (Table 1). In particular, the number of pairs in large colonies (>5 pair) did not increase significantly through the years (x^ = 2.12, d£ = 4, P = 0.710). Instead, the number of pairs in small colonies (^4 pair) progressively decreased (x^ = 9.76, df = 4, P = 0.040), except in 1995 (Fig. 1) when a new two- pair colony was created by settlement by a new pair near a previously solitary pair. The percentage of solitary pairs did not increase significantly over time (x^ = 4.92, df = 4, P = 0.301, Fig. 1). Mean NND was 306 m for colonial pairs (±13.7, N = 151) and 1288 m for solitary pairs (±166, N = 24). NNDs ranged from 60-690 m for colonial pairs and from 450-4250 m for solitary pairs. Mean NND of solitary pairs did not increase through the years (Kruskal-Wallis x^ = 195, df = 3, P = 0.583, Table 2). For three large colonies (A, B, and C, Table 2), we tested the effect of year and colony on nest spacing by means of a two-way ANOVA: the interaction of year and colony was almost signifi- cant (partial Fg gg = 1.86, P = 0.076) and both year and colony significantly affected mean NND inde- pendendy (respectively; partial F 4 93 = 11.01, P = 0.000; partial Fg gs = 20.78, P = 0.000). Variation in NND within colonies among years was mainly caused by the progressive increase in territorial pairs in two colonies around the lake. New pairs 210 Sergio and Boto VoL. 33, No. 3 30 tn '5 1992 1993 1994 1995 1996 Colony size large (5-1 3 pr) small (2-4 pr) solitary pairs Year Figure 1. Number of Black Kite pairs breeding solitarily or in small (S4 pairs) and large (^5 pairs) colonies between 1992-96 in the Italian pre-Alps. Table 2. Mean (±SE) Black Kite nearest neighbor distances for 24 solitary pairs and at seven colonies in the Italian pre-Alps between 1992-96. Mean Nearest Neighbor Distance (m) Nest Type 1992 1993 1994 1995 1996 1992-96 Solitary 934 ± 161 1203 ± 199 1304 ± 332 1508 ± 437 1288 ± 166 (4)® 0.82” (7)® 0.72” (4)® 0.68” (8)® 0.38” (24)® 0.54” Colony A 394 ± 40 373 ± 64 288 ± 49 252 ± 28 241 ± 33 295 ± 20 OY 0.881^ (8)® 0.67” (9)® 0.64” (12)® 0.79” (13)® 0.68” (49)® 0.67” Colony B 292 ± 12 169 ± 24 (7)® 0.89” 167 ± 15 150 ± 25 183 ± 10 OY 0.96*’ (8)® 0.67” (9)® 0.87” (8)® 0.67” (39)® 0.78” Colony C 679 ± 24 355 ± 51 290 ± 23 237 ± 44 174 ± 23 307 ± 34 (3)^ O.lOb (5)® 0.85” (5)® 0.95” (5)® 0.75” (7)® 0.83” (25)® 0.59” Colony D 597 ± 27 597 ± 27 567 ± 37 567 ± 37 433 ± 123 552 ± 29 (3)^ 0.99” (3)® 0.99” (3)® 0.98” (3)® 0.98” (3)® 0.74” (15)® 0.92” Colony E 401 ± 3 317 ± 45 640 ± 0 640 ± 0 640 ± 0 500 ± 44 (3)“ 0.10” (3)® 0.93” (2)“ (2)® (2)® (12)® 0.83” Colony F 213 ± 10 206 ± 0 206 ± 0 206 ± 0 208 ± 3 (3)^ 0.99” (2Y (2)® (2)® (9)® 0.99” Colony G 403 ± 0 (2) 403 ± 0 (2) All nests 430 ± 58 (27)® 394 ± 48 (33)® 492 ± 75 (35)® 388 ± 63 (39)® 493 ± 115 (41)® 441 ± 36 (175)® ^ Sample size. G statistic. September 1999 Black Kite Breeding Ecology 211 settling in these colonies did not place their nests in the outer parts of colonies or far away from tra- ditional pairs, but selected areas near already- occupied nests. Thus, mean within-colony NND de- creased as the number of pairs in the colonies increased though the years (colony A; F4 44 = 2.64, P= 0.046; colony C; F4 20 = 11.29, P= 0.000, Table 2). Mean NND did not vary significantly with time in colony C, where the number of pairs remained stable through the years (F4 34 = 2.07, P = 0.106). Due to this, NNDs are shown separately for each colony in Table 2. The G-statistic declined progressively for solitary pairs through the years, indicating a regular dis- persion of nest sites between 1993-95, but not in 1996 (Table 2). Pooling years, the overall disper- sion of solitary nests was not regular (G = 0.540), probably because of the progressive increase in the mean NND of solitary nesters over time. Instead, the G-statistic indicated a regular dispersion of nest sites within 22 of 23 colonies with >3 pairs. Nest Sites. All territorial pairs built nests {N = 175 pairs in 48 nest areas). After settling on terri- tories, most pairs brought nest material to one nest, but 13 pairs in six nest areas brought material to two nests during the beginning of the pre-laying period, before selecting the nest in which they eventually laid eggs. Altitude of the nest sites ranged from 240-870 m, averaging 515 m (±14, N = 175). The 175 censused territorial nestings took place in 84 different nest structures, each one oc- cupied for one to five consecutive years (x = 2.1 ±0.1, A =84). Of these nests, 58% were placed in trees while the remainder were on cliffs {N = 84) . Of 49 tree nests, 39% were in sweet chestnut, 33% in sessile or downy oak, 18% in locust, 6% in European hop- hornbeam, 2% in common lime {Tilia europaea), and 2% in Scotch pine {Pinus silvestris) trees. The 35 cliff nests were placed on bare rock ledges (29%) and at the base (54%) or in the canopy (17%) of trees, generally downy oaks, growing on cliff faces. Many pairs used nest structures origi- nally built by other species, to which they added new material: 23% of 84 occupied nest structures were originally built by Common Buzzards (Buteo buteo) and 5% were old Raven nests. Some nests were being used by the original occupants when they were taken over by Black Kites. Buzzard nests were taken over during the original owners’ pre- laying period and one Raven nest still contained nestlings when it was occupied by kites, which preyed on the chicks before starting to add new material to the nest. Breeding Season. There were no year-to-year dif- ferences in mean laying date (F440 = 0.68, P = 0.762, Table 3). Laying dates ranged from 12 April-8 May, averaging 25 April (Table 3) . No cases of replacement clutches in the same nest were ob- served after breeding failures, even when these Table 3. Mean (±SE) laying date and productivity estimates of an Italian pre-Alpine Black Kite population between 1992-96. Variable 1992 1993 1994 1995 1996 1992-96 Laying date in April 24.6 ± 1 (7)^ 23.3 ± 2 (7)- 25.5 ± 2 (8)- 28.7 ± 3 (9)- 25.3 ± 2 (14)- 25.6 ± 1 (45)- Territorial pairs’’ % reproductive 17 31 31 33 31 143 Pairs { N ) 100 (17)^ 76 (21)- 86 (22)- 86 (21)- 74 (19)- 84 (100)- Clutch size { N ) 2.5 ± 0.2 2.5 ± 0.2 1.8 ± 0.3 1.8 ± 0.4 2.6 ± 0.2 2.3 ± 0.1 (42)- (lO)^’ (11)- (8)- (5)- (8)- % hatching success 68 (25)‘^ 89 (27)*= 79 (14)^ 89 (9)^ 100 (21)^ 84 (96)*= % successful pairs 71 58 48 42 61 55 Young fledged per territorial pair 1.06 ± 0.2 1.16 ± 0.2 0.71 ± 0.2 0.67 ± 0.2 1.32 ± 0.2 0.97 ± 0.1 Young fledged per 1.00 ± 0.2 1.24 ± 0.2 0.86 ± 0.2 0.76 ± 0.2 1.72 ± 0.3 (18)- 1.11 ± 0.1 (97)- reproductive pair (16)- (21)- (21)- (21)- Young fledged per 1.50 ± 0.2 2.00 ± 0.2 1.47 ± 0.2 1.57 ± 0.2 2.16 ± 0.2 (19)- 1.78 ± 0.1 (78)- successful pair (12)- (18)- (15)- (14)- ‘‘Sample size (when different from that in column “territorial pairs”). ’’ Those checked for nesting success. Number of eggs. Sample size of nests same as for clutch size. 212 Sergio and Boto VoL. 33, No. 3 Table 4. Diet of breeding Black Kites in the Italian pre- Alps (1992-97), as estimated by food remains collected under nests. Prey Category Number of Items (%) Fishes 190 (61.9) Rudd {Scardinius erythrophthalmus) 59 (19) Chub {Lends cus cephalus) 56 (18) Unidentified Ciprinidae 37 (12) Others^ 38 (12) Birds 85 (27.7) Blackbird ( Turdus merula) 20 (7) Others'’ 65 (21) Unidentified Passeriformes 20 (7) Mammals 24 (7.8) Unidentified Microtidae 15 (5) Others'^ 9 (3) Reptiles'' 5 (1.5) Amphibians'^ 1 (0.3) Invertebrates' 2 (0.7) Total 307 ® Includes: perch {Perea fluviatilis N = 13), bleak {Alburnus alhidus alborella, N = 11), roach {Rutilus rutilus, N= 8), pumpkinseed (Lepomis gibbosus, N = A), large-mouth black bass {Micropterus sal- moides, N = 2). Includes: Rock Dove {Columba livia, N — II), Starling {Sturnus vulgaris, N = 6), Jay {Garrulus glandarius, N — 5), Italian Sparrow {Passer domesticus italiae, N — A), Black Kite nestling {N = 2), Honey Buzzard nestling {Pemis apivorus, N= 2), Tawny Owl {Strix aluco, N = 2), Hooded Crow {Corvus corone cornix, N = 2), Song Thrush {Turdus philomelos, N = 2), Moorhen {GaUinula chloropus, N — 1), Collared Dove {Streptopelia decaocto, N= \), Barn Swallow {Hirundo rustica, N - 1), House Martin {Delichon urbica, N = \), Chaffinch {FringiUa coelebs, N — 1), Great Tit {Parus major, N = 1), Reed Bunting {Embetiza schoeniclus, N = 1). Includes: dormouse {Myoxus glis, N = 3), common mole ( Talpa europaea, N = 3), rabbit ( Oryctolagus cuniculus, N = 2), unidenti- fied mammal from slaughterhouse {N =1). ^ Includes: Aesculapian snake {Elaphe longissima, N = 2), uniden- tified Colubridae {N = 3). Includes: Rana spp. {N = 1). ^Includes: Cetonia spp. {N = 2). happened early in incubation. All nestlings fledged by 15 July and all Black Kites departed the study area by 10 August each year. Productivity. Each year, 7^100% of territorial pairs laid eggs (Table 3), with no clear trend over time. Year did not enter a stepwise logistic regres- sion model with laying or nonlaying of eggs as a dependent variable. The overall mean number of laid eggs (x = 2.3, N = 42) varied significandy among years (F 4 37 = 2.75, P = 0.043) and was sig- nificanUy lower in 1994 than in 1992, 1993, and 1995 (Duncan’s Multiple Range Test, P< 0.05, Ta- ble 3). On the whole, 84% of 96 eggs checked hatched. Stepwise logistic regression analysis showed a significant effect of year on probability of egg hatching (B = 0.65, Wald = 5.88, df = 1, P = 0.015, reclassification rate = 84%). The number of young„ fledged per territorial pair averaged 0.97 {N = 143) and differed signifi- candy among years (F 4138 = 2.49, P = 0.045). Pro- ductivity was significantly higher in 1996 than in 1994 and 1995 (Duncan’s Multiple Range Test, P < 0.05, Table 3). Significant among-year differenc- es also were detected for mean number of fledged young per reproductive pair (F 492 = 2.68, P = 0.036). Again, average values were significantly higher in 1996 than in 1994 and 1995 (Duncan’s Multiple Range Test, P < 0.05, Table 3). Mean number of fledged young per successful pair was 1.78 {N = 78) and also varied among years (F 4 73 = 3.69, P = 0.009), with more young fledged on average in 1996 than in 1992, 1994, and 1995 and in 1993 than in 1994 (Duncan’s Multiple Range Test, P < 0.05, Table 3). Finally, the percentage of successful territorial pairs in the population ranged from 42% in 1995 to 71% in 1992, with no significant among-year differences (x^ = 4.98, df = 4, P = 0.301, Table 3). Sample size allowed us to test the effect of year and colony on mean number of fledged young in three colonies (A, B, and C) by means of a two-way ANOVA. The effect of the interaction between year and colony on mean number of young fledged per territorial pair was significant (partial Fg 83 = 3.01, P = 0.005) , but only approached significance when using mean number of fledged young per repro- ductive pair or per successful pair as the depen- dent variable (respectively: partial Fg gg = 1.80, P = 0.092; partial Fg ,42 = 1.96, P = 0.093). In these two cases, colony was the only significant main effect (respectively: partial F 2 gg = 3.32, P = 0.042; partial F‘ 2,42 ~ 3.54, P = 0.038). These patterns were main- ly caused by the productivity of colony C, where fledging success remained similar to that in other colonies until 1995, when no young were raised in any of the eight nests checked. In 1996, only one young was raised in seven nests checked in colony C. The reason for such a dramatic decline in pro- ductivity was not clear. Overall, causes of breeding failure were seldom identified: 89% of failures occurred during incu- bation {N =19) and the rest occurred during the early part of the fledging period. Probable causes of failure included: failed hatching of eggs and September 1999 Black Kite Breeding Ecology 213 Table 5. Nesting density of Black Kite populations in Europe, 1966-96. Area (Period) Habitat (N) Density (pr/lOOkm^) Nearest Neighbor Distance (AO Reference Matas Gordas, Spain (1987-89) Grassland and marshland 21-45 700-1500 Vinuela et al. 1994 Rhone Plain, France (1970) River plain 140 609 Sermet 1980 Donana, Spain (1981-84) Marshland 80 266.6 206 (47)^ Hiraldo et al. 1990 Lac Leman, Switzerland (1975-90) Farmland and lake 319 100.6 Henrioux and Henrioux 1995 Neuchatel, Switzerland (1968) Farmland and lake 337 69.6 Sermet 1980 Lorraine, France (1966) Woodland and pas- ture 66 44.6 Thiollay 1967 Lake Lugano, Italy (1992- 96) Woodland and lake 27-41 24-38 441 (175) This study Castelporziano, Italy (1991- 92) Woodland and farmland 16 33.3 103 (16) De Giacomo et al. 1993 Lake Constance, Germany (1968-69) Farmland and lake 25-30 18.5-22.2 Heckenroth 1970 Limousin, France (1976- 78) Bocage'’ 21 9.5 Nore 1979 Dromling, Germany (1993- 94) Farmland 8 7.02 2330 (8)^^ Seelig et al. 1996 Monti della Tolfa, Italy (1973-80) Woodland and pas- ture 42 4.9 Petretti and Petretti 1981 Brandenburg, Germany (1979) Farmland 215 0.74 Fiuczynski 1981 “Estimate from Bustamante and Hiraldo (1990), for the period 1985—88. Pasture and woodland enclosed by stone walls or hedges. Calculated from the published map. consequent desertion (three cases), predation (two cases) and disturbance by rock climbers (one case) . Diet. Fish and birds dominated the diet and ac- counted for 61.9% and 27.7% of 307 identified prey items, respectively (Table 4). Mean length of 15 fish found uneaten in nests was 18.6 cm (±1.8, range = 7-30 cm). Of 54 avian prey items, 26% were juveniles. Discussion The number of territorial pairs of Black Kites increased steadily during our study. It is difficult to say whether this was caused by a real population increase or a simple population fluctuation. Spatial and temporal variations in Black Kite population density and productivity have been reported by many authors (Fiuczynski and Wendland 1968, Vinuela et al. 1994, Doumeret 1995, Bijlsma 1997). Results of the analyses on NND and fledging suc- cess in the three large colonies showed that each colony could be considered a separate unit within the population. Each colony was characterized by its own nest spacing distance, which tended to be regular, but which differed from that of other col- onies and sometimes varied from one year to the next. Similarly, productivity varied among colonies within and among years within each colony. These results emphasized the importance of monitoring populations composed of more than just one col- ony for as many years as possible to obtain an un- biased estimate of density, nest dispersion, and pro- ductivity. To date, many studies on Black Kites have been conducted in just one colony or in areas of homogeneous high density that are functionally similar to one large colony (e.g., Desai and Mal- hotra 1979, Koga et al. 1989, Vinuela et al. 1994). Overall, the population showed remarkable op- 214 Sergio and Boto VoL. 33, No. 3 Table 6. Productivity of Black Kite populations in Europa and Asia, 1966—96. Area (Period) Habitat N Clutch Size Hatching Success‘s (Eggs) Breeding Success Donana, Spain (1987-89) Marshland 166 92% Germany (1992-95) 599 79% Nagasaki, Japan (1983-86) Fishing port 32 2.18 (28)’’ 79% (61)” 75% Limousin, France (1976-78) Bocage^ 22 68% Berlin, Germany (1940-79) Farmland 215 62% Lake Lugano, Italy (1992-96) Woodland and 143 2.29 (42)^^ 84% (96)” 55% lake Slovakia (1975-89) 162 2.98 (44) Lac Leman, Switzerland Farmland and (1975-90) lake 165 2.25 Lorraine, France (1966) Woodland and pasture 66 2.26 (45)’’ New Delhi, India (1973-76) Urban 45 2.3 (60)’’ 55% (102)” ^Data on hatching success not shown in table: 75% {N = 36 eggs from 14 nests, Hakel, Germany, 1957; Stubbe 1961) and 64% (N = 28 eggs from 10 nests, Mazio, Italy, date unknown; Petretti 1992). Sample size (when different from that in column “N”). Data also from Mammen and Stubbe (1995, 1996). Pasture and woodland enclosed by stone walls or hedges. portunism and elasticity. First, variability in the ex- tent of coloniality between and within years was paralleled by high flexibility in NNDs, the highest observed NND being 70 times greater than the lowest one. Interestingly, as the population in- creased over the years, new pairs settling in large colonies positioned their nests near existing pairs, possibly as a result of conspecific attraction (Stamps 1988), as was also reported by Vihuela et al. (1994). Advantages of nesting near traditional older pairs (Vihuela 1993) could include local en- hancement (Hagan and Walters 1990), informa- tion parasitism (e.g., Green 1987) and greater op- portunities for kleptoparasitism (Vihuela et al. 1994). The opposite trend was recorded for soli- tary pairs whose NNDs increased as new pairs col- onized the area, suggesting pronounced overall variations in individual pairs nest dispersion, prob- ably in relation to food distribution (Newton 1979, Vihuela et al. 1994). Second, Black Kite nests were found in a variety of situations; in mature and young trees, within large forests, in small clumps of two or three trees, in isolated trees and on large (>100 m high) and small (<20 m high) cliffs. Nests originally built by other species, especially Common Buzzards, also were occupied readily and taken over by persistent harassment of the original occupants. Cliff nesting, which is very common all over the Italian pre-Alps, is very rare elsewhere in the Black Kite’s range (Cramp and Simmons 1980) and only locally re- ported in Sicily (Massa 1985), in an area of central Spain (Blanco 1997) and near some Swiss lakes (Sermet 1980). Finally, diet was dominated by fish and birds on a numeric basis. However, all vertebrate classes and some occasional invertebrate prey were also rep- resented. This finding agreed with the common definition of Black Kites as opportunistic feeders with local specialization in the most available prey (Delibes 1975, Arroyo 1978, Vihuela and Veiga 1992, Blanco 1997). Despite the overall opportunism and adaptabili- ty, productivity estimates for our study population were remarkably low when compared to other pub- lished estimates in Europe, Japan and India (Table 5). Nesting densities across Europe were highest near large wetlands and intermediate or very low in farmed or extensive livestock rearing land- scapes. Preference for proximity to large wetlands and higher densities near water bodies have been reported by many authors (Heckenroth 1970, De- libes 1975, Hiraldo et al. 1990, Henrioux and Hen- rioux 1995). Density in the Italian pre-Alpine area was much lower than that of populations near large wetlands (Sermet 1980, Hiraldo et al. 1990, September 1999 Black Kite Breeding Ecology 215 Table 6. Extended. Mean Number of Fledged Young Territorial Breeding Successful Pair Pair Pair Reference 1.77 1.92 (153)'’ Jones and Manez 1990 1.63 2.07 (471)'’ Gedeon 1994*^ 1 00 1.14 (28)'^ 1.33 (24)'’ Koga et al. 1989 1.32 1.93 (15)'’ Nore 1979 1 20 1.90 (133)'’ Fiuczyn.ski 1981 0.97 1.1 (95)^^ 1.78 (78)'’ This study 2.31 Danko 1989 2.02 Henrioux and Henrioux 1.32 1995 1.58 (55)'’ Thiollay 1967 0 98 Desai and Malhotra 1979 Vinuela et al. 1994, Henrioux and Henrioux 1995) . It was similar to that along Lake Constance in the 1960s and in low intensity farmland with scattered wetlands (Heckenroth 1970, Thiollay 1967, Seelig et al. 1996), but higher than that in farmed or pas- ture habitats with few wetlands (Nore 1979, Fiuc- zynski 1981, Petretti and Petretti 1981). Thus, the observed density was intermediate, but low, com- pared with other populations near large lakes or wetlands. Clutch size was similar to that of other popula- tions (Table 6) and hatching success also was com- parable to that of other published estimates, ex- cept the one of New Delhi, although all of these data were from studies conducted outside Europe or during the pesticide era (1960-70s; Newton 1979). In contrast, the percentage of successful pairs and mean number of fledged young per ter- ritorial, reproductive, or successful pair were the lowest ever recorded in Europe and similar only to those in India and Japan. Reasons for such low na- tality were unclear. During the 1960s, trapping and persecution were common practice in the county of the study area. For example, Bianchi et al. (1969) reported gamekeepers killing 22 Black Kites in one month in a l-km^ area. However, this practice now is rare and we observed no cases of nest robbing, trapping, or persecution. Clutch and nestling predation was minimal, as important po- tential predators such as Goshawks {Acdpiter gentil- is) and Eagle Owls {Bubo bubo) were very scarce in the study area. Thus, the remaining potential caus- es of low natality could be low food availability or water pollution and consequent prey contamina- tion. DDT contamination of water and fish con- tamination was reported recently in nearby Lake Maggiore (Ceschi et al. 1996). Even though lake euthrophication and pollution could benefit this species by providing dead or moribund fish (Ger- oudet 1965, Bijeleveld 1974, Doumeret 1995), pes- ticide contamination clearly can be harmful for Black Kites (e.g., Jenni-Eiermann 1996, Bijlsma 1997) causing low natality and eventual population decline in the long term. The aim of current and future research on Black Kites in northern Italy is to monitor pesticide concentrations in eggs, pop- ulation density and productivity levels in different lakes and correlate them with levels of pollution in water and fish prey. There is need for studies as- sessing the level of water contamination above which toxic chemicals concentrations start to be harmful, both for Black Kites and their fish prey. More published estimates of density and produc- tivity, especially hatching success, are also needed from other European countries. Acknowledgments We thank G. Bogliani for continuous encouragement during each phase of the research and G. Bogliani, C.M Perrins, M. Henning, P.G.W. Salaman, K. Steenhof, and D.E. Varland for constructive comments on a first draft of the paper. We thank L. Marchesi for identification of avian prey remains. Part of this study was funded by Re- gione Lombardia, Servizio Faunistico. 216 Sergio and Boto VoL. 33, No. 3 Literature Cited Arrigoni Degli Oddi, E. and E. Moltoni. 1931. La,ni- dificazione del Nibbio bruno in provincia di Mantova. Riv. Ital. Ornitol. 1:105-128. Arroyo, B. 1978. La alimentacion del milano negro {Mil- vus migrans) en una localidad de Espana central. Ar- deola 25:47-57. Baligliniere, J.L. and H. Le Louarn. 1987. Caracteris- tiques scalimetriques des principales especes de pois- sons d’eau douce de France. Bull. Frangais de la Peche et de la Pisciculture 306:1—39. Belloni, M. 1975. II clima della provincia di Como e Varese in relazione alio studio dei dissesti idrogeolo- gici. Publ. 17, 1st. Geol. Paleontol. Universita di Mil- ano, Milano, Italy. Bianchi, E., L. Martire and A. Bianchi. 1969. Gli uccelli della provincia di Varese (Lombardia). Riv. Ital. Or- nitol. 39:384—401. Bijeleveld, M. 1974. Birds of prey in Europe. Macmillan Press Ltd., London, U.K. Bijlsma, R.G. 1997. Black Rite. Pages 132-133 in W.J.M. Hagemeijer and M.J. Blair [Eds.], The EBCC Atlas of European breeding birds, their distribution and abun- dance. T. & A.D. Poyser, London, U.K. Blanco, G. 1997. Role of refuse as food for migrant, floater and breeding Black Kites (Milvus migrans).]. Raptor Res. 31:71-76. Brown, C.R. and M.B. Brown. 1996. Coloniality in the Cliff Swallow: the effect of group size on social be- havior. Univ. Chicago Press, Chicago, IL U.S.A. Brown, D. 1975. A test of randomness of nest spacing. Wildfowl 26:102-103. Brown, L. and D. Amadon. 1968. Eagles, hawks and fal- cons of the world. Hamlyn Publishing Group, Lon- don, U.K. Bustamante, J. and F. Hiraldo. 1990. Adoptions of fledglings by Black and Red Kites. Anim, Behav. 39: 804-806. AND . 1993. The function of aggressive chases by breeding Black and Red Kites Milvus mig- rans and M. milvus during the post-fledging depen- dence period. Ibis 139:139-147. Ceschi, M., M. De Rossa and M. Jaggli. 1996. Organic and inorganic pollutants and radionuclides in the ich- thyofauna of the Lakes Ceresio and Verbano (Swiss Basins). Trav. Chim. Aliment. Hyg. 87:189-211. Cramp, S. and K.E.L. Simmons. 1980. Handbook of the birds of Europe, the Middle East and North Africa. Vol. 2. Hawks and bustards. Oxford Univ. Press, Ox- ford, U.K. Danko, S. 1989. Five young in the nest of a Black Kite {Milvus migrans). Buteo 4:87-92. De Giacomo, U., O. Martucci and A. Tinelli. 1993. L’alimentazione del Nibbio bruno {Milvus migrans) nella Tenuta di Castelporziano (Roma). Avocetta 17: 73-78. Delibes, M. 1975. Alimentacion del milano negro {Milvus migrans) en Dohana (Huelva, Espana). Ardeola 21: 183-207. Desai, J.H. and A.K. Malhotra. 1979. Breeding biology of the pariah kite Milvus migrans at Delhi Zoological Park. Ibis 121:320-325. Doumeret, a. 1995. Milan noir Milvus migrans. Pages 160-163 in D. Yeatman-Berthelot and G. Jarry [Eds.], Nouvel atlas des oiseaux nicheurs de France, 1985- 1989. Societe Ornithologique de France, Paris, France. FIUCZYNSKI, V. 1981. Berliner milan-chronik {Milvus mig- rans and Milvus milvus). Beitr. Vogelkd. 27:161-196. AND V. Wendland. 1968. Zur Populationsdynamik des Schwarzen milans {Milvus migrans) in Berlin Beo- bachtungen 1952-1967./. Ornithol. 109:462-471. Garcia Ferre, D. and A. deJuan. 1983. Mila negre Milvus migrans. Pages 55-56 in]. Muntaner, X. Ferrer and A Martinez-Vilalta [Eds.], Atlas dels ocells nidificants de Catalunya i Andorra. Ketres Editora, Barcelona, Spain. Gedeon, K. 1994. Monitoring GreifVogel und Eulen. Grundlagen und Moglichkeiten einer Langfristigen Uberwachung von Bestandsgroben und Reproduk- tionsdaten Jahresber. Monitoring Greifvogel Eulen Euro- pas 1:1-118. Geroudet, P. 1965. Les rapaces diurnes et nocturnes d’Europe. Delachaux et Niestle, Neuchatel, Switzer- land. Green, E. 1987. Individuals in an Osprey colony discrim- inate between high and low quality information. Na- ture 829:289-241 . Hagan, J.M. andJ. R. Walters. 1990. Foraging behavior, reproductive success, and colonial nesting in Ospreys. Auk 107:506-521. Heckenroth, V. 1970. Der Greifvogelbestand des Bod- anriicks (Bodensee) 1968 and 1969. Anz. Ornithol. Ges Bayern 9:47-51. Henrioux, P. and J. Henrioux. 1995. Seize ans d’etude sur les rapaces diurnes et nocturnes dans I’Ouest le- manique (1975-1990). Nos Oiseaux 48:\— 2%. Hernandez, L., M.C. Rico, M.J. Gonzalez, M.A. Hernan AND M.A. Fernandez. 1986. Presence and time trends of organochlorine pollutants and heavy metals in eggs of predatory birds of Spain. J. Field Ornithol. 57:270- 282. Hiraldo, F, J.P. Veiga and M. Manez. 1990. Growth of nestling Black kites Milvus migrans: effects of hatching order, weather and season./. Zx)ol. Lond. 222:197-214. Jenni-Eiermann, S., U. Buhler and N. Zbinden. 1996. Vergiftungen von Greifvogeln durch Carbofuranan- wendung im Ackerbau. Ornithol. Beobachter 98:^9-77 . Jones, A.M. and M. Manez. 1990. Cannibalism by Black Kite {Milvus migrans).] Raptor Res. 24:28-29. Koga, K., S. Siraishi and T.A. Uchida. 1989. Breeding ecology of the Black-eared Kite Milvus migrans lineatus September 1999 Black Kite Breeding Ecology 217 in the Nagasaki Peninsula, Kyushu, Ornithol 57; 57-66. Mammen, U. and M. Stubbe. 1995. Jahresbericht 1994 zum monitoring Greifvogeln und Eulen Europas Jah- resber. Monitoring Greifvdgel Eulen Europas 7:1-78. AND M. Stubbe. 1996. Jahresbericht 1995 zum monitoring Greifvogeln und Eulen Europas Jahres- ber. Monitoring Greifvdgel Eulen Europas 8:1-92. Massa, B. 1985. Nibbio bruno Milvus migrans. II Natural- ista Siciliano 9:47. Matthews, J.D. 1989. Silvicultural systems. Oxford Univ. Press, Oxford, U.K. Micheli, a. 1990. Nibbio bruno. Page 57 in P. Brichetti and M. Fasola [Eds.], Atlante degli uccelli nidificanti in Lombardia, 1983-1987. Editoriale Ramperto, Bre- scia, Italy. Newton, I. 1979. Population ecology of raptors. T. & A.D. Poyser, Berkhamsted, U.K. Nore, T. 1979. Rapaces diurnes communs en Limousin pendant la periode de nidification (Buse, Bondre, Milan noir, Busards saint-martin et Cendre). Alauda 47:183-194. Norusis, M.J. 1993. SPSS Advanced statistics. McGraw- Hill, New York, NY U.S.A. Petretti, F. 1992. Nibbio bruno. Pages 459-465 in P. Bri- chetti, P. De Franceschi and N. Baccetti [Eds.], Fauna d’ Italia. Edizioni Calderini, Bologna, Italy. Petretti, A. and F. Petretti. 1981. A population of di- urnal raptors in central Italy. Gci/aw# 71:143— 156. Pinna, M. 1978. L’atmosfera e il clima. U.T.E.T., Torino, Italy. Richard, A. 1934. Le Milan noir. Nos Oiseaux 117:298- 306. Seelig, K., H. Benecke, F. Braumann and B. Nicolai. 1996. Die Vogel im Naturpark Dromling. Abh. Ber. Mus. Heineanum 3, Sonderh, Germany. Sermet, E. 1980. Milan noir. Pages 88-89 in A. Schifferli, P. Geroudet and R. Winkler [Eds.], Atlas des oiseaux nicheurs de Suisse. Station Ornithologique de Sem- pach, Sempach, Switzerland. Stamps, J. A. 1988. Conspecific attraction and aggregation in territorial species. Am. Nat. 131:329-347. Steenhof, K. 1987. Assessing raptor reproductive success and productivity. Pages 157-170 in B.A. Giron Pen- dleton, B.A. Millsap, K.W. Kline and D.M. Bird [Eds ], Raptor management techniques manual. Natl. Wildl Fed., Washington, DC U.S.A. Steinmetz, B. and R. Muller. 1939. An atlas of fish scales and other bony structures used for age determina- tion; non-salmonid species found in European fresh waters. Samara Publ., Cardigan, U.K. Stubbe, V. 1961. Die Besiedelungsdichte eines Abges- chlossenen Waldgebietes (Hakel) mit Greifvogeln im Jahre 1957. Beitr. Vogelkd. 7:155-224. Thiollay, J.-M. 1967. Ecologie d’une population de ra- paces diurnes en Lorraine. La Terre et la Vie 114:116- 183. ViNUELA, J. 1993. Variacion en la fecha de puesta de una poblacion de milano negro {Milvus migrans). Efecto de la experiencia de los reproductores. Ardeola 40:55- 63. and G.P. Veiga. 1992. Importance of rabbits in the diet and reproductive success of Black Kites in southwestern Spain. Ornis Scand. 23:132-138. AND C. Sunyer. 1994. Black Kite Milvus migrans. Pages 148-149 in G.M. Tucker and M.F. Heath [Eds.], Birds in Europe; their conservation status. BirdLife Conserv. Sen No. 3, Cambridge, U.K. , R. Villafuerte and C. De Le Court. 1994. Nest- ing dispersion of a Black Kite population in relation to location of rabbit warrens. Can. J. Zool. 72:1680- 1683. Received 29 August 1998; accepted 3 April 1999 J Raptor Res. 33(3) :218-226 © 1999 The Raptor Research Foundation, Inc. STATUS OF NESTING BALD EAGLES IN ARIZONA Daniel E. Driscoll, Ronald E. Jackman and W. Grainger Hunt Predatory Bird Research Group, Long Marine Laboratory, University of California, Santa Cruz, CA 95060 U.S.A. Greg L. Beatty, James T. Driscoll and Richard L. Glinski Arizona Game and Fish Department, 2221 Greenway Road, Phoenix, AZ 85203 U.S.A Thomas A. Gatz and Robert L Mesta^ us. Fish and Wildlife Service, 2321 West Royal Palm Road, Suite 103, Phoenix, AZ 85021 U.S.A. Abstract. — ^A small, isolated population of Bald Eagles (Haliaeetus leucocephalus) breeds along the desert rivers of central Arizona. The extent to which it followed the continental pattern of decline during the DDT era cannot be known because of the paucity of data before 1970 but, from 1970-93 the number of known breeding pairs increased from two to 34. Some of this growth was an artifact of increased surveys, but much was real. During 1987—93, territory occupancy rate was 90%, a higher than normal value for the species, but nest success (45%) and productivity (0.69 young per occupied site, SE = 0.08) were lower than reported for other populations in the coterminous U.S. Much of the annual variation in nesting success and productivity resulted from heat stress and flooding, factors that impacted many breeding areas simultaneously. We recorded 41 breeder fatalities or replacements during 262 monitored breeder/years, and additional fatalities and replacements likely went unrecorded. If replacements equat- ed to fatalities, estimated maximum annual breeder survival was 0.84 (95% C.l. 0.78-0.88), a lower value than elsewhere estimated. Of 131 monitored pairings during the seven years of our research, 24 (18%) contained subadults. This higher than normal rate for Bald Eagles suggests a paucity of floaters (non- territorial adults) , although other factors may be involved. Notwithstanding these suggestions of reduced demographic potential, a continuing upward trend in the number of territories is apparent. Key Words: Bald Eagle; Haliaeetus leucocephalus; Arizona', nesting, population status; productivity, breeder turnover, subadult recruitment. El estatus de aguilas calvas anidando en Arizona Resumen. — Una pequena y aislada poblacion de aguilas calvas {Haliaeetus leucocephalus) se reproduce a lo largo de los rios del desierto en el centro de Arizona. Hasta donde esta poblacion siguio el patron continental de disminucion poblacional durante la era del DDT, no se ha podido comprobar debido a la falta de informacion existente antes de 1970. Entre 1970-93 el numero de parejas reproductivas conocidas aumento de dos a 34. Parte de este aumento fue sobreestimado. Durante 1987-93 la tasa de ocupacion del territorio fue del 90%, un valor mayor del normal para la especie, el exito de anidacion fue del 45% y la productividad de 0.69 juveniles por sitio ocupado, SE = 0.08; Lo cual fue mas bajo que lo reportado para otras poblaciones de Estados Unidos. Estas variaciones fueron atribuidas al im- pacto del estres causado simultaneamente por el calor y las inundaciones en varies sitios de reproduc- cion. Registramos 41 fracases de los reproductores o remplazos en 262 monitoreos de reproduccion por ano. Creemos que hubo mas fracases que no pudieron ser registrados. Si los remplazos fueron iguales a los fracasos, la sobrevivencia anual de reproduccion fue de 0.84 (95% C.L 0.78-0.88), un valor mas bajo de lo estimado en otros lugares. De las 131 parejas monitoreadas durante los 7 afios de nuestra investigacion, 24 (18%) contenian individuos subadultos. Estas cifras son mas altas de lo normal para las aguilas calvas. Lo anterior pudo ser ocasionado por el estancamiento de los individuos “floaters” (adultos sin territorio) , aunque otros factores pudieron ser los causantes de esta situacion. A pesar de esto y sin subestimar esta situacion de reducido potencial productive, se observa una tendencia creciente de territories ocupados. [Traduccion de Cesar Marquez] ^ Present address: U.S. Fish and Wildlife Service, 2493 Portola Road, Suite B, Ventura, CA 93003 U.S.A. 218 September 1999 Status of Bald Eagles in Arizona 219 / / / i Figure 1. The distribution of breeding Bald Eagles in Arizona. Arizona supports a nesting population of Bald Eagles {Haliaeetus leucocephalus) primarily along the Salt and Verde Rivers in the central part of the state (Fig. 1). This small population is somewhat isolated from other centers of breeding activity and occupies habitat drier, warmer, and less vegetated than is typical for the species. Large trees are scarce along the desert rivers and many pairs use cliff nests. The population has been classified as threatened by the U.S. Fish and Wildlife Service (USFWS 1995), and there is concern that the rap- idly increasing human population in Arizona may impact its numbers. Herein, we summarize data on occupancy and productivity, and discuss factors af- fecting survival and recruitment. Drawing upon historical records and reports, and upon field data obtained during 1987-93, we examine the evi- dence for population self-sustainability. History of Occurrence When the USFWS began monitoring nesting Bald Eagles in Arizona in 1970, only two occupied territories were known, both on the Verde River (Rubink and Podborny 1976). From 1970-93, the number of known territories increased to 34. To what extent this increase resulted from greater numbers of eagles is unknown, although it was clear that many territories had been newly estab- lished. Others may have been long occupied and simply found by greater search efforts and more extensive surveys, an uncertainty that has affected studies of other populations (Sprunt et al. 1973, Henny and Anthony 1989, Frenzel 1991). Newly discovered sites we examined often contained nu- merous old nests, but whether these were recently occupied or were remnants of many years past could not be determined since stick nests on cliffs in Arizona may persist for many decades. Also uncertain, because of the paucity of data before 1972, is the extent to which the Arizona population followed the continental pattern of de- cline during the DDT era. DDT was applied exten- sively to cotton and vegetable crops in Arizona and adjacent regions, and the highest DDE levels re- corded during the 1967-79 National Starling Mon- itoring Program were from nearby Maricopa, Ari- zona and Chaves, New Mexico (Hunt et al. 1986). Prey fish probably contained very little DDE because relatively few farms existed within the drainages occupied by breeding eagles; however, migratory and nomadic waterbirds were likely con- taminated. There is evidence for the existence of a nesting population of Bald Eagles in Arizona prior to 1970. Early sightings of individual eagles were reported by Coues (1866), Henshaw (1875), Willard (1916) and Hargrave (1939). Mearns (1890) described the first breeding record, a tree nest near Stoneman’s Lake on the Mogollon Plateau. Bent (1937) re- ported breeding pairs at Fort Whipple and the Salt River Bird Reservation, an area flooded by Roose- velt Reservoir in 1911. According to Jenks and Ste- venson (1937), Bald Eagles bred in the White Mountains and were resident in central eastern Ar- izona, primarily in the Transition and Upper Son- oran zones. Cited records of two nests in saguaro cacti ( Cereus giganteus) on the lower Verde River in 1937 and a nesting pair on the Little Colorado Riv- er in 1951. Phillips et al. (1964) reported nesting records from the 1930-60s on the Salt and Verde rivers, including pairs breeding during 1930-36 near Saguaro Reservoir (constructed during 1928— 30), since the early 1940s near Bartlett Reservoir (completed in 1939), and in the free-flowing Salt River Canyon in 1935, 1944, and 1949. C. Mc- Collough (field notes) observed nesting on the Tonto Arm of Roosevelt Reservoir from 1951-55. 220 Driscoll et al. VoL. 33, No. 3 Rubink and Podborny (1976) described a nest on the upper Verde River in the 1960s and a pair near the mouth of the Verde River in 1968. L. Forbis (U.S. Forest Service, unpublished interview with F. Thompson 1979) noted three nests on the upper Verde River occupied during the early- to mid- 1960s. G. Gibbons (pers. comm.) observed nesting at Canyon Reservoir on the Salt River in the 1950s and 1960s. Study Area In Arizona, Bald Eagles breed mainly in open, desert landscapes of the Upper and Lower Sonoran Life-Zones (Lowe 1964) at elevations ranging from 329-1719 m. An- nual precipitation averages from 39 cm at higher eleva- tions to 25 cm in desert scrub habitats, where tempera- tures may reach 50°C. Livestock grazing, particularly during the close of the 19th century, created conditions leading to extreme loss of soils and vegetation, including formerly extensive riparian forests (Hayden 1965, Has- tings 1959, Hastings and Turner 1965, Davis 1982). The resulting need for flood control and water storage prompted the construction of dams in the early 1900s (Hayden 1965). Considerable human recreational activity occurs along the water courses. Some Bald Eagle terri- tories are easily accessible, while others are in remote canyons. Reservoirs and riverine sections contain both native and introduced fishes. Species most commonly eaten by Bald Eagles are catfish {Ictalurus punctatus and Pylodictis ohvaris), sucker {Catostomus clarki and C. insignis), com- mon carp {Cyprinus carpio), and perciforms {Pomoxis ni- gromaculatus, Morone mississippiensis and Micropterus salmo- ides) (Haywood and Ohmart 1986, Grubb 1995, Hunt et al. 1992). Of those listed, only the suckers are native to Arizona. In winter, waterfowl are important prey of the breeding population, especially at reservoirs. Methods During 1987-93, we collected occupancy and repro- ductive data both on foot and by monthly helicopter flights throughout the breeding season (January-June). Data recorded during the flights included the number of adults seen, location of new nests, number of nestlings and their approximate age. At many territories, nest war- dens of the Arizona Bald Eagle Nest Watch Program (ABENWP) reported the exact dates of egg laying, hatch- ing and fledging, and recorded observations of distur- bance and fatalities (Forbis et al. 1985). Status terminol- ogy followed Postupalsky (1974). Using techniques described in Hunt et al. (1992) and Jackman et al. (1993, 1994), we radiotagged 15 juveniles, eight subadults and 12 breeders at 10 territories. We monitored survival of radio-tagged eagles by airplane. Breeder replacement was determined on the basis of identifying characteristics (e.g., bands on right versus left tarsi, adults preceding subadults, and known deaths of unmarked pair members [Gould and Fuller 1995]). Except in cases where our ac- tivities might have jeopardized nestling survival, we en- tered all occupied nests during 1987-93 and banded all eaglets surviving to 6 wk of age {N = 119) with standard USGS tarsal bands and similarly-sized, color-anodized alu- minum visual identification (VID) bands engraved with various symbols. We recorded standard body measurements of breeding eagles as described by Bortolotti (1984a, 1984b) and Gar- celon et al. (1985), including hallux length, culmen length, beak depth, tarsus width, length of tail and eighth primary, and weight. A cloth tape was used to measure wing chord length across the dorsum. We compared the means for each measurement (Rest) from Arizona (10 males, 4 females), northern California (9 males, 11 fe- males) and the Greater Yellowstone Ecosystem in Wyo- ming and Montana (12 males and 6 females; A. Harmata unpubl. data). Results and Discussion Reproduction. Over the 7-year period of our study (1987-93), the occupancy rate in central Ar- izona was 90% (183 of 204 known territory/ years) as compared to a mean of 71 % for numerous other North American populations reviewed hy Stalmas- ter (1987). Nest success in Arizona was 45%, mean brood size 1.5 young and productivity 0.69 young (SE ± 0.08) per occupied site (Table 1). Except for brood size, these figures are the lowest report- ed in a sample of nine other populations in the coterminous U.S. for roughly the same period (Ta- ble 2). On average, only 11.9 (range = 9-16) of the 23- 30 pairs were productive in any one year (Table 1) . For the most part, the low productivity resulted not from failure to lay eggs but from loss of eggs or young. These results were surprising in view of the nest protection and frequent enhancement of nest- ling survival by the ABENWP, active since 1978. Young blown from nests by high winds and those threatened by impending inundation or the loss of a parent were replaced or fostered into other nests (Grubb 1984). Egg Mortality. During our 7-yr study, 73 (31%) of 237 known eggs perished or were infertile, a figure comparable to the 25% reported by Stal- master (1987) for other North American popula- tions. Of 24 eggs (in 13 clutches) for which we knew the mortality cause, 11 involved human dis- turbance, four (two clutches) were lost to nest in- undation by filling reservoirs, six (three clutches) to a polygynous relationship where the male assist- ed only his primary mate, two when a female ceased incubating because of apparent physiologi- cal distress, and one involving an exceptionally small egg. Seven addled eggs we collected during 1986-89 averaged 5.5 ppm DDE (range = 2. 3-9.5) and 2.2 September 1999 Status of Bald Eagles in Arizona 221 Table 1. Known productivity at Arizona bald eagle breeding areas during 1970-93.^ Terminology follows Postupalsky (1974). Year Known Breeding Areas Occupied Nests Successful Nests Young Fledged % Nest Success Mean Brood Size Productivity 1970 2 2 2 3 100.0 1.5 1.50 1971 3 3 3 4 100.0 1.3 1.33 1972 4 3 0 0 0.0 0.0 0.00 1973 6 5 5 7 100.0 1.4 1.40 1974 8 5 3 6 60.0 2.0 1.20 1975 9 7 4 5 57.0 1.2 0.71 1976 9 5 4 7 80.0 1.7 1.40 1977 9 7 3 6 42.9 2.0 0.86 1978 11 10 6 9 60.0 1.5 0.90 1979 12 10 4 6 40.0 1.5 0.60 1980 12 9 3 5 33.3 1.7 0.56 1981 13 10 8 16 80.0 2.0 1.60 1982 14 13 8 14 61.5 1.7 1.07 1983 15 13 7 13 53.8 1.9 1.00 1984 18 17 8 15 47.1 1.9 0.88 1985 20 19 13 22 68.4 1.7 1.16 1986 21 16 12 17 75.0 1.4 1.06 1987 26 24 11 20 45.8 1.9 0.83 1988 27 23 15 24 65.2 1.6 1.04 1989 28 25 9 13 36.0 1.4 0.52 1990 28 26 9 14 34.6 1.6 0.54 1991 29 26 13 20 50.0 1.5 0.77 1992 32 30 10 14 33.3 1.4 0.47 1993 34 29 16 22 55.2 1.4 0.76 ' Data from 1970 through 1985 from Rubink and Podborny (1976), Hildebrandt and Ohmart (1978), Hildebrandt (1981), Ohmart and Sell (1980), Haywood and Ohmart (1980, 1981, 1982, 1983), Grubb et al. (1983), Grubb (1984, 1986). Table 2. Productivity comparison of bald eagle populations within the coterminous United States over a similar period. Region Period Occupied Nests Percent Nest Successes Mean Brood Size Productivity Source Florida 1988 399 69 ? 1.1 Wood et al. 1990 Louisiana 1988 36 67 ? 1.1 Wood et al. 1990 Texas, Southeast 1981-90 193 64 1.5 0.98 Mabie et al. 1994 Colorado + Wyoming 1981-89 85 63 1.9 1.21 Kralovec et al. 1992 South Carolina 1988 50 82 ? 1.4 Wood et al. 1990 Virginia 1988 84 77 ? 1.4 Wood et al. 1990 Chesapeake Bay 1981-90 1448 70 1.7 1.21 Buehler et al. 1991 Maryland 1988 97 79 ? 1.4 Wood et al. 1990 California 1987-93 69 58 1.6 0.90 Jackman (unpubl. data) Arizona 1987-93 183 45 1.5 0.69 (this study) 222 Driscoll et al. VoL. 33, No. 3 ppm PCBs (range = 1.2-3. 9) wet weight (Jenkins et al. 1994). No DDE effect on productivity was apparent in the breeding areas from which eggs were obtained, although Wiemeyer et al. (1984) reported that DDE levels of this magnitude were associated with a 50% depression of productivity in a 14-state study during 1969-79. A mean thinning rate of 6.9% for 156 eggshell fragments (including membrane) was well below the 10% level associat- ed with reduced productivity in other bald eagle populations (Wiemeyer et al. 1984, Nisbet 1989). These findings are similar to those reported by Grubb et al. (1990) for Bald Eagles in Arizona from 1977-85. Nestling Fatalities. During our study, 37 (22.6%) nestlings died among the 164 known to have hatched, as compared with 15% reported by Stal- master (1987) for other populations. However, few studies were likely to have monitored nests as close- ly as those in Arizona where many were watched continuously throughout the breeding season. We recorded an additional six fatalities around the time of first flights from the nest. Weather-related factors were the apparent cause of death in 23 (53%) of the 43 total cases. In 1988 and 1989, heat stress appeared to be the primary agent in the death of 11 young (see Nelson 1969, Beecham and Kochert 1975, Hayes and Gessaman 1980). During 1992 and 1993, heavy rains and flood conditions were the apparent cause of 11 fatalities and the destruction of 10 nests. Additional causes of death included Mexican chicken bug (Haematosiphon in- odorus) infestations {N = 5) (see also Grubb et al. 1986), falling from (or being blown out of) nests (A = 4), human disturbance before self-thermo- regulation (A = 3), Great Horned Owl {Bubo vir- gtnianus) predation (A = 2), entanglement in monofilament fishing line (A = 2), and bacterial infections ( A = 2) . Heavy rains not only directly caused nesding mortality, they also transformed rivers and creeks into muddy torrents, reducing fish availability (Grubb 1995), Only two of 14 breeding pairs on free-flowing river reaches successfully raised young in 1992, a flood year. We examined prey remains from the two successful nests and found signifi- cantly higher mammalian numerical composition (X^ = 18.4, df = 1, P < 0.005 and x^ = 7.4, df = 1, P < 0.01) compared to remains we collected during 1986-90 from these same nests. Breeder Fatalities and Turnover. Thirteen fatal- ities were documented for Arizona Bald Eagle breeders from 1951-93. Two were killed by other Bald Eagles, one by a Golden Eagle {Aquila chry- saetos), one by a Peregrine Falcon {Falco peregrinus) , five were shot, one died of impact injuries, and three died of unknown causes. We infer an addi- tional 46 breeder fatalities (24 males, 21 females, 1 unknown) from replacement of pair members (Gould and Fuller 1995). No marked breeder was seen after its replacement. The two oldest known breeders were 14 years old in 1993, with site ten- ures of 11 and 10 years. Considering only the 1987-93 period, we record- ed 41 fatalities or replacements during 262 moni- tored breeder/years. If replacements equate to fa- talities, the (maximum) point estimate of breeder survival becomes 0.84 (95% C.I. — 0.78—0.88) (Trent and Rongstad 1974), a figure somewhat low- er than that reported in three other studies (i.e., 0.88 by Bowman et al. (1995) in Alaska, 0.93 by Hodges et al. (1987) in Alaska and 0.93 by Gerrard et al. (1992) in Canada. Because some deaths and replacements in Arizona likely went undetected, our estimate was doubtless biased toward higher survival, as may have been the case with the esti- mate of Gerrard et al. (1992). The estimates for Alaska were based on samples of radio-tagged ea- gles. Breeding Eagles in Subadult Plumage. Bald ea- gles normally defer first breeding until at least 5- 8 years of age (Hansen and Hodges 1985, Buehler et al. 1991, Gerrard et al. 1992, Bowman et al. 1995). Healthy raptor populations typically contain many floaters unable to obtain breeding territories because suitable habitat is saturated by breeding pairs (Newton 1979). However, during the seven years of our study, the Arizona population ap- peared to contain relatively few floaters, as evi- denced by a high incidence of breeding eagles dis- playing subadult (nondefinitive) characteristics (24 of 131 monitored pairings, 18.3%). This and the incidence of replacements by young eagles (>44%) were far higher than reported elsewhere (Herrick 1924, Bent 1937, Gerrard et al. 1978, Sherrod et al. 1976, Stalmaster 1987, Mabie et al. 1994, Anthony et al. 1994). In a long-term study involving thousands of Bald Eagle pairings in Sas- katchewan, Gerrard et al. (1983) noted only two subadults paired with full-adults and neither pair laid eggs. Among hundreds of pairs at Chesapeake Bay, Buehler et al. (1991) observed no subadults and Hansen and Hodges (1985) remarked that subadults were not known to breed in southeastern September 1999 Status of Bald Eagles in Arizona 223 Alaska. Swenson et al. (1986) noted five subadults paired with full-adults in 11 years of data for 40- 50 pairs in the Greater Yellowstone Ecosystem (1972-82); only two such pairs repaired a nest or laid eggs and none reproduced. Some of the pairings and replacements in Ari- zona by eagles in subadult plumage may reflect a possibly atypical delay in the acquisition of full adult (definitive) plumage. J. Driscoll noted that six of 23 eagles of known age (5 females, 1 male) retained eye-stripes at greater than five years of age, and two of these still displayed eye-stripes at seven years of age. McCollough (1986) reported that two of 13 individuals in Maine retained eye- stripe remnants at six to eight years. Pairs with members displaying subadult plumage in Arizona were more productive than those in oth- er studies. Among the 40 pairings containing sub- adults at 35 occupied nests since 1970, 16 (46%) successfully fledged young, three (9%) hatched young that died in the nest, four (11%) laid eggs that did not hatch, and the remaining 12 (34%) apparently did not lay eggs. Of the five pairs with both members in subadult plumage, three appar- ently did not lay eggs and two hatched two young each. Although significantly more full-adult pairs laid eggs than did subadult or mixed pairs (x^ = 8.0, df = 1, P < 0.005), we found no significant difference in nest success or productivity. Relationship to Other Populations. There is some evidence that the Arizona Bald Eagle popu- lation may be augmented by immigration from other regions. We were able to read the bands of 14 of 29 banded breeders during 1987-93. All 14 had been banded as nestlings in Arizona. Six of the unidentified birds were banded on the left tar- sus, but according to banding records, only four Arizona nestlings had been so banded. Thus, at least two breeders may have originated from out- side Arizona. In 1994, an eagle banded as a nest- ling by Mabie et al. (1994) in southeast Texas (1410 km away) bred at a high altitude reservoir (2438 m) in the pine-forested mountains of east- ern Arizona near the New Mexico border. Arizona adults were, on average, smaller in all morphological characteristics than those from Alaska, California, or the Greater Yellowstone re- gion, a finding that supports an affinity of the Ar- izona population with those in the southeastern U.S. as proposed by Amadon (1983), Stalmaster (1987) , and Palmer (1988) . Arizona males were sig- nificantly smaller in 13 of 16 measurements (Table 3) and females in eight of 16 comparisons, a dif- ference probably arising from the small number of females sampled. Morphological measurements of two adults from the Rio Yaqui, Sonora, Mexico did not differ from those in Arizona. These results suggested that the Bald Eagle pop- ulation in the southwest either rebounded from a depleted population or derived from immigration from the southeastern U.S., or both. If gene flow into Arizona from north or west (where eagles are larger) had recently occurred, it should at least be reflected in the overall variance of mensural char- acters, assuming that the environmental compo- nent of such variation is weak compared with the genetic component. We examined this hypothesis by comparing coefficients of variation for the var- ious characters we measured in the four popula- tions, but found no suggestion of greater variance within the Arizona sample. Summary of Status Despite the uncertainty of determining whether newly discovered pairs equate to newly occupied territories, the weight of evidence indicates a con- tinuing upward trend in the number of Bald Eagle pairs in Arizona since 1970. This trend is occurring with little direct evidence of augmentation by im- migrants and despite relatively low estimates of productivity and breeder survival. The substantial incidence of pairs containing members in subadult plumage suggests a deficient age structure in the floating segment and thus a relatively small recruit- ment buffer (Hunt 1998). A paucity of floaters may, in part, result from continued population ex- pansion (i.e., eagles that would otherwise become floaters are founding new territories) . However, we do not discount the role of low vital rates (e.g., breeder survival) in retarding floater accrual. We therefore recommend monitoring of breeding sites and continuation of population-wide VID banding. Because of the proven benefit of the ABENWP in enhancing reproduction, we recom- mend its continuance at sites where human distur- bance is considered significant. Acknowledgments Funding for this study and its publication was provided by the U.S. Bureau of Reclamation (USBR) with assis- tance in some years from the Arizona Game and Fish Department. We thank H. Messing and D. Busch (USBR) for help with the project. The USBR and Salt River Pro- ject (SRP) provided helicopters and pilots B. Barnes and D. Blakely for survey flights, nestiing emergencies, and Table 3. Means and standard deviations for selected measurements (mm) and weights (kg) of adult male Bald Eagles from Arizona, northern California, Alaska (data from T. Bowman and P. Schempf) , and the Greater Yellowstone Ecosystem (Wyoming and Montana, data from A. Harmata) . 224 Driscoll et al. VoL. 33, No. 3 z 0 H 1 0 J * * * * * * CH * * * * * * 00 Oi H 2 p ^ ^ ^ ^ m 00 d +1 +1 o oq rH CO + 1 +1 q cq GO rH i> m CH oo d +1 * * * * * * * * GO ^ GO GO GO 00 CO GO GO M II II II 00 r— I O +1 +1 +1 ^ GO o +1 +1 GO M q GO 00 GM 1> C>J lO ■X- ^ 'Sfr ji * * * 2 * * •:^‘ * y— s. y — V y— s ^ 05 .. 05 ^ II II ^ M II 'gH d + 1 d GO d ^ d d O rH rH + 1 +1 +1 +1 9 d d rH GO iO .P r-H CO m ^ O ^ 00 05 CM q q rH <d CD d +1 G GO + 1 +1 Tf O J> GO (M d oj ^ d +1 +1 9 00 00 GO u (L) 4 -J gS 'u -5 ^ :2 X "rt w: 3 VI Sh be 3 (U 3 U be • ^ JO o o o V * * * q d V * X- in o d V CL. 4f G O N </! 0) G O 'u CIh E o access to remote breeding areas for nest climbs. We thank M. Greenburg and T. Noble (SRP) and C. Thelan- der (BioSystems Analysis, Inc., BSAI) and numerous BSAI employees for their assistance during our study. We acknowledge the work of D. Rubink (USFWS), R. Ohmart, T. Hildebrandt, D. Haywood and R. Sell (Ari- zona State University), T. Grubb, W. Eakle and L. Forbis (USFS), and the hundreds of ABENWP personnel who have watched over Arizona’s nesting eagles since 1978. We acknowledge D. Bland, L. Kiff, and R. Risebrough for work concerning eggshell thinning and contaminants. We thank J. Linthicum, T. Hunt, and two anonymous re- viewers for valuable comments on the manuscript. Literature Cited Amadon, D. 1983. The Bald Eagle and its relatives. Pages 1-4 in D.M. Bird, N.R. Seymour and J.M. Gerrard [Eds.], Biology and management of Bald Eagles and Ospreys. Harpell Press, MacDonald Raptor Research Center of McGill Univ., Montreal, Quebec, Canada Anthony, R.G, R.W. Frenzel, F.B. Isaacs and M.G Garrett. 1994. Probable causes of nesting failures in Oregon’s Bald Eagle population. Wildl. Soc. Bull. 22’ 376-582. Beecham, J.J. AND M.N. Kochert. 1975. Breeding biology of the Golden Eagle in southwestern Idaho. Wilson Bull. 87:506-513. Bent, A.C. 1937. Life histories of North American birds of prey. Part 1. Smithsonian Institution, U.S. Nat. Mus Bull. 167. U.S. Government Printing Office, Washing- ton, DC U.S.A. Bortolotti, G.R. 1984a. Sexual size dimorphism and age-related size variation in Bald Eagles./. Wildl. Man- age. 48:72—81. . 1984b. Criteria for determining age and sex of nestling Bald Eagles. J. Field Ornithol. 55:467-480. Bowman, T.D., P.F. Schempf andJ.A. Bernatowicz. 1995 Bald Eagle survival and population dynamics in Alas- ka after the Exxon Valdez oil spill. J. Wildl. Manage 59:317-324. Buehler, D.A., J.D. Fraser, J.K.D. Seeger, G.D. Therres AND M.A. Byrd. 1991. Survival rates and population dynamics of Bald Eagles on Chesapeake Bay. /. Wildl. Manage. 55:608-613. CouES, E. 1866. List of the birds of Fort Whipple, Ari- zona. Proc. Acad. Nat. Sd. Phila. 18:39-100. Davis, G.P. 1982. Man and wildlife in Arizona: the Amer- ican exploration period, 1824—65. Arizona Game and Fish Dept., Phoenix, AZ U.S.A. Forbis, L.A., T.G. Grubb and W.D. Zeedyk. 1985. “Eagle Beagles”; a volunteer Bald Eagle nest watcher pro- gram on Arizona national forests. Pages 246-254 in J.M. Gerrard and T.N. Ingram [Eds.], The Bald Eagle in Canada, Proc. Bald Eagles Days, 1983. White Horse Plains Publishers, Headingly, Manitoba, Canada. Frenzel, L.D. 1991. Bald Eagle and Osprey. Pages 23—31 in Proceedings of the Midwest raptor management September 1999 Status of Bald Eagles in Arizona 225 symposium and workshop. Nad. Wildl. Fed., Washing- ton, DC U.S.A. Garcelon, D.K., M.S. Martell, RT. Redig and L.C. Buoen. 1985. Morphometric, karyotypic and laparo- scopic techniques for determining sex in Bald Eagles. /. Wildl. Manage. 49:595-599. Gerrard, J.M., D.W.A. Whitfield, P. Gerrard, P.N. Ger- RARD AND W.J. Maher. 1978. Migratory movements and plumage of subadult Saskatchewan Bald Eagles. Can. Field-Nat. 92:375-382. , P.N. Gerrard, G.R. Bortolotti and D.W.A. Whitfield. 1983. A 14-year study of Bald Eagle repro- duction on Besnard Lake, Saskatchewan. Pages 47-57 m D.M. Bird, N.R. Seymour andJ.M. Gerrard [Eds.], Biology and management of Bald Eagles and Ospreys. Harped Press, MacDonald Raptor Research Center of McGill Univ., Montreal, Quebec, Canada. , P.N. Gerrard, G.R. Bortolotti and E.H. Dzus. 1992. A 24-year study of Bald Eagles on Besnard Lake, Saskatchewan./. Raptor Res. 26:159-166. Gould, W.R. and M.R. Fuller. 1995. Survival and pop- ulation size estimation in raptor studies: a comparison of two methods./. Raptor Res. 29:256-264. Grubb, T.G. 1984. 1984 Arizona Bald Eagle research re- port. USES, Rocky Mountain Forest and Range Ex- periment Station, Tempe, AZ U.S.A. . 1986. Arizona Bald Eagle research 1983-85. USES, Rocky Mountain Forest and Range Experiment Station, Tempe, AZ U.S.A. . 1995. Food habits of Bald Eagles breeding in the Arizona desert. Wilson Bull. 107:258-274. , W. Ea ki.e and J.L. Ganey. 1983. 1983 Arizona Bald Eagle research. USES, Rocky Mountain Forest and Range Experiment Station, Tempe, AZ U.S.A. , W.L. Ea rl e and B.N. Tuggle. 1986. Haematosi- phon inodorus (Hemiptera: Cimicidae) in a nest of a Bald Eagle in Arizona./. Wildl. Diseases 22:125-127. , S.N. Wiemeyer and L.E. Kief. 1990. Eggshell thin- ning and contaminant levels in Bald Eagle eggs from Arizona, 1977-85. Southwest. Nat. 35:298-301. Hansen, A.J. and J.L Hodges. 1985. High rates of non- breeding Bald Eagles in southeastern Alaska. /. Wildl. Manage. 49:454—458. Hargrave, L.L. 1939. Winter bird notes from Roosevelt Lake Arizona. Cowrfor 41:121-123. Hastings, J.R. 1959. Vegetation change and arroyo cut- ting in southeastern Ariz./ Ariz. Acad. Sci. 1:60-67. and R.M. Turner. 1965. The changing mile. Univ. Arizona Press, Tucson, AZ U.S.A. Hayden, C. 1965. A history of the Pima Indians and the San Carlos Irrigation Project. Compiled in 1924, 89th Congress, First Session, Senate Document 11. U.S. Government Printing Office, Washington DC U.S. A. Hayes, S.R. and J.A. Gessaman. 1980. The combined ef- fects of air, temperature, wind and radiation on the resting metabolism of avian raptors. /. Therm. Biol. 5: 119-125. Haywood, D.D. and R.D. Ohmart. 1980. Study of the breeding biology of Bald Eagles in Arizona: status re- port. USFWS, Phoenix, AZ U.SA. and . 1981. Study of the breeding biology of Bald Eagles in Arizona: 1979-80. USFWS, Phoenix, AZ U.S.A. and . 1982. Study of the breeding biology of Bald Eagles in Arizona: 1981. USFWS, Phoenix, AZ U.S.A. AND . 1983. Study of the breeding biology of Bald Eagles in Arizona: 1982. USFWS, Phoenix, AZ U.S.A. AND . 1986. Utilization of benthic-feeding fish by inland breeding Bald Eagles. Condor 88:35-42. Henny, CJ. AND R.G. Anthony. 1989. Bald Eagle and Os- prey. Pages 66-82 in Proceedings of the western rap- tor management symposium and workshop. Natl Wildl. Fed., Washington DC U.S.A. Henshaw, H.W. 1875. Annotated list of the birds of Ari- zona. /n U.S. Geographic and Geological Survey west of the One-hundredth Meridian, by Lieut. George M Wheeler. U.S. Government Rep., Washington, DC U.S.A. Herrick, F.A. 1924. Nests and nesting habits of the Amer- ican eagle. Auk 41:213-231. Hildebrandt, T.D. 1981. The ecology of breeding south- ern Bald Eagles in Arizona, 1977 and 1978. M.S. the- sis, Arizona State Univ., Tempe, AZ U.S.A. AND R.D. Ohmart. 1978. Arizona breeding Bald Eagle investigations: 1977 annual report. USES, Phoe- nix, AZ U.S.A. Hodges, J.L, E.L. Boeker and A.J. Hansen. 1987. Move- ments of radio-tagged Bald Eagles in and from south- eastern Alaska. Can. Field-Nat. 101:136—140. Hunt, W.G. 1998. Raptor floaters at Moffat’s equilibrium. Oikos 82:191-197. , B.S. Johnson, C.G. Thelander, B.J. Walton and R.W. Risebrough. 1986. Environmental levels of p,p'- DDE indicate multiple sources. Environ. Toxicol. Chem. 5:21-27. , D.E. Driscoll, E.W. Bianchi and R.E. Jackman. 1992. Ecology of breeding Bald Eagles in Arizona. USBR, Boulder City, NV U.S.A. Jackman, R.E., W.G. Hunt, D.E. Driscoll and F. Lapan- SKY. 1994. Refinements to selective capture techniques for Bald and Golden Eagles./. Raptor Res. 28:268-273. , , , AND J.M. Jenkins. 1993. A modified floating-fish snare for capture of inland Bald Eagles. N. Am. Bird Bander 18:98-101. jENpNS, J.M., R.M. Jurek, D.K. Garcelon, R. Mesta, W.G. Hunt, R.E. Jackman, D.E. Driscoll and RW Risebrough. 1994. DDE contamination and popula- tion parameters of Bald Eagles (Haliaeetus leucocephal- us) in California and Arizona, U.S.A. Pages 751-756 in B.-U. Meyburg and R.D. Chancellor [Eds.], Raptor conservation today. World Working Group on Birds of Prey and Owls. Pica Press, London, U.K. 226 Driscoll et al. VoL. 33, No. 3 JENKS, R. AND J.O. Stevenson. 1937. Bird records from central-eastern Arizona. Cowrfor 39:87-90. Kralovec, M.L., R.L. Knight, G.R. Craig and R.G. Mc- Lean. 1992. Nesting productivity, food habits and nest sites of Bald Eagles in Colorado and Southeastern Wy- oming. Southwest. Nat. 37:356-361. Lowe, C.H. 1964. The vertebrates of Arizona. Univ. Ari- zona Press, Tucson, AZ U.S.A. Mabie, D.W., M.T. Merendino and D.H. Reid. 1994. Dis- persal of Bald Eagles fledged in Texas. J. Raptor Res. 28:213-219. McCollough, M.A. 1986. The post-fledging ecology and population dynamics of Bald Eagles in Maine, Ph.D. dissertation, Univ. Maine, Orono, MA U.S.A. Mearns, E.A. 1890. Observations on the avifauna of por- tions of Arizona. Auk 7:45—55. Nelson, M.W. 1969. The status of the Peregrine Falcon in the northwest. Pages 61—72 in].]. Hickey [Ed.], Per- egrine Falcon populations: their biology and decline. Univ. Wisconsin Press, Madison, WI U.S.A. Nisbet, I.C.T. 1989. Organochlorines, reproductive im- pairment, and declines in Bald Eagle populations: mechanisms and dose-response relationships. Pages 483-489 in B.-U. Meyburg and R.D. Chancellor [Eds.], Raptors in the modern world. Third world conference on birds of prey and owls, Eilat, Israel. World Working Group on Birds of Prey and Owls, Ber- lin, Germany. Newton, I. 1979. Population ecology of raptors. Buteo Books, Vermillion, SD U.S.A. Ohmart, R.D. AND R.J. Sell. 1980. The Bald Eagle of the southwest, with special emphasis on the breeding population of Arizona. U.S. Dept. Interior, Water and Power Resources Service, Phoenix, AZ U.S.A. Palmer, R.S. 1988. Handbook of North American birds. Vol. 4. Yale Univ. Press, New Haven, CT U.S.A. Phillips, A, J. Marshall and G. Monson. 1964. The birds of Arizona. Univ. Arizona Press, Tucson, AZ U.S. A. POSTUPALSKY, S. 1974. Raptor reproductive success: some problems with methods, criteria and terminology. Raptor Res. Rep. No. 2:21-31. Rubink, D.M. and K. Podborny. 1976. The southern Bald Eagle in Arizona: a status report. USFWS Endangered Species Rep. 1. Albuquerque, NM U.S.A Sherrod, S.K., C.M. White and F.S.L. Williamson. 1976. Biology of the Bald Eagle on Amchitka Island, Alaska. Living Bird 15:143-182. Sprunt, a., IV, W.B. Robertson, Jr., S. Postupalsky, R.J. Hensel, C.E. Knuder and FJ. Ligas. 1973. Compara- tive productivity of six Bald Eagle populations. Trans. N. Am. Wildl. Nat. Resour. Conf. 38:96-106. Stalmaster, M.V. 1987. The Bald Eagle. Universe Books, New York, NY U.S.A. Swenson, J.E., K.L. Alt and R.L. Eng. 1986. Ecology of Bald Eagles in the Greater Yellowstone Ecosystem. Wildl. Monogr. No. 95. Trent, T.T. and O.J. Rongstad. 1974. Home range and survival of cottontail rabbits in southwestern Wiscon- sin. J. Wildl. Manage. 38:459-472. U.S. Fish and Wildlife Service. 1995. Endangered and threatened species; Bald Eagle reclassification; final rule. 50 CFR Part 17, 12 July 1995. Federal Register 36000-36010. Wiemeyer, S.N., T.G. Lamont, C.M. Bunk, C.R. Sindelar, FJ. Gramlich, J.D. Fraser and M.A. Byrd. 1984. Or- ganochlorine pesticide, polychlorobiphenyl and mer- cury residues in Bald Eagle eggs (1969-79) and their relationships to shell thinning and reproduction. Arch. Environ. Contam. Toxicol. 13:529-549. Willard, F.C. 1916. Notes on the Golden Eagle in Ari- zona. Cowfior 18:200-201. Wood, P.B., D.A. Buehler and M.A. Byrd. 1990. Bald Eagle. Pages 13-21 in Proceedings of the southeast raptor management symposium and workshop. Natl Wildl. Fed., Washington, DC U.S.A. Received 17 May 1998; accepted 14 March 1999 J. Raptor Res. 33(3):227-232 © 1999 The Raptor Research Foundation, Inc. THE INFLUENCE OF WEAPONS-TESTING NOISE ON BALD EAGLE BEHAVIOR Bryan T. Brown SWCA, Inc., Environmental Consultants, 230 South 500 East, Suite 230, Salt Lake City, UT 84102 U.S.A. G. Scott Mills SWCA, Inc., Environmental Consultants, 343 South Scott Avenue, Tucson, AZ 85701 U.S.A. Cornelius Bowels SWCA, Inc., Environmental Consultants, 53 Steeplechase Lane, North East, MD 21901 U.S.A. William A. Russell MCHB-DC-EEN, Environmental Noise, US. Army Center for Health Promotion and Preventive Medicine, Aberdeen Proving Ground, AID 21010 U.S.A. Glenn D. Therres Maryland Department of Natural Resources, 580 Taylor Ave. (E-1), Annapolis, MD 21401 U.S.A. James J. Pottie us. Army Directorate of Safety, Health, and Environment, STEAP-SH-ER/Bldg. E4430, Aberdeen Proving Ground, MD 21010 U.S.A. Abstract. — ^We studied the influence of weapons-testing noise on Bald Eagle {Haliaeetus leucocephalus) behavior at the Aberdeen Proving Ground (APG), Maryland, in 1995. Our objectives were to document and compare eagle behavior at times with and without weapons-testing noise, determine if the frequency of behavior after noise increased with increasing sound levels and compare nest success and productivity on APG with that of adjacent areas of Maryland. Most roosting (72.7%) and nesting (92.7%) eagles showed no activity (i.e., perched motionless) in the 2-sec interval following weapons-testing noise. The most frequent activity following noise was a head turn, exhibited by 18.2% of roosting and 0.7% of nesting eagles; other eagle activities following noise (e.g., body movement, vocalization and flight) were rare at both roosts (9.1%) and nests (6.6%). Frequency of activity after noise differed between adults and juveniles at nests, but did not differ between adults and immatures at roosts. Activity after noise occurred significantly more in roosting than nesting eagles. For roosting eagles, frequency of activity after noise was similar to activity at times without noise. Frequency of no activity versus activity after noise did not vary at sound intensity levels >110 and <110 dBP for either nesting or roosting eagles. Nest success and productivity on APG did not differ from nest success and productivity in adjacent counties of Maryland from 1990-95, suggesting that weapons-testing noise did not influence eagle re- production at the population level. Key Words: Bald Eagle; Haliaeetus leucocephalus; behavior, Chesapeake Bay; human disturbance, Maryland; noise effects. Influencia del ruido de prueba de armas en el comportamiento de aguilas calvas Resumen. — Estudiamos la influencia del ruido de la prueba de armas en el comportamiento de las aguilas calvas {Haliaeetus leucocephalus) en el Campo de Pruebas de Aberdeen (CPA), Maryland en 1995. Nuestros objetivos fueron los de comparar el comportamiento de las aguilas con y sin ruidos de prueba de armas, determinar si la frecuencia de comportamiento despues del ruido aumenta con los niveles de sonido y comparar el exito de anidacion y productiyidad en el CPA y en areas adyacentes de Mary- land. Las aguilas en perchas (72.7%) y en anidacion (92.7%) no mostraron ninguna actividad en el intervalo de 2 segundos despues de la prueba de armas. La actividad mas frecuente despues del ruido 227 228 Brown et al. VoL. 33, No. 3 fue la de girar la cabeza, exhibida por el 18.2% de las aguilas en perchas y 0 . 7 % de las aguilas en anidacion. Otras actividades despues del ruido (movimientos de cuerpo, vocalizaciones y vuelo) fueron raras en las aguilas en perchas (9.1%) y en nidos (6.6%). La frecuencia de actividad despues del ruido difirio entre adultos y juveniles en los nidos, pero no entre adultos y juveniles en perchas. La actividad fue significativamente mayor en las aguilas en perchas que en anidacion. Para las aguilas en perchas la frecuencia de actividad despues del ruido fue similar a la de actividad sin ruido. La frecuencia de respuesta de inactividad versus actividad despues del ruido no vario con la intensidad del sonido 110 y 110 dBP tan to para las aguilas en anidacion o en perchas. El exito de anidacion y productividad en el CPA no difirio del exito de anidacidn y productividad de los condados adyacentes de Maryland entre 1990-95, lo cual sugiere que el ruido de la prueba de armas no afecta la reproduccion de las aguilas a nivel poblacional. [Traduccion de Cesar Marquez] Despite studies direcdy or indirectly addressing the influence of unnatural sound energy, hereafter referred to as noise, from military activities on rap- tors (e.g., Andersen et al. 1986, Manci et al. 1988, Andersen et al. 1989, Stalmaster and Kaiser 1997), there is little consensus on the overall influence of noise on them. Nine of 17 Red-tailed Hawks {Buteo jamaicensis) , not previously exposed to helicopter overflights, flushed from nests exposed to helicop- ter activity (Andersen et al. 1989), although noise and visual parameters of helicopter disturbance were not examined separately. Grubb et al. (1992) reported noise from artillery fire located a median distance of 1.5 km from nesting Bald Eagles {Hal- laeetus leucocephalus) elicited no visible behavioral response in 100% of 25 eagle-noise observations. Stalmaster and Kaiser (1997) reported that 8% of 1452 Bald Eagles flushed during 373 weapons-fir- ing events on the Fort Lewis Army Reservation, Washington. The influence of weapons- testing noise on raptor behavior or reproductive fitness has not been quantitatively examined in other studies, and decibel levels associated with raptor behavior following weapons-testing noise have not previously been documented. Because military in- stallations comprise approximately 9.7 million ha in the U.S. (Pfister 1988), applied information on the effects of weapons-testing noise could be useful in developing raptor management guidelines for military installations. We studied the influence of noise from military weapons-testing at the Aberdeen Proving Ground (APG) , Maryland, on nesting and roosting Bald Ea- gle behavior, nest success, and productivity. Our objectives were to document eagle behavior follow- ing weapons-testing noise, determine if frequen- cies of behavior after noise differed by age, test the null hypothesis that roosting eagle behavior after noise did not differ from behavior at times without noise, test the null hypothesis that the frequency of active behaviors after noise did not increase with increasing sound levels and compare nest success and productivity on APG with that of adjacent ar- eas of Maryland. In addition to using nest success and productivity on APG as an indirect measure of the influence of weapons-testing noise, this com- parison served as a relative indicator of habitat quality. The study area was an ideal locale to ex- amine the influence of noise on eagle behavior be- cause of the abundance of eagle nests and roosts (Buehler et al. 1991a, 1991b) and because of high levels of weapons-testing noise prior to and during our study. Methods APG is a 350 km^ military installation located on the western shore of the northern Chesapeake Bay, 30 km north of Baltimore, Maryland. Access to much of APG is restricted, greatly reducing human-associated activities that may negatively influence eagle behavior and distri- bution (Buehler et al. 1991c, Chandler et al. 1995). The area is dominated by forests of various ages except for the developed Aberdeen and Edgewood cantonment ar- eas and scattered test ranges with open fields. Most of APG is at or near sea level with a largely undeveloped shoreline characterized by marshes and forested wet- lands. Testing of ordnance and weapons has been the primary mission of APG since 1917. Up to several thou- sand impulsive (<1 sec) noise events/day may occur at various test ranges across the installation as a result of explosive detonations and small arms, tank and artillery fire (U.S. Army 1994). Weapons-testing noise rarely oc- curs at night. Observations on eagle behavior after weapons-testing noise were gathered at three nests (11 total individuals sampled) in May and June 1995; observations on roosting eagle behavior after noise events were made throughout 1995 at two large communal roosts (^58 eagles/roost) . Monthly aerial surveys from 1993-95 located up to 94 nesting and roosting eagles on APG. Eagles were not banded or marked, and we were unable to differentiate between most eagles of similar age. Therefore, we were not able to determine the extent of pseudoreplication in roosting eagles, but estimated it to be very low due to September 1999 Bald Eagle Behavior and Noise 229 large daily variance in eagle abundance at roosts (Bueh- ler et al. 1991b) and seasonal turnover in roosting pop- ulations due to migration (Buehler et al. 1991a). Levels of prior exposure to weapons-testing noise was unknown for individual eagles. All nests and roosts were at least 0.5-4 km from test ranges and typically experienced noise events from more than one range. No ranges were visible from nests or roosts due to intervening forests. All noise events resulted from scheduled weapons-testing activities and none were staged for the purpose of this study. Nest observations were made throughout the day and roost observations were made from about 1.5 hr before sunset until dark. Observers used 15-45X spotting scopes from either a fixed blind or stationary vehicle. A 1-60 X video camera was used to record and later review some eagle behavior after noise. Eagle ages were classified as adult (white head and tail), immature (mottled or all-dark plumage), or juvenile (nestling) (Bortolotti 1984). Eagle behavior was recorded in an arbitrary 2-sec in- terval immediately following each weapons-testing noise event. A 2-sec interval was chosen because our prelimi- nary evaluations of eagle behavior following noise events suggested this was an appropriate interval to detect noise- related behaviors, if any, and because intervals >2 sec had an increasing probability of detecting behaviors un- related to noise. Behaviors were categorized as follows: 0 = no discernible activity (i.e., perched motionless) ; 1 = head turn; 2 = body or wing movement; 3 = vocalize; 4 = take to flight; 5 = preen; and 6 = other. We assumed that categories 0-4 described increasing energetic levels of activity; categories 5-6 represented miscellaneous be- haviors. A head turn toward the source of an auditory stimulus is known as an orienting response (Brown 1990). However, head turns recorded in our study in- cluded those toward the noise source, away from the noise source and up. We did not attempt to classify eagle behavior as a re- sponse or no response to noise because of inherent sub- jective assumptions involving cause and effect and be- cause all behaviors we classified were within eagles’ normal behavioral repertoire and could have occurred at any time, regardless of noise. Sample sizes of ^1 often resulted from single noise events if several eagles were under observation. Eagles exposed to occurrences that could have influenced their behavior after noise (e.g., interactions with other eagles) were eliminated from analysis. We collected control data to test the null hypothesis that no difference existed between roosting eagle behav- ior after noise compared to times without noise, using the seven previous categories. Control roosting behavior was collected in 30 consecutive 2-sec intervals/0.5 hr be- ginning about 1.5 hr before sunset and continuing until dark from January-December 1995 on the same days and at the same roosts that experimental roosting behavior was being gathered. Levels of weapons-testing noise were measured in un- weighted peak decibels (dBP) using a Larson Davis Lab- oratories 870 precision integrating sound level analyzer, a Larson Davis 2100 preamplifier and a Larson Davis 2541 microphone. Sound level analyzers were calibrated using either a Bruel and !^aer 4230 sound level calibrator or a Metrosonics CL304 acoustic calibrator. Microphones were located 3.0 m above ground and within approxi- mately 100 m of nests and roosts. Inaccuracies in dBP levels due to the distance between microphones and ea- gles were estimated to be <1 dBP based on the nature of impulsive sound energy and the dBP scale, the rela- tively large distances (S:0.5 km) from firing range to mi- crophone and the standard acoustical formula for deter- mining sound level differences between two receiving locations (Harris 1979). Observers synchronized their watches to the nearest sec with sound level analyzers so behavioral observations could be paired with correspond- ing noise events. Some dBP data were gathered using a Larson Davis portable sound level analyzer 800B (type 1), a Larson Davis 826B pre-amplifier and a Larson Davis 2559 microphone; the sound level analyzer was calibrated before and after use with a Metrosonics CL304 calibrator. We monitored eagle nest success and productivity on APG (experimental) versus adjacent areas (control) from 1990-95 to indirectly evaluate the possibility that weap- ons-testing noise influenced eagle reproduction on APG by affecting nest abandonment and failure. Nest sites were aerially monitored during three visits/season from February through May to determine occupancy and fate We analyzed four measures of eagle reproduction: nest success (% of successful nests/occupied territory) and productivity (number of young assumed fledged/occu- pied territory, young assumed fledged/breeding pair and young assumed fledged/successful nest [Postupalsky 1974, Steenhof 1987]). Adjacent areas of Maryland in- cluded Baltimore, Cecil, Harford and Kent counties. Behavior categories 1—6 were combined and compared to category 0 in all statistical tests because categories 2- 6 were rarely observed. Thus, all behavior comparisons evaluated the difference between frequencies of no dis- cernible activity versus some activity. Chi-square tests for association were used in all behavioral comparisons and in comparison of nest success. Mean nesting productivity on APG versus adjacent counties from 1990-95, was com- pared using independent sample Rests with SPSS soft- ware (Norusis 1993) . Statistical significance was accepted at P < 0.05. Results The most common eagle behavior in the 2-sec interval following weapons-testing noise at nests and roosts was no activity, recorded after 92.7% and 72.7% of noise events, respectively (Table 1). The most frequent activity after noise at nests and roosts was a head turn, recorded after 0.7% and 18.2% of noise events, respectively. Other activity categories (2-6) were recorded for 6.6% and 9.1% of eagle behavior after noise at nests and roosts, respectively. Nesting adults and juveniles showed activity after noise in 1 (1.8%) of 55 and 10 (10.4%) of 96 ob- servations, respectively. Roosting adults and im- matures exhibited activity after noise in 8 (25%) of 32 and 28 (28%) of 100 observations, respec- 230 Brown et al. VoL. 33, No. 3 Table 1. Nesting and roosting Bald Eagle behavior in the 2-sec interval following weapons-testing noise (experi- mental) and at times without noise (control), Aberdeen Proving Ground, Maryland, 1995. Behavior categories in- cluded: 0 = no discernible activity; 1 = head turn; 2 = body or wing movement; 3 = vocalize; 4 = take to flight; 5 = preen; and 6 = other. Eagle Location Noise Number of Eagle Behaviors by Category Total Status 0 1 2 3 4 5 6 Nest Experimental 140 1 0 0 0 10 0 151 Nest Control None gathered Roost Experimental 96 24 2 4 0 6 0 132 Roost Control 5596 1201 76 72 15 1038 229 8227 tively. Frequencies of no activity compared to activ- ity behavior categories after noise differed between adults and juveniles at nests (x^ = 3.82, df = P = 0.05), but did not differ between adults and im- matures at roosts (x^ = 0.11, df = 1, P = 0.74), although our study was not designed to test for differences between age classes. Therefore, all age classes were combined for subsequent analyses. The frequency of active behaviors following noise was higher for roosting eagles than for nesting ea- gles (x^ = 20.32, df = 1, P< 0.001). The most frequent behavior recorded for roost- ing eagles at times without weapons-testing noise was no activity, accounting for 68.7% of control ob- servations (Table 1). Frequencies of no activity ver- sus activity categories did not differ between con- trol and experimental roost observations (x^ = 1.28, df = 1, P = 0.26). Because it appeared un- likely that preening and other behavior (categories 5 and 6, respectively) were reactions to noise events, we compared control and experimental roosting data in two additional ways; without inclu- Table 2. Nesting and roosting Bald Eagle behavior in the 2-sec interval following weapons-testing noise by deci- bel (dBP) level, Aberdeen Proving Ground, Maryland, 1995. Behavior categories have been summarized into no activity (category 0) and activity (categories 1-6). Number of Eagle Behaviors by Category dBP Nesting Eagles Roosting Eagles Levels 0 1-6 Total 0 1-6 Total 80-89 90-99 100-109 — — — 1 1 2 52 7 59 32 21 53 110-119 68 2 70 2 3 5 120-129 20 2 22 6 10 16 sion of these two categories and including these categories as no activity. We could not detect a dif- ference between control and experimental behav- ior at roosts (x^ = 3.12, df = 1, P = 0.08 and x^ — 1.27, df = 1, P = 0.26, respectively) in either com- parison. Despite a lack of difference between experimen- tal and control roost behavior, a small number of activity behaviors following ordnance noise ap- peared to be a direct result of noise. For example, a roosting immature eagle that was preening ap- peared to lose its balance and nearly fell off its perch immediately after an explosion measuring 120.1 dBP on 11 September. However, we did not observe any eagles taking to flight immediately af- ter noise during the study period, although we ob- served this activity once after noise during a 1994 preliminary study. Behavioral observations at nests and roosts fol- lowing weapons-testing noise were paired with dBP levels ranging from 82-126 dBP (Table 2). We re- corded dBP data for all observations at nests but for only 58% of observations at roosts because of wind interference and equipment malfunction. Be- cause of small sample sizes for some dBP levels and because we had no reason to divide dBP categories at any particular level for analysis, we chose to com- pare frequencies of activity after noise between <110 dBP and ^110 dBP. The 110 dBP threshold was chosen to attempt to obtain approximately equal sample sizes for roosts and nests. We did not detect a difference in frequencies of no activity ver- sus activity at sound levels si 10 than at <110 dBP for nesting (x^ = 3.01, df = P = 0.08) and roost- ing eagles (x^ = 2.94, df = P = 0.09). Overall nest success did not differ for 1990-95 between APG (61 occupied territories, 41 success- ful nests, 67% nest success) and adjacent areas of September 1999 Bald Eagle Behavior and Noise 231 Table 3. Summary measures of Bald Eagle reproduction {N = 209 occupied territories) on Aberdeen Proving Ground (APG) compared to adjacent areas (ADJ) along the northern Chesapeake Bay, Maryland, 1990-95. Test statistics are from analyses to determine if differences existed between APG and ADJ for 1990-95. Summary Data Test Parameters Statistics Statistic APG ADJ t df P Young fledged/occupied territory N mean ± SD 61 1.13 ± 0.90 148 1.09 ± 0.96 0.26 207 0.80 Young fledged/breeding pair N mean ± SD 59 1.17 ± 0.89 136 1.19 ± 0.94 -0.15 193 0.88 Young fledged/ successful nest N mean ± SD 41 1.68 ± 0.52 96 1.69 ± 0.64 -0.04 135 0.97 Maryland (148 occupied territories, 96 successful nests, 65% nest success; ~ 0.11, df = 1, P = 0.75) . Overall numbers of young/ occupied terri- tory, young/breeding pair, and young/ successful nest for 1990—95 combined were not significantly different on APG compared to adjacent areas of Maryland (Table 3). Discussion Behaviors that were likely indications of severe noise disturbance, such as body or wing movement and flight, occurred infrequently or were absent during both control and experimental observa- tions. Although some eagles apparentiy reacted, our findings suggest that most eagles have habitu- ated to weapons-testing noise. We did not demon- strate that habituation has occurred, but our find- ings were consistent with a habituation hypothesis. Habituation is an active learning process that per- mits individuals to discard a response to a recur- ring stimulus for which constant response is bio- logically inappropriate without impairment of their ability to respond to other stimuli (Lorenz 1965, Alcock 1979, Peeke and Petrinovich 1984). Because this constitutes tolerance for prolonged and repetitive activities, then the thousands of noise events caused by weapons testing on a typical day at APG would be a likely basis for habituation. Habituation could occur in a relatively short time even for nonresident eagles that migrate into the area. Apparent habituation by many vertebrates to similar noise has been widely documented (e.g., Andersen et al. 1989, Grubb and King 1991, Wei- senberger et al. 1996). Perhaps most unexpected was our finding of apparent eagle habituation to most weapons-testing noise exceeding 120 dBP. For comparison, naturally-occurring thunder ranges from 82-103 dBP at distances of 700-2100 m (Holmes et al. 1971). An alternative hypothesis is that some eagles re- acted to weapons-testing noise by more frequently ceasing activity (i.e., they “froze”). For example, a decrease in flight activity was reported in cave- roosting bats exposed to noise from low-level su- personic aircraft overflights at Organ Pipe Cactus National Monument in Arizona (V.M. and D.C. Dalton unpubl. data) . However, we did not address this hypothesis, which would require evaluating be- havior immediately before and after noise. The null hypothesis that eagle activity after noise did not increase with increasing sound level was not rejected. Sensitization, defined as successively stronger responses to specific stimuli (Peeke and Petrinovich 1984), apparently did not occur in ea- gles exposed to weapons-testing noise at APG. Based on our finding that most eagles exhibited no activity following relatively loud noise events, we concluded that Bald Eagles at nests and roosts at APG do not show a significant behavioral reaction to weapons-testing noise. This conclusion is sup- ported by the finding that sensitization to noise was apparendy not occurring. Our finding that eagle nest success and productivity from 1990-95 was similar for APG and adjacent areas of Maryland suggests that weapons-testing noise did not influ- ence overall reproductive performance of the nest- ing eagle population at APG. Loud noise can induce stress in some animals, resulting in physiological changes such as in- creased heart and respiratory rates, altered blood chemistry and hormone production, hypertension and vasoconstriction (Manci et al. 1988). For ex- ample, Weisenberger et al. (1996) reported that heart rates of ungulates increased relative to in- 232 Brown et al. VoL. 33, No. 3 creasing noise levels produced by simulated jet air- craft overflights but returned to pre-disturbance conditions in 60-180 sec. Our study addressed vis- ible behavior only, and no existing studies of the influence of noise have examined raptor physiol- ogy. We recommend that future research on the effects of noise on Bald Eagles or other raptors should focus primarily on physiology and should attempt to test for a quantitative link between noise, physiology and reproductive fitness. Acknowledgments I. Bon, J. Forch, J. Sandler, and T. Stoflet assisted with field observations. We thank the staff of APG for their support and cooperation, particularly S. Benjamin, J. Broome, K. Broska, H. Hash, J. Kaffl, N. Lewis, and J. Ondek. We thank tower operators in range control for coordinating our access to study sites adjacent to firing ranges. S. Carothers, B. Hoskin, and S. Ruhr of SWCA, Inc., Environmental Consultants provided encourage- ment and administrative support. Early drafts of this manuscript were improved by comments from D.E. An- dersen, M.J. Bechard, G.R. Bortolotti, T.G. Grubb and P.R. Krausman. Funding was provided by the U.S. Army; this paper received OPSEC security clearance number 1488-A-3. Literature Cited Alcock, j. 1979. Animal behavior: an evolutionary ap- proach. Sinauer Associates Publishers, Inc., Sunder- land, MA U.S.A. Andersen, D.E., O.J. Rongstad and W.R. Mytton. 1986. The behavioral response of a Red-tailed Hawk to mil- itary training activity. Raptor Res. 20:65-68. , and . 1989. Response of nesting Red-tailed Hawks to helicopter overflights. Condor 9\\ 296-299. Bortolotti, G.R. 1984. Sexual size dimorphism and age- related size variation in Bald Eagles. J. Wildl. Manage. 48:72-81. Brown, A.L. 1990. Measuring the effect of aircraft noise on seabirds. Environ. Int. 16:587-592. Buehler, D.A., T.J. Mersmann, J.D. Fraser and J.K.D. Seegar. 1991a. Differences in distribution of breed- ing, nonbreeding, and migrant Bald Eagles on the northern Chesapeake Bay. Condor 93:399-408. , , and . 1991b. Nonbreeding Bald Eagle communal and solitary roosting behavior and roost habitat on the northern Chesapeake Bay./. Wildl. Manage. 55:273-281. , , and . 1991c. Effects of hu- man activity on Bald Eagle distribution on the north- ern Chesapeake Bay./. Wildl. Manage. 55:282-290. Chandler, S.K., J.D. Fraser, D.A. Buehler and J.K.D. Seegar. 1995. Perch trees and shoreline development as predictors of Bald Eagle distribution on Chesa- peake Bay. /. Wildl. Manage. 59:325-332. Grubb, T.G. and R.M. King. 1991. Assessing human dis- turbance of breeding Bald Eagles with classification tree models./ Wildl. Manage. 55:500-511. , W.W. Bowerman, J.P. Giesy and G.A. Dawson. 1992. Responses of breeding Bald Eagles, Haliaeetus leucocephalus, to human activities in northcentral Michigan. Can. Field-Nat. 106:443-453. Harris, C.M. 1979. Handbook of noise control. McGraw- Hill Book Co., New York, NY U.S.A. Holmes, C.R., M. Brook, P. Krehbiel and R. McCrory. 1971. On the power spectrum and mechanism of thunder./. Geophys. Res. 76:2106-2115. Lorenz, K. 1965. Evolution and modification of behavior. Univ. Chicago Press, Chicago, IL U.S.A. Manci, K.M., D.N. Gladwin, R. Villella and M.G. Cav- endish. 1988. Effects of aircraft noise and sonic booms on domestic animals and wildlife: a literature synthesis. NERC Rep. 88/29, U.S. Fish and Wildlife Service, Ft. Collins, CO U.S.A. Norusis, M.J. 1993. SPSS for Windows: base system user’s guide, release 6.0. SPSS, Inc., Chicago, IL U.S.A. Peeke, H.V.S and L. Petrinovich [Eds.]. 1984. Habitua- tion, sensitization and behavior. Academic Press, New York, NY U.S.A. Pfister, A.R. 1988. Current and potential management on military installations. Pages 228-231 in R.L. Glin- ski, B.G. Pendleton, M.B. Moss, M.N. LeFranc, Jr., B.A. Millsap and S.W. Hoffman [Eds.], Proceedings of the southwest raptor management symposium and workshop. Sci. and Tech. Series No. 11, Natl. Wildl. Fed., Washington, DC U.S.A. PosTUPALSKY, S. 1974. Raptor reproductive success: some problems with methods, criteria, and terminology. Raptor Res. Rep. 2:21-31. Stalmaster, M.V. and J.L. Kaiser. 1997. Flushing re- sponses of wintering Bald Eagles to military activity. /. Wildl. Manage. 61:1307-1313. Steenhof, K. 1987. Assessing raptor reproductive success and productivity. Pages 157-170 in B.G. Pendleton, B.A. Millsap, K.W. Cline and D.M. Bird [Eds.], Raptor management techniques manual. Sci. and Tech. Se- ries No. 10, Natl. Wildl. Fed., Washington, DC U.S.A. U.S. Army. 1994. Environmental noise consultation No. 52-34-Q3UV-94; noise contours for range activities, Aberdeen Proving Ground, Maryland, June-August 1994. U.S. Army Environmental Hygiene Agency, Ab- erdeen Proving Ground, MD U.S.A. Weisenberger, M.E., P.R. Krausman, M.C. Wallace, D.W. DeYoung and O.E. Maughan. 1996. Effects of simulated jet aircraft noise on heart rate and behavior of desert ungulates./. Wildl. Manage. 60:52-61. Received 20 May 1996; accepted 10 April 1999 J. Raptor Res. 33(3):233-239 © 1999 The Raptor Research Foundation, Inc. EFFECTS OF RECREATIONAL TRAILS ON WINTERING DIURNAL RAPTORS ALONG RIPARIAN CORRIDORS IN A COLORADO GRASSLAND Robert J. Fletcher, Jr. Shawn T. McKinney and Carl E. Bock Department of Environmental, Population, and Organismic Biology, University of Colorado, Boulder, Boulder, CO 80309 U.S.A. Abstract. — Different types of human activity may influence raptors in various ways, potentially affecting their abundance, distribution, habitat use and productivity. We studied the effects of recreational trails on wintering raptor populations in grasslands of eastern Boulder County, Colorado, from December 1995-March 1996. We conducted strip transects to survey raptor populations at six study sites. All sites consisted of short and/ or tallgrass prairie, and all contained a riparian corridor. Three sites contained recreational trails running adjacent to the riparian corridor (trail), while three sites contained no trails (control). Species richness, abundance and perch use were compared between control and trail sites. Species richness was consistently greater in control sites. Abundance of total raptors observed was greater in control sites. Abundance of Bald Eagles {Haliaeetus leucocephalus) was greater in control sites, while abundance of Red-tailed Hawks {Buteo jamaicensis) was similar for control and trail sites. Perching dis- tances from riparian corridors were greater in trail sites than in control sites. In addition, raptors perched along riparian corridors more frequently in control sites. Results of this study suggest that recreational trails may have affected habitat selection of some raptor species in this grassland ecosystem. Key Words: raptors', human activity, recreational trails', perch use, riparian corridors', grasslands. Corredores riberehos en un pastizal de Colorado Resumen. — Diferentes tipos de actividad humana pueden influenciar a las aves rapaces de varias ma- neras; potencialmente afectando su abundancia, distribucion, uso de habitat, y productividad. Estudia- mos los efectos de los senderos recreativos en las poblaciones de aves rapaces del este del Condado de Boulder, Colorado desde Diciembre 1995-Marzo 1996. Efectuamos transectos lineares para documentar las poblaciones de aves rapaces en seis areas de estudio. Los sitios de estudio fueron en pastizales cortos y/o altos que contenian un corredor ribereho (sendero), mientras que tres sitios no contenian ningun sendero (control) . La riqueza de especies, abundancia, utilizacion de perchas fueron comparadas entre el control y los sitios con senderos. La riqueza de especies fue consistentemente mayor en los sitios de control. La abundancia total de aves rapaces observadas fue mayor en los sitios de control. La abun- dancia de aguilas calvas {Haliaeetus leucocephalus) fue mayor en los sitios de control, mientras que la abundancia de Buteo jamaicensis fue similar para ambos sitios. La distancia entre perchas de los corre- dores riberenos fue mayor en los sitios con senderos que en los sitios de control. Adicionalmente, las aves rapaces utilizaron con mas frecuencia las perchas a lo largo de los corredores riberehos en los sitios de control. Los resultados de este estudio sugieren que los senderos recreativos pudieron haber afectado la seleccion de habitat de algunas especies de aves rapaces en este ecosistema de pastizales. [Traduccion de Cesar Marquez] Raptor abundance and distribution can be af- fected by many factors, including prey availability (Howard and Wolfe 1976, Preston and Beane 1996, Gietzen et al. 1997, Plumpton and Andersen 1998), vegetation (Wakeley 1978), and the availability of ^ Present address: Department of Animal Ecology, 124 Science Hall II, Iowa State University, Ames, lA 50011 USA perch sites (Stahlecker 1978, Jones 1989, Janes 1994, Widen 1994). While open space areas and green belts provide potential habitats for grassland raptors in urban environments (Cringan and Hor- ak 1989), such areas are also characterized by a variety of human activities. These human activities may affect raptor habitat use and may influence raptor abundance. Certain effects of human activity on raptor pop- 233 234 Fletcher et al. VoL. 33, No. 3 ulations have been extensively studied, yet with equivocal results. For example, Stalmaster and Newman (1978) found that Bald Eagles {Haliaeetus leucocephalus) avoid areas of high human activity. Smallwood et al. (1996) also observed many spe- cies of raptors avoiding human settlements. How- ever, other research suggests that human activity may not affect raptor distribution and productivity (Mathisen 1968, Grier 1969, Preston and Beane 1996). Human activity may affect raptors in differ- ent ways, and some types of human activity may affect raptors more than others (Preston and Be- ane 1996). For instance, Fraser et al. (1985) found that nesting success of Bald Eagles was not affected by human activity; however, eagles did choose nest sites far from human development. Both Skagen (1980) and Holmes et al. (1993) found that raptors were more sensitive to humans approaching on foot than to humans approaching in vehicles. We studied the potential effects of recreational trails on wintering grassland raptor populations in Boulder County, Colorado, during the winter of 1995-96. We compared species richness, abun- dance, and perch use along riparian corridors in areas with trails adjacent to corridors and areas without trails. We also compared distances between raptor perches and riparian corridors in areas with and without trails along corridors. Study Area and Methods We studied raptors in the grasslands of east Boulder County, Colorado (40°00'N, 105°20'W) from December 1995 until March 1996. We chose six study sites of similar habitat on the City of Boulder Open Space property. All study sites contained riparian corridors, with adjacent vegetation consisting of short and/or tallgrass prairie. Previous research in this region showed a high correla- tion of raptor abundance to perch sites (Jones 1989). We chose sites containing riparian corridors with numerous potential perches to control for perch availability. All sites also contained many perches outside the riparian area, at a wide range of distances from the corridor (0-440 m). Three sites included recreational trails running ad- jacent to the riparian corridor (trail sites), within 15 m of the corridor at all locations. Three sites had no trails (control sites) . Recreational trails were 2-3 m in width and composed of gravel. Throughout the study period, trails were used frequently each day by hikers and/or bicyclists, but were closed to vehicle use. Study site size was determined by open space boundaries, and was es- timated using 7.5 min U.S. Geological Survey topograph- ic quadrangle maps. Overall, trail sites included 72 ha {x = 23.9 ± 6.1, ±SE) of grassland and riparian habitat, and control sites included 54 ha {x = 18.0 ± 5.1). Pre- dominant vegetation along riparian areas included cot- tonwoods {Populus sargenti), willows {Salix spp.) and Rus- sian olives {Elaeagnus angustifolia) . Potential prey items for raptors included lagomorphs (Sylvilagus spp. and Lepus spp.), black-tailed prairie dogs {Cynomys ludovidanus) , and other small rodents (e.g., Mi- crotus ochrogaster) . None of the sites contained prairie dog colonies, but colonies were potentially within the home ranges of raptors seen on each of our study sites. We estimated the distance from the center of each site to the edge of the closest prairie dog colony, based on 1997 Boulder County Open Space digitized maps. The average distance to prairie dog colonies for control sites (2.3 ± 0.6 km) and trail sites (1.9 ± 0.3 km) did not differ sta- tistically (two-tailed, two sample ^-test; t = 0.61, df = 4, P = 0.58). We conducted strip transects to survey raptor popula- tions twice a week at each site from 0700-1130 H (MST) . Each transect was 350 ra long, parallel to the riparian corridor, at distances of 5-30 m from each corridor. Tran- sect width was delineated based on public land bound- aries, ranging from 315-450 m. For each survey, a single observer walked slowly along the transect for 15 min, re- cording the species and number of raptors observed within the study site, the type of perch used by perching raptors, and mapping perches used by raptors. Perch dis- tances (i.e., distance from the perch used to the edge of the riparian corridor) were later measured when raptors were not present. We rotated the order of sites surveyed each day to control for survey time differences. We did not conduct surveys during inclement weather (e.g., snowstorms) because inclement weather could affect the detectability of raptors and there was probably less rec- reational activity during inclement weather. We conduct- ed a total of 77 surveys, ranging from 10-15 for each site. We conducted 41 surveys at trail sites and 36 surveys at control sites. We sampled vegetation along the riparian corridors and in surrounding grassland vegetation at each site. To sample riparian corridors, we walked one 300 m transect running along the grassland/riparian edge at each site, recording the species and diameter breast height (dbh) of the closest tree in each cardinal direction at 50 m intervals {N = 24 for each site). For sampling grassland characteristics at each site, we walked four 300 m tran- sects running parallel to each riparian corridor at dis- tances of 50, 100, 150, and 200 m from the edge of the corridor. Every 50 m we recorded the coarse defining characteristic of the vegetation (shortgrass prairie, tail- grass prairie, broad-leaf shrub, tree, bare ground) and visually estimated the percent ground vegetation cover within a 1 m^ radius aluminum ring placed on the ground {N = 24 for each site). Riparian corridors provided a high, uniform density of potential perches at each of the six sites. We quantified surrounding perch density (i.e., not including the ripar- ian perch density) for each site by counting the total number of natural perches (i.e., trees, shrubs and snags) >2 m in height (Janes 1985) at each study site, and di- viding this total by the area of the site (in ha). We did not include perch types other than trees and shrubs (e.g., utility poles), because we only observed raptors using trees and shrubs for perching in our study sites. To directly compare surveys conducted in study sites of differing size, we analyzed the abundance of raptors as the number of raptors observed per km^ surveyed, and September 1999 Effects of Trails on Raptors 235 Figure 1. The total abundance and species richness (x ± SE) of all raptors observed, December 1995-March 1996. Abundance is defined as the number of raptors observed per km^ surveyed. Species richness is defined as the number of species observed per km^ surveyed. £ £ o +-< Q. CO cu £ 3 6 5 4 3 2 1 0 □Trail □ Control I - -- I " " ■ I “ " " I P' p- ^ species richness as the number of species observed per km^ surveyed. Both species richness and abundance in- dices were analyzed using two-way analysis of variance (ANOVA) with a repeated measures design, using tran- sect type (trail or control) as one factor and sampling date as the repeated measure. Using sampling date as a repeated measure allowed us to investigate any potential temporal differences in abundance and species richness within the sampling period. We used the procedure, PROG MIXED on SAS statistical software, to run the re- peated measure ANOVA analyses, because this procedure can accomodate unbalanced repeated measure designs (SAS 1997). The data collected for perching distances from ripari- an corridors were not normally distributed and therefore were analyzed using a Kruskal-Wallis nonparametric test (with Chi-square approximation, SAS 1989). We investi- gated potential temporal differences in perching dis- tance from corridors by regressing the response variable, perching distance, on sampling date (PROG GLM, SAS 1989). The frequency of use of riparian areas for perch- ing was analyzed using a contingency table, vdth tran- sect type (trail or control) as one criterion, and number of perching events on and off riparian areas as a second criterion. We analyzed dbh, percent ground cover, and perch availability for trail and control sites using one way AN- OVAs. Species composition of the riparian canopy was analyzed using a goodness-of-fit test against the null expectation of equal numbers of trees in trail and control sites. For all analyses, we chose a = 0.05. Arithemetic means ±SE are reported. Results Species Composition. Seven species of raptors were observed during our study, including (in de- Figure 2. Abundance indices {x ± SE) of each species observed at trail and control sites, December 1995- March 1996. Abundance is defined as the number of rap- tors observed per km^ surveyed. creasing frequency of observation): Red-tailed Hawks {Buteo jamaicensis) , Bald Eagles, Rough-leg- ged Hawks {Buteo lagopus), Ferruginous Hawks {Buteo regalis), Northern Harriers {Circus cyaneus), Golden Eagles {Aquila chrysaetos) and Prairie Fal- cons {Falco mexicanus). Seven species were ob- served in control sites, while only three were ob- served in trail sites. We observed a consistent pattern of greater species richness in control sites (ANOVA: F = 12.67, df = 1, P < 0.001) (Fig. 1). Raptor Abundance. Abundance indices for total raptors were greater at control sites (ANOVA: F = 9.70, df = 1, P = 0.003) (Fig. 1). Abundance in- dices for Red-tailed Hawks were similar for control and trail sites (ANOVA: F = 1.52, df = 1, P = 0.224) (Fig. 2). Bald Eagle abundance indices were greater in control sites (ANOVA: F = 6.25, df = 1, P = 0.016). We pooled all other species of raptors, due to small sample size. Abundance indices for these raptors were greater in control sites (ANO- VA: F = 7.29, df = 1, P = 0.010). For all abundance and species richness analyses, sampling date as the repeated measure was not significant (P > 0.15), and there were no statistically significant interac- tions among the two factors (P ^ 0.3), suggesting 236 Fletcher et al. VoL. 33, No. 3 Perch Distance from Riparian Corridor (m) Figure 3. The distribution of raptor perching distances from riparian corridors in trail and control sites, Decem- ber 1995-March 1996. Perch distance was measured from the perch used to the riparian corridor (m) . both the abundance and species richness of rap- tors did not change temporally during the sam- pling period. Perch Use. During the study period, 60.2% (N — 56) of all raptors were observed perching. In all observations, raptors used deciduous trees and snags for perching, even though fence posts, power lines and utility poles were present at the study sites. The distribution of perching distances from riparian corridors was different in control and ex- perimental sites (Fig. 3), where the mean perch distance from riparian corridors was significantly greater in trail sites (155.9 ± 28.0 m) than in con- trol sites (42.3 ± 9.8 m) (Kruskal-Wallis: = 18.97, df = 1, P < 0.001). In addition, raptors perched along riparian corridors significantly more often in control sites (54.8%, = 17) than in trail sites (24.0%, = 6) (x^ = 5.61, df = 1, P = 0.018). Raptors perching on riparian corridors were often flushed by the observer conducting the transect, due to the proximity of the walking tran- sect to the riparian area. However, this occurred in both trail and control sites, and did not appear to change overall habitat use by raptors. Throughout the study period, there was no evidence for a tem- poral change in perching distances from riparian corridors in control (F = 0.12, df = 1, 27, P = 0.74) and trail sites {F— 0.62, df = 1, 24, P = 0.44). Table 1. Habitat characteristics {x ± SE) for trail and control sites, Boulder County, CO, December 1995- March 1996. For riparian canopy composition, each spe- cies is reported as the percent occurrence sampled at control and trail sites (N = 72) . Percent ground cover is the percent sampled within 1 m^ radius points at control and trail sites {N = 72). Site Characteristics Trail Control Riparian canopy: Populus sargenti 45.8 ± 9.6 37.5 ± 16.7 Salix spp. 26.4 ± 8.4 33.3 ± 7.2 Elaeagnus angustifolicP 6.3 ± 3.7 23.6 ± 12.1 Dbh (cm) 26.4 ± 4.3 25.5 ± 1.8 Grass cover (%) 81.4 ± 8.7 83.9 ± 0.7 Bare ground (%) 14.4 ± 7.4 15.7 ± 0.6 Shrub cover (%)^ 4.2 ± 1.5 0.4 ± 0.2 Perches/ha 1.7 ± 0.3 2.7 ± 0.8 ^p< 0.05. Habitat Sampling. Three species of trees, Populus sargenti, Salix spp., and Elaeagnus angustifolia com- prised 84.4% (N = 123) of all trees sampled along riparian corridors at all sites (Table 1). P sargenti did not differ in the percent sampled between con- trol and trail sites (x^ = 0.60, df = 1, P = 0.438), nor did the total number of Salix spp. (x^ = 0.86, df = 1, P = 0.354). However, the percent sampled of E. angustifolia was greater in control sites, due to the riparian canopy of one site being predomi- nately comprised of this species (x^ = 4.02, df = 1, P — 0.045) . The dbh of trees sampled along each riparian corridor did not differ between control and trail sites (ANOVA: F = 0.08, df = 1, P = 0.777). The surrounding vegetation at five of the sites consisted primarily of grazed shortgrass prairie. One trail site consisted primarily of tallgrass prai- rie. The percent of ground cover sampled was pre- dominately either grass, bare ground (no cover), or broad-leaf shrubs. Grass cover did not differ be- tween control and trail sites (ANOVA: F = 0.42, df = 1, P = 0.520), nor did the percent of bare ground (ANOVA: F = 0.13, df = 1, P = 0.723). Shrub cover was greater at trail sites than at control sites (ANOVA: F = 5.51, df = 1, P = 0.020). How- ever, shrub cover comprised <5% of all ground cover sampled (Table 1). The surrounding perch densities were similar between control and trail sites (ANOVA: F = 1.59, df = 1, P = 0.276). September 1999 Effects of Trails on Raptors 237 Discussion Human activity associated with recreational trails in our study area may have affected wintering rap- tor populations, in terms of species richness, abun- dance and perch use. Species richness was consis- tently greater in control areas, with few raptors other than Red-tailed Hawks being observed at trail sites. The abundance of total raptors was more than three times greater in areas without trails, and those raptors observed in areas containing trails perched farther from riparian corridors. Habitat characteristics at trail and control sites were gen- erally similar in both riparian canopy composition and dbh, and in the surrounding grassland habitat. Perch density was also similar in control and trail sites. A predominant difference in trail and control sites was the presence or absence of a recreational trail. Prior research has indicated that humans ap- proaching on foot may affect raptors more than vehicle disturbance (Holmes et al. 1993), but no research to date has correlated the potential ef- fects of recreational trails on raptor distribution and abundance. The potential effects of recreational trails on raptors may be species-specific. Abundance of Bald Eagles was greater at control areas. Bald Eagles avoiding areas of human activity has been docu- mented (Stalmaster and Newman 1978, Fraser et al. 1985), but we are unaware of any previous stud- ies documenting recreational trail effects on Bald Eagles. Abundance of Red-tailed Hawks did not dif- fer between control or trail sites. This was the only species of raptor commonly observed in trail areas. Red-tailed Hawks are generalist raptors, both in diet (Errington 1933, Hansen and Flake 1995) and in distribution (Brown and Amadon 1968). Rnight and Kawashima (1993) observed Red-tailed Hawks exploiting power lines more often than expected, while Minor et al. (1993) found no significant dif- ferences in density and productivity of Red-tailed Hawks in urban and nonurban environments. In Boulder County, both wintering and breeding dis- tributions of Red-tailed Hawks have not been sen- sitive to landscape urbanization (Berry et al. 1998). These studies suggest that Red-tailed Hawks have adapted to human development and associated ac- tivity. Our study suggests that Red-tailed Hawks tol- erate human activity along recreational trails. Al- though our sample size was small for species of raptors other than Red-tailed Hawks and Bald Ea- gles, overall these species were more abundant in areas with no recreational trails. Perch use in control and trail sites varied greatly in our study, where the mean perch distance of raptors from riparian corridors was greater in trail sites. Although the overall density of perches out- side of the riparian areas was similar between con- trol and trail sites, we did not quantify raptor perching distance in relation to the distribution of potential perches from corridors. Nevertheless, raptors did use riparian corridors for perching more in control areas than in trail areas. Other research also indicates that riparian areas are used frequently by raptors, both for wintering habitat (Tingle 1989, Smallwood et al. 1996) and for nest- ing (Hansen and Flake 1995). Our study suggests both that riparian corridors are important areas for wintering raptors and that trails may displace raptor perch use away from riparian habitat. One potential reason for differences in raptor abundance between trail and control sites could be from the increased shrub cover on trail sites. An increase in shrub cover could affect hunting strat- egies of raptors and could conceal prey, making prey less vulnerable (Craighead and Craighead 1956, Wakeley 1978). However, because shrub cov- er comprised <5% of all ground vegetation sam- pled, it is unlikely that this variable would control for the large differences of raptor abundance we observed. The distribution of prey items may affect raptor distribution (Plumpton and Andersen 1998). We did not quantify prey availability at our sites. Prairie dog colonies, which are an important prey base for Bald Eagles, Red-tailed Hawks, and Ferruginous Hawks (Jones 1989), were not present at any of our sites. Distances to the closest prairie dog town were similar for both control and trail sites. Be- cause distances were similar among sites, we be- lieve that prey availability for these larger raptors did not confound our results. If these species were concentrating hunting efforts on prairie dogs, then foraging activities may have been infrequent in our sites. Smaller rodents, such as voles (e.g., Microtus ochrogaster) , are important prey items for Rough- legged Hawks and Northern Harriers. Although vegetation was generally similar for control and trail sites, we cannot ascertain if vole populations were different among sites. Because of large home ranges and movement ability of wintering raptors, effects of human activ- ity and urbanization on wintering raptors should 238 Fletcher et al. VoL. 33, No. 3 be viewed with a landscape context (Berry et al. 1998). In areas of human activity, other factors af- fecting wintering raptors in the landscape could be mitigated, such as enhancing foraging opportuni- ties (Preston and Beane 1996) or increasing perch availability (Widen 1994). Buffer zones have been suggested as a means to reduce conflicts between human activities and rap- tor populations (Stalmaster and Newman 1979, Knight and Skagen 1988, Holmes et al. 1993). Al- though prior research has documented potential buffer zones for wintering raptors (Holmes et al. 1993), future research should focus on determin- ing appropriate buffer zones from recreational trails and other specific types of human activity during the nonbreeding season. The type of activ- ity should be considered because different effects could potentially occur from different types of ac- tivity (Preston and Beane 1996). By integrating perch availability (e.g., Marion and Ryder 1975), habitat type, prey abundance, and types of human activity, suitable indices for predicting optimal rap- tor environments and also appropriate buffer zones may be determined. Acknowledgments We thank D. Andersen, J. Chace, D. Craig, A. Cruz, J. Moreno, and one anonymous reviewer for helpful sug- gestions on earlier drafts of this manuscript. We are grateful to C. Miller and the City of Boulder Open Space Department for permitting this research on their prop- erty. Literature Cited Berry, M.E., C.E. Bock and S.L. Haire. 1998. Abundance of diurnal raptors on open space grasslands in an ur- banized landscape. Co^^^^or 100:601-608. Brown, L. and D. Amadon. 1968. Eagles, hawks, and fal- cons of the world. McGraw-Hill Book Co., New York, NY U.S.A. Craighead, JJ- AND EC. Craighead. 1956. Hawks, owls and wildlife. Stackpole Co., Harrisburg, PA U.S.A. Cringan, A.T. AND G.C. Horak. 1989. Effects of urbani- zation on raptors in the western United States. Pages 219-228 in B.G. Pendleton [Ed.], Proceedings of the western raptor management symposium and work- shop. Natl. Wildl. Fed., Washington, DC U.S.A. Errington, P.L. 1933. Food habits of southern Wisconsin raptors part II: hawks. Condor 35:19-29. Fraser, J.D., L.D. Frenzel and J.E. Mathisen. 1985. The impact of human activities on breeding Bald Eagles in north central Minnesota. J. Wildl. Manage. 49:585- 592. Gietzen, R.A., S.R. Jones and RJ. McKee. 1997. Hawks, eagles, and prairie dogs: population trends of winter- ing raptors in Boulder County, 1983-1996./. Colorado Field Omithol. 31:75—86. Grier, J.W. 1969. Bald Eagle behavior and productivity responses to climbing nests. J. Wildl. Manage. 33:961- 966. Hansen, R.W. and L.D. Flake. 1995. Ecological relation- ships between nesting Swainson’s and Red-tailed Hawks in southeastern Idaho. J. Raptor Res. 29:166- 171. Howard, R.P. and M.L. Wolfe. 1976. Range improve- ment practices and Ferruginous Hawks./. Wildl. Man- age. 29:33-37. Holmes, T.L., R.L. Knight, L. Stegall and G.R. Craig. 1993. Responses of wintering grassland raptors to hu- man disturbance. Wildl. Soc. Bull. 21:461-468. Janes, S.W. 1985. Habitat selection in raptorial birds. Pag- es 159-188 in M.L. Cody [Ed.], Habitat selection in birds. Academic Press, Orlando, FL U.S.A. . 1994. Partial loss of Red-tailed Hawk territories to Swainson’s hawks: relations to habitat. Condor 96: 52-57. Jones, S.R. 1989. Populations and prey selection of win- tering raptors in Boulder County, Colorado. Proc. N. Am. Prairie Con. 11:255-258, Knight, R.L. and S.K. Skagen. 1988. Effects of recrea- tional disturbance on birds of prey: a review. Pages 355-359 in R.L. Glinski, B.G. Pendleton, M.B. Moss, M.N. LeFranc, Jr., B.A. Millsap and S.W. Hoffman [Eds.], Proceedings of the southwest raptor manage- ment symposium and workshop. Natl. Wildl. Fed., Washington, DC U.S.A. ANdJ.Y Kawashima. 1993. Responses of raven and Red-tailed Hawk populations to linear right-of-ways./ Wildl. Manage. 57:266-271. Lingle, G.R. 1989. Winter raptor use of the Platte and North Platte River valleys in south central Nebraska. Prairie Nat. 21:1-16. Marion, W.R. and R.A. Ryder. 1975. Perch-site prefer- ences of four diurnal raptors in northeastern Colo- rado. Cowdor 77:350— 352. Mathisen, J.E. 1968. Effects of human disturbance on nesting of Bald Eagles./. Wildl. Manage. 32:1—6. Minor, W.F., M. Minor and M.F. Ingraldi. 1993. Nesting of Red-tailed Hawks and Great Horned Owls in a cen- tral New York urban/ suburban area. /. Field Ornithol. 64:433-439. Plumpton, D.L. and D.E. Andersen. 1998. Anthropogen- ic effects on winter behavior of Ferruginous Hawks./. Wildl. Manage. 62:340-346. Preston, C.R. and R.D. Beane. 1996. Occurrence and distribution of diurnal raptors in relation to human activity and other factors at Rocky Mountain Arsenal, Colorado. Pages 365-374 in D.M. Bird, D.E. Varland and J.J. Negro [Eds.], Raptors in human landscapes: adaptations to built and cultivated environments. Ac- ademic Press, San Diego, CA U.S.A, SAS Institute, Inc. 1989. SAS/ STAT user’s guide. Ver- September 1999 Eefects oe Trails on Raptors 239 sion 6, fourth edition. Vol. 1. SAS Inst. Inc., Cary, NC U.S.A. . 1997. SAS/STAT software: changes and enhance- ments through release 6.12. SAS Inst. Inc., Cary, NC U.S.A. Skagen, S.K. 1980. Behavioral responses of wintering Bald Eagles to human activity on the Skagit River, Washington. Pages 231-241 mR.L. Knight, G.T. Allen, M.V. Stalmaster and C.W. Servheen [Eds.], Proceed- ings of the Washington Bald Eagle symposium. The Nature Conservancy, Seattle, WA U.S.A. Smallwood, S.K., B.J. Nakamoto and S, Geng. 1996. As- sociation analysis of raptors on a farming landscape. Pages 177-190 in D.M. Bird, D.E. Varland and J.J. Ne- gro [Eds.], Raptors in human landscapes: adaptations to built and cultivated environments. Academic Press, San Diego, CA U.S.A. Stahlecker, D.W. 1978. Effect of a new transmission line on wintering prairie raptors. Condor 80:444—446. Stalmaster, M.V. and J.R. Newman. 1978. Behavioral re- sponses of wintering Bald Eagles to human activity, J. Wmi Manage. 42:506-513. and . 1979. Perch-site preferences of win- tering Bald Eagles in northwest Washington. J. Wildl Manage. 43:506—513. Wakeley, J.S. 1978. Factors affecting the use of hunting sites by Ferruginous Hawks. Cowofor 80:316-326. Widen, P. 1994. Habitat quality for raptors: a field exper- iment, y. Avian Biol. 25:219—223. Received 29 August 1998; accepted 2 April 1999 Short Communications /. Raptor Res. 33(3) : 240-244 © 1999 The Raptor Research Foundation, Inc. Northern Goshawks Nesting on a Private Industrial Forest in Western Washington Thomas Bosakowski, Bruce McCullough, Frank J. Lapsanskyand Martin E. Vaughn^ Beak Consultants Incorporated, 12931 NE 126th Place, Kirkland, WA 98034 U.S.A. Key Words: Northern Goshawk, Accipiter gentilis; nest, habitat, productivity, nesting density, forestry. Pacific Northwest. Very little is known about Northern Goshawk {Accipiter gentilis) nesting on private, intensively-managed timber- lands in the Pacific Northwest. Previous information about goshawk nesting on the west side of the Cascades (Fleming 1987, U.S. Forest Service 1994) is known pri- marily from pristine National Parks and Wilderness Are- as, as well as National Forest lands which have generally received light to moderate levels of timber harvesting since European settlement. In 1995, the Murray Pacific Corporation implemented a multi-species Habitat Conservation Plan (HCP) which requires periodic monitoring of goshawk populations on their Mineral Tree Farm in Lewis County, Washington U S.A. (Beak Consultants Incorporated 1995). To meet the requirements of the HCP, we designed and conduct- ed a monitoring plan to determine the occupancy and breeding status of goshawks in all suitable and marginal nesting habitats on the tree farm. Additional information was obtained on goshawk breeding density, productivity, prey, nest sites, and habitat characteristics which are also presented. Study Area and Methods The study area was the Mineral Tree Farm in Lewis County, WA, an industrial tree farm of about 21 600 ha m size which is owned and operated by the Murray Pa- cific Corporation. The tree farm is located along the western edge of the Southern Cascade Physiographic Province (Franklin and Dyrness 1984) and ranges in el- evation from approximately 300-1640 m. Approximately 19 600 ha of the tree farm are capable of supporting for- est, with the remaining area containing rocky ridgetops, alpine meadows, cliff/ talus slopes, roads, brush, standing water and gravel pits. The study area falls within the West- ern Hemlock {Tsuga heterophylla) Forest Zone (Franklin and Dyrness 1984), which is dominated by Douglas-fir {Pseudotsuga menziesii) and western hemlock. Western red cedar ( Thuja plicata) is locally abundant while Pacific sil- ^ Present address: Biota Pacific Environmental Sciences, Inc., 10516 East Riverside Drive, Bothell, WA 98011 US.A, ver fir {Abies amabilis) and noble fir (A. procera) are pre- sent at higher elevations. Commercial timber production is the dominant land use on the Mineral Tree Farm. In 1994, the landscape was covered by stands of various age classes including: recent clearcuts (0-5 yr old; 12%), sap- lings (6-26 yr old; 43%), pole forest (27-45 yr old; 28%) and commercially mature forest (>45 yr old; 15%). Scat- tered patches of remaining old-growth forest (250+ yr old) were lumped into the mature forest category be- cause they were rare (<2%) and most (55%) had expe- rienced partial (<50%) overstory removal within the past 80 yr. The area is characterized by a mild, wet maritime climate. Precipitation occurs mainly in winter and aver- ages 1.8-3. 6 m annually (Cummans et al. 1975). Murray Pacific’s Geographic Information System (GIS) timberlands database was queried to identify areas of po- tentially suitable goshawk nesting habitat on the Mineral Tree Farm. We divided habitat in two categories: (1) suit- able habitat was considered any conifer-dominated stand ^45 yr of age and (2) marginal habitat wsis considered any young coniferous stand 27-44 yr old or any stand of al- der/hardwoods. Alder stands were dominated by red al- der {Alnus rubra) and often contained a minor cohort of big-leaf maple {Acer macrophyllum) and black cottonwood {Populus trichocarpa) , especially in riparian areas. Surveys for nesting goshawks were conducted over the entire tree farm in 1995 and 1996 on all suitable and marginal habitat stands. In addition, we conducted a pi- lot survey for goshawks on the tree farm in 1994 (Bosa- kowski and Vaughn 1996) that was limited only to mature conifer habitat (28 cm dbh [diameter at breast height] , 24 m height, 250-750 trees/ha, ^70% conifer composi- tion) . The protocol for surveys is described in Bosakowski and Vaughn (1996) and includes several modifications to the current U.S. Forest Service protocol for surveying goshawks in the Pacific Northwest (Appendix B in U.S. Forest Service 1994). Nest site habitat attributes were measured from 15-m radius circular plots centered on nest trees. All trees and snags (dead trees) >10 cm in diameter within the plot were measured for dbh with a diameter tape and were identified to species. Heights of nests and nest trees were measured with a clinometer. Maximum slope through the nest site was measured with a clinometer and slope aspects were recorded. The presence or absence of can- opy cover at 20 points was determined with an ocular sighting tube (James and Shugart 1970). The 20 points were systematically positioned with five points along each of the four cardinal directions at 10-m intervals. At each 240 September 1999 Short Communications 241 Table 1. Characteristics of nest trees and nest sites for three Northern Goshawk territories on the Mineral Tree Farm in the Western Washington Cascades. Parameter East Fork Snow Creek Tilton Mineral Creek Mean (SD) Species Nest height (m) Douglas-fir 13.7 Nest Tree Western hemlock 14.7 pacific silver fir 13.4 13.9 (0.68) Nest tree height (m) 39.2 30.6 21.0 30.3 (9.10) Nest tree dbh (cm) 53 73 41 55.7 (16.17) dbh (cm) 29 Nest Site“ 32 24 25.5 (2.94) Canopy cover (%) 80 90 90 86.7 (5.77) Shrub cover (%) 20 20 45 28.3 (14.43) Live tree density (slO cm) (no. /ha) 496 665 1132 764 (329.4) Snag density (#/ha) 155 298 298 250 (82.56) Sapling density (5-10 cm) (no. /ha) 0 62 170 77.3 (86.0) Pacific silver fir density (no. /ha) 0 298 482 260 (243.2) Red alder density (no. /ha) 58 0 410 156 (221.9) Western hemlock density (no. /ha) 368 352 240 320 (69.74) Douglas-fir density (no. /ha) 70 15 0 28.3 (36.85) Elevation (m) 600 745 975 773 (189.1) Slope (%) 46 42 20 36 (14.0) Slope aspect N N NE — Stand age (yr) 40 54 43 45.7 (7.37) qdbh (cm) 29 34 22 28.3 (4.04) Patch size (ha)*’ 421 559 210 396.7 (175) “Measurements within a 15-m radius circle. Nest stands were silvicultural units at least 33 ha as delineated from previous even-aged timber harvests, but all formed a contiguous forest patch mosaic with other closed-canopy stands of various ages ^27 yr. point, the presence or absence of shrub/sapling cover within arms’ reach was also noted (Collins et al. 1982). Only woody stems >0.5 m in height and <10 cm dbh were considered shrub/ sapling cover. Stand-level data were obtained from recent stand in- ventory records for stand age and stand quadratic mean diameter at breast height (qdbh). Qdbh was calculated using all trees ^10 cm dbh with the formula: qdbh = square root ((basal area/trees per ha) /0. 005454) . A landscape-level spatial analysis with GIS was used to calculate the proportion of closed-canopy forest cover (S27 yr) and mature forest cover (s45 yr) within con- centric circles at increasing distances (radii) from the nest. The range of distances tested was 0.2-3. 2 km from the nests and was inclusive of the 2400-ha home range, 240-ha territory or postfledging family area (PFA) and 12-ha nest site as defined by Reynolds et al. (1992) for the Northern Goshawk in the southwestern U.S. Prey remains were found under nests or at nearby prey- plucking posts. These were identified and enumerated using standard methods described by Reynolds and Mes- low (1984). Results During the 1994-96 study period, we found goshawks nesting in three different territories on the Mineral Tree Farm. Only one nest was occupied in 1995 (Mineral Creek), producing at least one fledgling. In 1996, three different occupied nests were monitored, each produc- ing two fledglings for a productivity rate of 2.0 young fledged per nesting attempt. Overall, the nesting results translated into a nesting density ranging from 0.0046— 0.0139 nests per 100 ha. Nest stands were 40-54 yr old and were dominated by conifers (Table 1). Nest stands were densely stocked (496—1132 live trees/ha) on north-facing slopes with high canopy cover (80-90%), low to moderate shrub/ sapling cover (20-45%) and high snag densities (155— 298/ha). We attributed high snag densities to high sup- pression mortality, since all stands were still within the stem exclusion stage (Oliver and Larson 1990) for un- managed (never thinned) second-growth stands in west- ern Washington. At the densest nest site (Mineral Creek) , 242 Short Communications VoL. 33, No. 3 0 1000 2000 3000 4000 RADIUS (m) Figure 1. Percent forest cover at increasing distances (circle radii) from three Northern Goshawk nests on the Mineral Tree Farm in western Washington in 1996. Note that the data in Fig. la (top graph) is for all forest cover S27 yr and is inclusive of the data shown in Fig. lb (bot- tom graph) for all forest cover ^45 yr. The postfledging family area (PFA) is as described in Reynolds et al. (1992). the nest tree was located on an old, overgrown logging road where the forest was more open and had a higher component of alder. In all cases, nests were built in dom- inant trees (mean dbh = 55.7 cm) of the stands which averaged only 28.3 cm in qdbh. Nest heights were con- sistently similar (range = 13.4—14.7 m), despite different nest tree heights and stand ages/types. All nests were in large contiguous forest patches (range = 210-559 ha) comprised of a mosaic of closed-canopy forest (suitable and marginal habitat) of various ages and tree compositions. Landscape-level analyses of the three goshawk nests showed a similar trend for the amount of closed-canopy forest (Fig. la), with mean values of 100% forest cover within 196 m of the nest (“nest site”), 60% cover within 880 m (“PFA”) and 30% cover within 2780 m (“home range”). However, the same trend was not apparent for older forest (Fig. lb), where two of three territories had <10% mature forest cover >45 yr old. All nests were >1.6 km from other landowners and within this distance, there was no old-growth forest at Mineral and Snow Creek nests and only 45 ha at the East Fork nest. A total of 37 prey items were recorded with grouse representing the most numerous prey item (56.7%), Al- though we did not differentiate between Blue Grouse {Dendragapus obscurus) and Ruffed Grouse {Bonasa umbel- lus) in most cases. Blue Grouse were detected at 25% of point counts whereas Ruffed Grouse were detected at only 5% of point counts on the tree farm (Bosakowski 1997). Steller’s Jays {Cyanodtta stelleri) were the second most numerous prey items (16.2%), followed by snow- shoe hare {Lepus americanus, 8.1%) and Northern Flicker {Colaptes auratus, 5.4%). Prey items found only once were Pileated Woodpecker (Dryocopus pileatus) , Varied Thrush {Ixoreus naevius), Swainson’s Thrush {Catharus ustulatus) and Gray Jay {Perisoreus canadensis). Discussion We found goshawks nesting successfully in younger and denser stands than previously reported for north- western goshawks (Reynolds 1983, Fleming 1987, Reyn- olds 1989, Marshall 1992). The spatial analysis revealed that closed-canopy conifer forest cover (>27 yr old) was a consistent feature at the landscape level for all three nests, with 100% forest cover within the 12-ha nest site, about 60% cover within the 240-ha PFA, and about 30% cover within the 2400-ha home range. Plot samples showed that nest sites were composed of young (40-54 yr old) second-growth conifer-dominated forest with high tree and snag densities. None of these stands had re- ceived either commercial or pre-commercial thinning and average qdbh was low (28.3 cm) for goshawks even though saplings (trees <10 cm dbh) were not included in the calculations. Previous measurements of goshawk nest sites in the western Washington Cascades have shown an average stand dbh of 48.3 cm and minimum of 34.5 cm (Fleming 1987). Nesting stands in our study were at or younger than the usual harvest age (45-60 yr) for industrial forest land in western Washington. Growing conditions in the mesic Western Cascades fosters rapid tree growth, allowing suit- able nesting habitat to develop in as little as 40 yr. This age is about half the minimum age (75 yr) needed for goshawk nesting habitat as predicted by a model devel- oped in drier interior Douglas-fir forests of Idaho (Lil- ieholm et al. 1993). Even though the nesting stands in this study were relatively young, it is noteworthy that all three of the nest trees were dominants or remnants (East Fork Tilton nest), having the largest (or second largest) diameter in the sample plot. Similarly, Fleming (1987) also noted that in younger second-growth stands, gos- hawks usually selected the largest trees to provide suitable support for the large, heavy nest since larger trees are more likely to have thicker, stronger support branches. These studies indicate the value of remnant trees (resid- uals) for goshawks and confirm a definite benefit to September 1999 Short Communications 243 maintaining green leave-tree clumps and individual trees after timber harvesting. We noted that grouse were more prevalent in the diet on the tree farm in comparison to Northwest National Forests, where goshawks relied primarily on forest-inte- rior species (Reynolds and Meslow 1984, Bull and Hoh- mann 1994). Since one of the principal foods of the Blue Grouse is Douglas-fir needles (Martin et al. 1951), we speculate that replanting clearcut areas with (primarily) Douglas-fir may have helped support successful goshawk nesting under less than optimal habitat conditions. Our finding of three occupied nests in 1996 on the Mineral Tree Farm represents a nesting density of about one-half to one-fourth of the nesting density that we cal- culated for three to seven occupied nests located by DeStefano et al. (1994) on the Wallowa-Whitman Nation- al Forest in the Oregon Cascades. At least in one year, goshawk reproduction on the Mineral Tree Farm was similar or higher than reproduction occurring in more pristine National Forests of Oregon (DeStefano et al. 1994, Bull and Hohmann 1994). Overall, this study sug- gests that goshawks may be breeding more commonly on private industrial forests than previously predicted, which might be contributing substantially to the regional pop- ulation of goshawks. In the Cascades of Washington and Oregon, the frequent intermingling of National Forest land and private industrial timberland further under- scores the importance of managing for goshawks under both types of ownerships. Resumen. — Esta investigacion revelo que Accipiter gmtilis anido en un bosque comercial privado (21 600 ha) en el Condado de Lewis, Washington, Estados Unidos. El nu- mero maximo de nidos activos fue de 1-3 lo que se tra- duce a una densidad de 0.0046-0.0139 nidos por 100 ha. Cada uno de estos tres nidos produjo dos juveniles en 1996, para una productividad de 2.0 juveniles por intento de anidacion. Los arboles del nido eran coniferas de 40- 54 anos. Todos los nidos estaban ubicados en parches continuos de bosque (entre 210-559 ha) que incluian un mosaico de dosel cerrado (>27 anos) de distintas edades y composiciones. Los niveles de analisis de paisaje de los tres nidos revelaron una tendencia similar en relacion a la cantidad de bosque de dosel cerrado (>27 anos) , con valores medios del 100% de cobertura boscosa dentro de 196 mts del nido (sitio del nido), 60% de cobertura den- tro de 880 mts (area de la familia posterior a la eclosion) y 30% de cobertura dentro de los 2.780 mts (rango de hogar). Sinembargo, este mismo patron no es aparente en bosques mas antiguos donde dos de los tres territorios tenian <10% de cobertura madura (>45 anos) dentro de su rango de hogar. [Traduccion de Cesar M^quez] Acknowledgments Funding for this research project was provided by the Murray Pacific Corporation, Tacoma, Washington. We also thank Gene McCaul from Murray Pacific for the GIS calculations and mapping, K.J. Jorgensen, D. Doyle, R.M. Scharpf, D.L. Olson, and N. Miller for assistance with fieldwork, and Brian Woodbridge, John Marzluff, Dave Hays and Anna Bruce for reviewing the manuscript. Literature Cited Beak Consultants Incorporated. 1995. Amendment to the habitat conservation plan and incidental take per- mit PRT-777837 for the Northern Spotted Owl on tim- berlands owned by the Murray Pacific Corporation, Lewis County, Washington. Murray Pacific Corp., Ta- coma, WA U.S.A. Bosakowski, T. 1997. Breeding bird abundance and hab- itat relationships on a private industrial forest in the Washington Cascades. Northwest Sci. 7L87- 96. AND M.E. Vaughn. 1996. Developing a practical method for surveying Northern Goshawks in man- aged forests of the western Washington Cascades. West.]. Appl. Far. 11:109-113. Bull, E.L. and J.H. Hohmann. 1994. Breeding biology of Northern Goshawks in northeastern Oregon. Stud Avian Biol. 16:103—105. Collins, S.L., EC. James and P.G. Risser. 1982. Habitat relationships of Wood Warblers (Parulidae) in north- ern central Minnesota. Oikos 39:50-58. CuMMANS, J.E., M.R. COLLINGS AND E.G. Nassar. 1975 Magnitude and frequency of floods in Washington U.S. Geological Survey Open File Report 74-336, Ta- coma, WA U.S.A. DeStefano, S., S.K. Daw, S.M. Desimone and E.C. Mes- low. 1994. Density and productivity of Northern Gos- hawks: implications for monitoring and management. Stud. Avian Biol. 16:88-91. Fleming, T.L. 1987. Northern Goshawk status and habitat associations in western Washington with special em- phasis on the Olympic Peninsula. USDA Forest Ser- vice, Pacific Northwest Forest and Range Experiment Station, Olympia, WA U.S.A. Franklin, J.F. and C.T. Dyrness. 1984. Natural vegetation of Oregon and Washington. Oregon State Univ. Press, Corvallis, OR U.S.A. James, F.C. and H.H. Shugart, Jr. 1970. A quantitative method of habitat description. Audubon Field Notes 24: 727-736. Lilieholm, R.J., W.B. Kessler and K. Merrill. 1993. Stand density index applied to timber and goshawk habitat objectives in Douglas-fir. Environ. Manage. 17: 773-779. Marshall, D.B. 1992. Status of the Northern Goshawk in Oregon and Washington. Portland Audubon Soci- ety, Portland, OR U.S.A. Martin, A.C., H.S. Zim and A.L. Nelson. 1951. American wildlife and plants: a guide to wildlife food habits. Do- ver Publ., Inc., New York, NY U.S.A. Oliver, C.D. and B.C. Larson. 1990. Forest stand dynam- ics. McGraw-Hill Inc., New York, NYU.S.A. 244 Short Communications VoL. 33, No. 3 Rewolds, R.T. 1983. Management of western coniferous forest habitat for nesting Accipiter hawks. USDA For. Serv. Gen. Tech. Rep. RM-102, Rocky Mountain For- est and Range Experiment Station, Fort Collins, CO U.S.A. . 1989. Accipiters. Pages 92-101 mB.G. Pendleton, C.E. Ruibal, D.L. Krahe, K. Steenhof, M.N. Kochert, and M.N. LeFranc, Jr. [Eds.], Proceedings of the west- ern raptor management symposium and workshop. Natl. Wildl. Fed., Scientific and Tech. Series No. 12, Washington, DC U.S.A. AND E.C. Meslow. 1984. Partitioning of food and niche characteristics of coexisting Accipiter during breeding. Auk 101:761-779. , R.T. Graham, M.H. Reiser, R.L. Bassett, P.L. Kennedy, D.A. Boyce, Jr., G. Goodwin, R. Smith and E.L. Fisher. 1992. Management recommendations for the Northern Goshawk in the southwestern United States. USDA For. Serv. Gen. Tech. Rep. RM-217, Fort Collins, CO U.S.A. U.S. Forest Service. 1994. Goshawk monitoring, man- agement, and research in the Pacific Northwest: 1994 status report. T. Schommer and G. Silovsky [Compil- ers], USDA For. Serv., Portland, OR U.S.A. Received 29 August 1998; accepted 3 April 1999 J. Raptor Res. 33(3):244-247 © 1999 The Raptor Research Foundation, Inc. Notes on Eggs of the Bicolored Hawk Accipiter bicolor Russell Thorstrom and Lloyd F. Kjff The Peregrine Fund, 566 West Flying Hawk Lane, Boise, ID 83709 U.S.A. Key Words: Bicolored Hawk, Accipiter bicolor; eggsr, nest, Neotropical raptor. Little is known of the breeding habits of Neotropical accipiters, including the widely distributed, but incon- spicuous, Bicolored Hawk {Accipiter bicolor). From 1991- 94, we conducted a study of the breeding biology of this species in Tikal National Park in NE Guatemala. Here, we provide details on Bicolored Hawk eggs found at Tikal and compare them to earlier published descriptions, in- cluding some that were probably misidentified. Fourteen Bicolored Hawk eggs from six nests at Tikal were dull white, unspotted, nonglossy and subelliptical in shape (Preston in Palmer 1962). The inner surface of the shells had a light bluish or greenish tinge, as is typical of many Amjfjiter species (Bent 1937, Schonwetter 1961). As measured in the field with Vernier calipers, they averaged 47.1 ± 1.1 mm (±SD, range = 44.9-49.0) X 36.5 ± 1.1 mm (range = 35.0-38.6), and the average mass of eggs at various stages of incubation was 33.5 ± 3.5 g (range = 28.0-38.0), as taken with a 100 g Pesola spring scale. Overall, the eggs closely resembled those of the related Cooper’s Hawk {Accipiter cooperii) , aside from their slightly greater “elongation” (length /breadth ratio) (i.e., 1.29 [this sample] vs. 1.26 [N = 172 for the Cooper’s Hawk, Kiff unpubl. data] ) . Only clutches of 1-3 eggs were ob- served at Tikal with an average clutch size of 2.4 eggs. These details differ from some published descriptions of Bicolored Hawk eggs and museum specimens attri- buted to this species, but they agree with others (Table 1) . In order of their collecting year, other purported eggs of this species include the following: Chubb (1910) re- ported that three eggs, probably representing a single clutch, of the race A. bicolor pileatus, taken in Paraguay (locality unknown) on 19 October 1902 were “dull white without any markings beyond a few underlying cloudings of a somewhat darker tint.” They measured 1.8-1. 9 in (46-48 mm) X 1.4— 1.5 in (36—38 mm), thus agreeing closely with the Bicolored Hawk eggs from Tikal. Based on these sparse details, there is no reason to doubt the authenticity of these eggs. A set of two purported Bicolored Hawk {Accipiter b. tri- color) eggs taken near Bartica, British Guiana on 13 April 1927 for the British collector G.D. Smooker is in the Western Foundation of Vertebrate Zoology collection (WFVZ 16 695). According to the accompanying data, the species identity was based on a description of the birds The eggs were in an advanced stage of incubation, and the nest was said to be “composed of sticks and placed high up in a tree — no further particulars given.” The egg measurements were noted on the data card as 42.0 X 32.8 and 42.6 X 33.1 mm. The collector of the set was noted as “a native,” probably indicating that Smooker did not visit the nest himself. Other misidentified raptor eggs have been reported from the Belcher-Smooker Trin- idad collection (e.g., Boyce and Kiff 1981, ffrench 1973, Kiff 1981); thus, all records from Smooker, who evidently relied heavily on native collectors, should be viewed with caution. September 1999 Short Communications 245 Table 1 . Comparison of Bicolored Hawk {Accipiter bicolor) eggs from Guatemala with earlier published accounts and unreported museum specimens. See text for additional details on egg appearance and nests. Authority (Year) (A) Length (mm) Breadth (mm) Country (Year Collected) Probable Authenticity This study {N =14) 47.1 (44.9-49.0) 36.5 (35.0-38.6) Guatemala (1991-93) Authentic Chubb (1910) {N= 3) 46-48 36-38 Paraguay (1902) Authentic G.D. Smooker (WFVZ 16 695) 42.02, 42.75 33.06, 33.31 Brit. Guiana (1927) Misidentified Hewitt (1937) 38 32.75 Venezuela (1936) Misidentified Housse (1945) 44 34 Chile (year?) Indeterminate Schonwetter (1961) 37.0, 41.3 31.0, 32.6 Not given Misidentified Schonwetter (1961) 46.5 35.2 Ecuador (1961) Authentic? Wolfe (1964) (WFVZ 16 694) 46.27 36.15 Mexico (1961) Authentic F.F. Nyc, Jr. (WFVZ 145 391) 46.9, 48.4 36.4, 36.1 Mexico (1970) Authentic This set was mentioned by Wolfe (1964), who de- scribed their appearance as a “grayish white ground col- or and both are sparingly flaked with light yellowish brown.” He gave the measurements as 42.0 X 32.8 and 42.6 X 33.1 mm. This set was reexamined in October 1998, and similar measurements (41.02 X 33.06 and 42.75 X 33.31 mm) were obtained. Both eggs are short subelliptical in shape (Preston in Palmer 1962), and they are dull white with faint light brown splotches and darker brown markings that may be nest stains. Since they differ so conspicuously in both color and size from the Tikal eggs, they were almost certainly misidentified. Hewitt (1937) attributed a single egg taken on 28 April 1936 in the Rio Orinoco District, Venezuela to the Bicol- ored Hawk. The egg was white with faint nest stains and was marked with thin, rust-colored streaks (interpreted as pigmentation, although they were probably blood stains) . It was said to exhibit a greenish tinge when held before a strong light. It measured only 38 X 32.75 mm, thus being about one-third smaller in volume and much more spherical (elongation = 1.16) than the Tikal eggs of this species. The nest was described as a small cup composed of dry sticks and lined with a few leaves and situated at the end of a branch 50 ft above the ground. The bird was said to have been identified on the nest. Seventeen Bicolored Hawk nests at Tikal were located in forks in main tree trunks, where they were supported by large branches or vines, or near the tops of trees sup- ported by several strong branches. None were found at the end of branches, a nesting situation more typical of several Neotropical kite species, and all were well-con- cealed from below. As Hewitt was not known to have per- sonally collected eggs in South America, he presumably obtained the specimen from some other collector or (more likely) from a commercial dealer. Given these dis- crepancies, we conclude that this specimen was misiden- tified. Housse (1937) provided various unconvincing details on nests and eggs purportedly of the Chilean race, Acap- iter bicolor chilensis (often regarded as a separate species, e.g., Johnson 1965, del Hoyo et al. 1994). Nests of the species were said to usually be in the forks of tall trees inside the forest, but close to a clearing or open country, and made of dry or green twigs, closely interlaced and lined with dry grass so as to provide a smooth and round- ed finish to the inner cup. Housse claimed that four eggs was the usual clutch, but that there were sometimes even five or six eggs in a nest. The eggs were described as nearly elliptical, smooth but somewhat thick-shelled, white with large yellowish or greenish blotches concen- trated around the larger pole and with numerous small yellow and gray spots elsewhere. Measurements were giv- en as 44 X 30 mm in the earlier reference (Housse 1937), but as 44 X 34 in Housse (1945). The clutch size, egg measurements, color of egg markings and the nest description is more suggestive of some other species, per- haps the Chimango Caracara (Milvago chimango)] at least they do not agree in these details with known authentic Bicolored Hawk eggs and nests. It is worth mention that the egg size and color description given in the Bicolored Hawk species account in Brown and Amadon (1968), which seem to be for A. bicolor chilensis, were actually those from Wolfe’s (1964) description of Mexican eggs. Schonwetter (1961) gave the measurements of two purported eggs of the race A. bicolor pileatus as 37.0 X 32.6 and 41.3 X 32.6 mm, far below the likely range of measurements for eggs of the Bicolored Hawk. No details on locality or collector were provided. Another egg listed 246 Short Communications VoL. 33, No. 3 by Schonwetter, apparently from western Ecuador and said to be from the Passler collection, has more credible measurements of 46.5 X 35.2 mm, and based solely on these data, it may be authentic. Because the size of this specimen differed from the one reported by Hewitt (1937), Wetmore (1965) regarded (wrongly, we think) the former specimen as genuine and this one as mis- identified. This egg would be from the nominate race bicolor, although birds from western Ecuador have some- times been regarded as a separate race, A.b. schistochlamys (cf. Hellmayr 1949). Wolfe (1964) described an egg of the race Accipiter bi- color jidens he collected in northern Veracruz, Mexico on 6 May 1961, with the assistance of R. Reyes, as “very light bluish gray without markings.” He gave the measure- ments as 46.2 X 36.1 mm. This specimen is in the West- ern Foundation of Vertebrate Zoology collection (WFVZ 16 684), where it now appears as dull white, lacking pig- ment, but with some faint nest stains. It is subelliptical in shape. The nest was placed about 80 ft (24.4 m) from the ground in an old nest of the Common Black-hawk {Buteogallus anthracinus) in thick woods along a stream. The old nest had been lined with a few small twigs and some green leaves. Wolfe was a well-respected authority on raptors and their eggs, and there is no question that this egg is authentic. The original color of the egg, i.e., “light bluish gray,” is typical of many fresh-laid Cooper’s Hawk eggs (Bent 1937), but the bluish color fades during incubation. Wolfe’s specimen was “fresh” when it was col- lected. The Western Foundation collection contains another set (WFVZ 145 391) taken by the late Fred F. Nyc, a Texan oologist who made frequent collecting forays into Mexico from the 1940s to the early 1970s, often in the company of fellow oologists Col. Wolfe and Dr. Travis Meitzen. We knew Nyc well and judged him to be a reliable, knowl- edgeable collector. According to the accompanying data card, Nyc collected the set about 12 mi (19.3 km) south- west of Tampico, Tamaulipas, Mexico “in the oil fields” on 20 March 1970. The eggs were fresh, and the collector suspected that the set was incomplete. Identification of one (or both?) of the birds was based on a half hour of observations with binoculars from a distance of 50 ft. The nest was described as about 10 ft (3.1 m) high in a “strag- gly” mesquite tree in open country, and it was lined with grass and a few green leaves. We examined and measured the Nyc eggs on 18 Oc- tober 1998. Their shape is subelliptical, and they measure 46.85 X 36.32 and 48.36 X 36.02 mm. Both are dull white with a sprinkling of fine brown spots on the large end; one of the eggs is more heavily marked than the other. In all details other than the markings, the egg is virtually identical to one (WFVZ uncat.) we collected at Tikal, al- though it also bears dark nest stains which might be in- terpreted as pigmented markings. Faint brown spots and splotches, many below the outer cuticle and often heavier at the larger end, are typical of Cooper’s Hawk eggs (Bent 1937, Palmer 1962), so it is not unlikely that some Bicolored Hawk eggs also bear such markings. The coastal locality and the open habitat, unlikely for this deep forest species (Howell and Webb 1995), might cast doubt on the authenticity of this set, or at least on the associated data. However, Nyc noted on the data card that he had found a nest of this species containing a clutch of five eggs in the same area on 20 May 1952. On this date, one of the eggs in this clutch had Just hatched, two had pipped and the other two were addled. All were so badly oil-stained that Nyc expressed surprise that they could have remained viable. His field notes indicate that a bird was shot at the nest, presumably to identify it. A clutch size this large would be unexpected for a tropical accipitrid, but perhaps it is typical at the relatively high latitude of southern Tamaulipas. In summary, we regard the specimens of Chubb (Par- aguay), Nyc (Mexico) and Wolfe (Mexico) to be authen- tic, and the eggs reported by Schonwetter from Ecuador were probably authentic. Based on a comparison with the Tikal sample and other details, the eggs reported by Hew- itt (Venezuela) , Schonwetter (A. b. pileatus) and Smooker (British Guiana) were probably misidentified. The am- biguous details provided by Housse on eggs from Chile do not permit any conclusion to be drawn about their authenticity. These examples illustrate both the strengths and weak- nesses of museum collections. Younger researchers, who may not have grown up in a collecting tradition, should regard museum specimens as an essential resource, but they should view specimen data (especially for egg sets) with the same critical eye they would apply to any other data set. Ideally, all trained field workers in poorly stud- ied regions should be encouraged to add material to mu- seum collections as possible, even if only opportunisti- cally from salvaged material, since even for such a widespread species as the Bicolored Hawk, which occurs from NW Mexico south to Tierra del Fuego, there is a dearth of usable oological specimens with reliable mod- ern data. Resumen. — Una comparacion de las descripciones pu- blicadas de huevos de Accipiter bicolor con las caracteristi- cas 14 huevos conocidos en seis nidos guatemaltecos es- tudiados entre 1991-94, indica que muchos de los reportes anteriores fueron probablemente basados en la identificacion equivocada de especimenes. Los huevos de Accipiter reportados en Paraguay (Chubb), Mexico (Nyc, Wolfe) y Ecuador (Schonwetter) parecen autenti- cos, pero los especimenes de Venezuela (Hewitt) y Guy- ana (Smooker) estan erroneamente clasificados. La iden- tificacion actual de huevos de Chile (Housse) es errada. [Traduccion de Cesar Marquez] Acknowledgments This study was part of The Peregrine Fund’s Maya Pro- ject, conducted in cooperation with the Institute Nacio- September 1999 Short Communications 247 nal de Antropologia y Historia (IDEAH), Centro de Es- tudios Conservationistas (CECON), and Consejo Nacional de Areas Protegidas (CONAP) of Guatemala. We thank Bill Burnham, David Whitacre, and J. Peter Jenny of The Peregrine Fund and participants in the Boi- se State University Raptor Biology program for various forms of assistance. Measurements of the eggs in the Western Foundation of Vertebrate Zoology collection were obtained with the assistance of Rene Corado and Sam Sumida. An earlier draft was much improved by the suggestions of Rob Bierregaard, C. Stuart Houston, and Manuel Marin. Literature Cited Bent, A.C. 1937. Life histories of North American birds of prey. Order Falconiformes (Pt. 1). U.S. Natl. Mus. Bull. 167:1-409. Boyce, D.A., Jr. and L.F. Kief. 1981. Have the eggs of the Orange-breasted Falcon {Falco deiroleucus) been de- scribed? Raptor Res. 15:89-93. Brown, L. and D. Amadon. 1968. Hawks, eagles, and fal- cons of the world. Vol. 2. McGraw-Hill, New York, NY U.S.A. Chubb, C. 1910. On the birds of Paraguay. Ibis 4:53-78, 263-285, 571-647. del Hoyo, H., a. Elliott and J. Sargatal. 1994. Hand- book of the birds of the world. Vol. 2. New World Vultures to Guineafowl. Lynx Edicions, Barcelona, Spain. ffrench, R. 1973. A guide to the birds of Trinidad and Tobago. Livingston Publ. Co., Wynnewood, PA U.S.A. Hellmayr, C.E. and B. Conover. 1949. Catalogue of birds of the Americas and the adjacent islands. Pt. I, no. 4. Cathartidae-Accipitridae-Pandionidae-Falconi- dae. Field Mus. Nat. Hist. Publ. Zool. Ser. 13(pt. 1, no. 4):l-358. Hewitt, V. 1937. A description of the eggs of Grey Crane- hawk, Guiana White-tailed Hawk, Eyebrowed Hawk, and Four-banded Sparrow Hawk. Oologists’ Rec. 17:12— 14. Housse, P.R. 1937. El Peuquito, Accipiter chilensis. Rev. Chil. Hist. Nat. 12:134-140. . 1945. Las aves de Chile en su classificacion mo- derna. Ediciones de la Univ. de Chile, Santiago, Chile. Howell, S.N.G. and S. Webb. 1995. The birds of Mexico and northern Central America. Oxford Univ. Press, Oxford, U.K. Johnson, A.W. 1965. The birds of Chile and adjacent re- gions of Argentina, Bolivia, and Peru. Vol. 1. Platt, Buenos Aires, Argentina. Kiff, L.F. 1981. Notes on the eggs of the Hook-billed Kite Chondrohierax uncinatus, including two overlooked nesting records. Bull. Br. Ornithol. Club 101:318-323. Palmer, R.S. 1962. Handbook of North American birds. Vol. 1. Yale University Press, New Haven, CT U.S.A. Schonwetter, H. 1961. Handb. Oologie 1:160. Wetmore, a. 1965. The birds of the Republic of Panama. Smithson. Misc. Collect. 150:190-192. Wolfe, L.R. 1964. Eggs of the Falconiformes further part. Oologists’ Rec. 38:50—57. Received 6 August 1997; accepted 2 April 1999 J Raptor Res. 33(3):248-251 © 1999 The Raptor Research Foundation, Inc. Nesting Ecology of Red-shouldered and Red-tailed Hawks in Georgia Christopher E. Moorman^ Daniel B. Warnell School of Forest Resources, University of Georgia, Athens, GA 30602-2152 US. A. Doug L. Howell PWBC, AFZA-PW-EW, Wildlife Branch, Fort Bragg, NC 28307-5000 U.S.A. Brian R. Chapman Daniel B. Warnell School of Forest Resources, University of Georgia, Athens, GA 30602-2152 U.S.A. Key Words: Red-shouldered Hawk, Red-tailed Hawk, Buteo lineatus; Buteo jamaicensis; reproductive success, nest-site reoccupancy’, nesting chronology. Because forest raptors occupy the top of many food chains, have relatively large home ranges and have a his- tory of human persecution, they generally are among the first groups of vertebrates to respond to alterations of habitat (Fuller 1996, Niemi and Hanowski 1997). There- fore, information on the biology and nesting ecology of forest raptors is needed to predict or mitigate possible impacts of habitat alterations associated with forest man- agement practices. This is especially true in the south- eastern U.S., where large areas are under intensive tim- ber management and little is known about the regional biology of resident raptor species. Although Breeding Bird Surveys indicate that resident populations of Red-shouldered {Buteo lineatus) and Red- tailed {B. jamaicensis) Hawks are stable or increasing in Georgia (Preston and Beane 1993, Crocoll 1994), little has been published concerning their nesting ecology in the southeastern U.S. Moorman and Chapman (1996) compared the macro- and micro-habitat factors that in- fluence nest-site selection of the two species, and Howell and Chapman (1997) described the home range and habitat use of nesting Red-shouldered Hawks in central Georgia. Here, we present data on the nesting chronol- ogy, nesting success and nest reoccupancy rates for both species in Georgia. Study Area and Methods Our study was conducted in 1994 and 1995 at the Bish- op F. Grant Memorial Forest (BGMF), a 5718-ha state wildlife management area located in Putnam County, Georgia. The BGMF contains a diversity of habitat types including bottomland and upland hardwoods, various- aged pine stands and pastures (Moorman and Chapman 1996, Howell and Chapman 1997). We searched for old hawk nests between January and ^ Present address: Extension Forestry, North Carolina State University, Box 8003, Raleigh, NC 27695-8003 U.S.A. early March 1994, and returned beginning in mid-March to check for signs of nesting activity. We continued to search for occupied nests through mid-June 1994. We lo- cated hawk nests by searching areas where they were seen perching or soaring. To prevent bias, nest searches also were conducted in all forested stands with trees >20 yr old where birds were not seen or heard. Nests were con- sidered occupied if incubating females down or young were observed in nests. Young could not be documented during ground visits until they were approximately 1 wk old. We monitored occupied nests of both species every 7-10 d and recorded the status (e.g., incubating, small young, large young) . A nest was considered successful if it fledged at least one young. Actual fledging dates were calculated as the median day between the last visit when young were present in nests and the visit when nests were vacant. Because nests were found throughout the breed- ing season, data from the early portion of the nesting season was lacking for some pairs. Return visits were made during 1995 to check activity of 1994 nests and to make observations on hawk nesting ecology. Nest revisits began on 31 March 1995 and were continued every 10 d until the end of June. Results Red-shouldered Hawk Nesting Ecology. Twelve Red- shouldered Hawk nests were found in 1994. Incubation was first observed on 24 March and it continued through 18 May at one nest. Nestlings were first observed on 25 April and most nests contained young by 7 May. Fledgling hawks were first observed on 5 June and all successful nests had fledged young by 25 June. Four nests failed to fledge young, two nests fledged one young and six nests fledged two young. In 1995, we revisited 10 Red-shouldered Hawk nests used in 1994. The two other nests from 1994 either had been damaged or had fallen from trees at the end of the 1994 breeding season so they were not revisited in 1995. We located occupied nests in six of the 10 nesting terri- tories used the previous year. Two pairs repaired and used alternate nests within the same breeding territory occupied in 1994, and four others reused 1994 nests. Three female hawks were banded and radio-collared in 1994 (Howell and Chapman 1997). In 1995, banded birds were seen entering or leaving nests in all three ter- 248 September 1999 Short Communications 249 Nestling Period Incubation Period RSHA- Species Fledging Begins 5 June RTHA- Fledging Begins 25 May 1 1 1 — February March April — I 1 1 — May June July — I August Figure 1. Nesting chronology of Red-shouldered (RSHA) and Red-tailed Hawks (RTHA) in Georgia. ritories in which females were marked during the previ- ous season, indicating that the same females reoccupied nests used in 1994. Nesting activity in 1995 was first ob- served on 6 April. Red-shouldered Hawk young were first observed on 18 April and all nests contained nestlings by 12 May. All Red-shouldered Hawk young fledged by 25 June. One nest failed, four nests fledged two young and one fledged three young. Red-tailed Hawk Nesting Ecology. Ten Red-tailed Hawk nests were located during the 1994 breeding season. Signs of incubation were first observed on 7 April but, because we failed to locate most Red-tailed Hawk nests early in the nesting season, incubation probably started earlier. Nestlings were first observed on 15 May and an unfledged nesding was last observed in a nest on 15 June. Fledglings were first observed on 14 June and most re- mained near the nest for several more weeks. Five of 10 Red-tailed Hawk nests fledged one young. Remains of dead nesdings were found under two of the nests that fledged young. Eight Red-tailed Hawk nests used in 1994 were revisit- ed in 1995. The two other nests from 1994 were not re- visited for logisdeal reasons. Six of eight 1994 Red-tailed Hawk nests were reoccupied. One nest occupied by Great Horned Owls {Bubo vir^nianus) in March 1995 was not used, although a pair of Red-tailed Hawks observed repeatedly in the area during later visits that year. During our inidal visit on 31 March 1995, signs of incubadon were observed at four of the nests used in 1994. Red- tailed Hawk nesdings were first observed on 12 May. Sev- eral of these nesdings were well developed and appeared to be approximately 2 weeks old. Fledgling Red-tailed Hawks were first observed on 6 June. Three of six nests fledged one young, two fledged two young and one nest failed. Discussion Nesting chronology for both hawk species in Georgia likely varies by individual pair and year. However, on av- erage, Red-shouldered Hawks begin incubation in late March, eggs hatch in late April to early May, and young fledge in early to mid-June (Fig. 1). Red-tailed Hawk nest- ing generally begins 2 wk earlier (Fig. 1). Burleigh (1958) reported incubation by a Red-shouldered Hawk as early as 22 March and as late as 1 May, and incubation by a Red-tailed Hawk as early as 15 February and as late as 10 April in Georgia. We documented later incubation dates for both species, although the birds observed incubating in our study might have been brooding small nestlings. Red-shouldered Hawks often fledge up to four young, but the average number/ breeding pair usually ranges from 1.1-1. 8 (Crocoll 1994). Red-tailed Hawk nest suc- cess ranges from 58-93%, but the number of young fledged/breeding pair is usually between 1.0-1 .5 (Pres- 250 Short Communications VoL. 33, No. 3 ton and Beane 1993). During the 1994 breeding season, Red-shouldered Hawk nesting success (66%) and num- ber of fledglings/breeding pair (1.17) were similar to that of previous studies, but Red-tailed Hawk nest success (50%) and number of fledglings/breeding pair (0.5) were lower than reports from other parts of North Amer- ica. Red-tailed Hawk dutch sizes generally increase from south to north and from east to west (Henny and Wight 1972, Preston and Beane 1993), so the low productivity may not be unusual for the southeastern U.S. These low numbers also may suggest that prey abundance was low during the hrst year of our study (Preston and Beane 1993), but data on prey abundance were unavailable. Two dead young were found below Red-tailed Hawk nests m 1994, which may indicate nestling starvation or sibling aggression (Stinson 1980), Records from our return visits in 1995 indicated that nest success (83% and 83%) and number of fledglings/breeding pair (1.83 and 1.17) in- creased for both Red-shouldered and Red-tailed Hawks, respectively. However, our 1995 results could be biased because early nest failures or pairs in new territories were not monitored. Breeding-site fidelity is common in both species (Bent 1937, Preston and Beane 1993, Crocoll 1994). In Wiscon- sin, Jacobs and Jacobs (1993) documented 50% nest reoccupancy and 83% territory reoccupancy by Red- shouldered Hawks. Burleigh (1958) reported one inci- dent of traditional nest-site use by Red-shouldered Hawks and one account of nest-site reoccupancy by the Red- tailed Hawk in Georgia. However, nest-site reoccupancy rates have not been reported for the two hawk species in the southeastern U.S. Forty percent of the Red-shoul- dered Hawk nests used in 1994 were reoccupied in 1995, and Red-shouldered Hawks occupied alternate nests in at least two more of the 10 territories used in 1994 (>60% territory reoccupancy). Seventy-five percent of the nests used in 1994 by Red-tailed Hawk pairs were reoccupied in 1995. We could not measure fidelity be- cause hawks were not uniquely marked, but the presence of banded female Red-shouldered Hawks indicated that the same females returned to the nesting territories oc- cupied the previous year. Red-tailed Hawk breeding territories associated with ephemeral, early-successional habitats such as clearcuts may shift with changes in the vegetative structure of these stands. However, nests located near permanent openings such as pastures may be reoccupied for many consecutive years. Red-shouldered Hawks, which nest in mature, more permanent bottomland forests (Bednarz and Dins- more 1981, Moorman and Chapman 1996), likely also maintain traditional territory boundaries for long peri- ods. Conservation of areas containing traditional Red- shouldered Hawk breeding territories may help prevent replacement by Red-tailed Hawks (Bednarz and Dins- more 1982). Resumen. — La ecologia de anidacion de Buteo lineatus y Buteo jamaicensis ha sido poco conocida en el sureste de los Estados Unidos. Documentamos la cronologia de ani- dacion, exito de anidacion y reocupacion de nidos para ambas especies entre 1994—95. Durante la epoca repro- ductiva de 1994, el exito de anidacion (66%) de Buteo lineatusy de productividad (1.17 pichones/pareja reprod- uctiva) fue muy inferior a los valores reportados anter- iormente. Las nuevas visitas a los nidos y las observa- ciones adicionales en 1995, indicaron que el exito reproductive incremento para las dos especies. Los cam- bios en el exito reproductivo de las dos especies pudo haber sido causado por fluctuaciones en las poblaciones de presas. Las tasas de re-ocupacion de nidos fueron del 75% y 40% para Buteo jamaicensis y Buteo lineatus respec- tivamente. La conservacion de las areas tradicionales de anidacion puede ser requerida para mantener las pobla- ciones locales de gavilanes en reproduccion. [Traduccion de Cesar Marquez] Acknowledgments We thank J.H. Brunjes, C.W. Eberly and C. Howell. This manuscript was improved by comments provided by D.S. Mizrahi, S.M. Lohr, and K.R. Russell. Funding was provided by the Daniel B. Warnell School of Forest Re- sources at the University of Georgia and Mclntire Stennis Project No. GEO-0074-MS. Literature Cited Bednarz, J.C. andJ.J. Dinsmore. 1981. Status, habitat use, and management of Red-shouldered Hawks in Iowa. J. Wild. Manage. 45:236—241. AND . 1982. Nest-sites and habitat of Red- shouldered and Red-tailed Hawks in Iowa. Wilson Bull 94:31-45. Bent, A.C. 1937. Life histories of North American birds of prey. Pt. 1. U.S. Natl. Mus. Bull. 167, Washington, DC U.S.A. Burleigh, T.D. 1958. Georgia birds. Univ. Oklahoma Press, Norman, OK U.S.A. Crocoll, S.T. 1994. Red-shouldered Hawk {Buteo linea- tus). In A. Poole and F. Gill [Eds.], The birds of North America, No. 107. The Philadelphia Academy of Nat- ural Sciences, Philadelphia, PA U.S.A. and American Ornithologists’ Union, Washington, DC U.S.A. Fuller, M.R. 1996. Forest raptor population trends in North America. Pages 167-208 in R. DeGraaf and R.I Miller [Eds.], Conservation of faunal diversity in for- ested landscapes. Chapman-Hall, London, U.K. Henny, C.J, and H.M. Wight. 1972. Population ecology and environmental pollution: Red-tailed and Coo- per’s Hawks. Pages 229—250 in Population ecology of migratory birds: a symposium. U.S. Dept. Int. Wildlf Res. Rep. 2, Washington, DC U.S.A. Howell, D.L. and B.R. Chapman. 1997. Home range and habitat use of Red-shouldered Hawks {Buteo lineatus) in Georgia. Wilson Bull. 109:131-144. Jacobs, J. and E. Jacobs. 1993. Summary of Red-shoul- dered Hawk reproduction in northeastern and cen- tral Wisconsin. Wisconsin Dept. Natural Resources, September 1999 Short Communications 251 Bureau of Endangered Resources, Madison, WI U.S.A. Moorman, C.E. and B.R. Chapman. 1996. Nest-site selec- tion of Red-shouldered and Red-tailed Hawks in a managed forest. Wilson Bull. 108:357—368. Niemi, G.J. and J.M. Hanowski. 1997. Preface raptor re- sponses to forest management: a Holarctic perspec- tive. /. Raptor Res. 31:93-94. Preston, C.R. and R.D. Beane. 1993. Red-tailed Hawk {Buteo jamaicensis) . In A. Poole and F. Gill [Eds.], The birds of North America, No. 52. The Philadelphia Academy of Natural Sciences, Philadelphia, PA U.S.A. and American Ornithologists’ Union, Washington, DG U.S.A. Stinson, C.H. 1980. Weather-dependent foraging success and sibling aggression in Red-tailed Hawks in central Washington. Condor 82:76-80. Received 9 September 1998; accepted 13 March 1999 J. Raptor Res. 33(3):251-254 © 1999 The Raptor Research Foundation, Inc. The Red Kite {Milvus milvus) Reintrodugtion Project: Modeling the Impact oe Translocating Kite Young Within England Ian Carter, Mick McQuaid, Nigel Snell and Peter Stevens English Nature, Northminster House, Peterborough PEI lUA, U.K. Key Words: Red Kite, Milvus milvus; reintroduction; trans- locations; England. The Red Kite (Milvus milvus) reintroduction project started in 1989 with the release of six Swedish kites at a site in northern Scotland and four Swedish and one Welsh kite at a site in southern England (Evans et al. 1991). From 1989-94, a total of 93 birds were released at each site with southern England birds coming mainly from Spain and northern Scotland birds from Sweden (Evans et al. 1997). As a result of these releases, small populations have been successfully established in both release areas. In 1997, the southern England breeding population reached about 55 pairs and is now considered to be self-sustaining. In order to improve the status of Red Kites in Britain and to increase their spread to other suitable areas, re- leases have started at two new sites in central Scotland and the English midlands (Garter 1995). At the latter site, a total of 29 birds, mainly from Spain, were released in 1995 and 1996. In 1997, another 10 kites from Spain were released but, due to concerns about declines in the Red Kite population in parts of Spain, Spanish authori- ties decided that it would be difficult for them to supply kites for the project in future years. Discussions with the recently formed Welsh Kite Trust led to an agreement that young rescued from vulnerable nests and for which suitable foster nests could not be found within Wales would be made available for translocation to the English midlands release site. However, this would involve only a few birds and, in some years, no young would be available for the reintroduction project. In order to make up this short-fall of birds for release, the translocation of young kites from the expanding southern England population was considered. To help assess the impact of any such translocations, we devised a simple model to show the likely effects on the southern England and midlands populations. Various scenarios were modeled, reflecting the range of options available. Methods The model used the following data and assumptions based mainly on monitoring work on the expanding southern England kite population up to 1996 (Evans et al. 1997, N. Snell, M. McQuaid and R Stevens unpubl. data): (1) 76% survival in the first year based on sightings of individually-identifiable, wing-tagged kites, released be- tween 1989-94 (A = 93); (2) 93.5% adult survival based on sightings of individually-identifiable, wing-tagged kites in their second and subsequent years (N = 136); (3) breeding productivity of 2.1 young per breeding attempt between 1991-96 (N = 94); (4) balanced immigration and emigration (this seemed reasonable because Red Kites are known to have a very high level of natal philo- patry [Newton et al. 1994]. No wing-tagged kites released or fledged in southern England have yet been found breeding elsewhere, although because some released birds have now lost their tags and not all young are fitted with tags each year, it is possible that a small number of cases have gone undetected; at least one continental im- migrant is known to have been recruited into the south- ern England breeding population [I. Evans pers comm.]); (5) age of first breeding at two years (in the southern England population, kites have occasionally bred in their first year but normally attempt to breed for the first time in their second year; in Wales where the habitat is less suitable, kites have been recorded breeding 252 Short Communications VoL. 33, No. 3 Table 1. Estimated number of breeding pairs of Red Kites in the southern England and the English midlands using the model assuming that 20 young kites are translocated from the southern England population. Year S. Eng. — Do Nothing S. Eng. — Remove 20 Chicks in 1997 Midlands — Release No Birds in 1997; 20 IN 1998 AND 1999 Midlands — Release 20 Birds Each Year 1997 TO 2000 1997 50 50 3 3 1998 75 75 9 9 1999 108 100 11 18 2000 157 150 24 30 2001 227 215 37 48 2002 329 313 52 68 for the first time at up to seven years of age [Newton et al 1987]); (6) equal sex ratio (a population of 40 birds, two years or older was assumed to have 20 breeding pairs) . Since it was a simple, deterministic model, density-de- pendent effects and stochastic events were not taken into account, although their implications were considered. The starting point for each scenario was the estimated kite population in southern England and the midlands in 1997, including the number of breeding pairs and the number of surviving young fledged from nests in 1996. For each year (x) , the population in the following year (x + 1) was calculated by summing the following values derived from the data and assumptions: (1) number of breeding adults surviving from year x to year x + 1; (2) number of young fledged in year x — 1 surviving to breed for the first time in year x + 1; (3) number of young fledged in year x surviving to year x -I- 1. In each case, the number of young removed or added to the popula- tion under a given scenario was taken into account in (2) and (3). Results Scenario 1. Remove 20 young from the southern Eng- land population in 1997 and release them in the mid- lands. We assumed that no further birds were translocat- ed from southern England in subsequent years but 20 birds from an alternative donor population were released m the midlands in 1998 and 1999. By the year 2002, this translocation resulted in an increase of 31% in the num- ber of pairs in the midlands compared to the release of no birds in 1997 (Table 1). The removal of the 20 birds from southern England resulted in a 4.9% decrease in that population. The effect of releasing the extra birds in the midlands was most noticeable in the years 1998 to 1999 when the population increased from nine to 18 pairs as opposed to an increase from nine to only 1 1 pairs if no birds were released. Scenario 2. In this scenario, no further birds were avail- able for release in the midlands from sources outside southern England. Fifteen birds were taken from south- ern England and released in the midlands in each year between 1997—99. In this example, the release of 15 birds in the midlands each year between 1997-99 resulted in a 74% increase in the midlands population by 2002 com- pared to doing nothing (Table 2). The loss of the 15 birds in each of three years from southern England re- sulted in a population reduction from 328 pairs to 303 pairs in 2002, a difference of 8%. In both the above scenarios, because the model does not take into account any density-dependent effects, the number of kites gained by the midlands population ex- actly matches the number lost to the southern England population. The percentage difference was, however, much greater for the midlands population than the well- established southern England population. Table 2. Estimated number of breeding pairs of Red Kites in southern England and the English midlands assuming that 15 birds are taken from southern England and released each year from 1997-99. Year S. Eng. — Do Nothing S. Eng. — Remove 15 Chicks in 1997, 1998 and 1999 Midlands — Do Nothing Midlands — Release 15 Birds in 1997, 1998 and 1999 1997 50 50 3 3 1998 75 75 9 9 1999 108 102 11 16 2000 157 146 17 27 2001 227 207 23 42 2002 329 303 34 59 September 1999 Short Communications 253 Discussion One of the requirements of any well-planned reintro- duction project is that it should be adequately monitored (lUCN 1987). This not only ensures that any problems will be identified and resolved at an early stage but also facilitates decisions about the future of the project Use of this simple model was only possible because the basic population and survival data were available from moni- toring work on the southern England Red Kite popula- tion. The model proved to be a valuable aid to decision making when considering the various options for trans- locating Red Kites for release in the midlands and, in particular, the option of taking birds from southern Eng- land in 1997. Having used the model to help consider the potential impact of translocating 10 birds in 1997, kite workers involved with the project agreed that the translocation should go ahead. It was accepted that the translocated birds would significandy improve the status of the small, vulnerable midlands population without having a significant impact on the larger donor popula- tion. In effect, the birds were thought to be more valu- able in helping to meet the project’s overall aim of re- storing kites to all suitable habitats in Britain if they were translocated and released in the midlands. Ten young were taken under licence from nests in June 1997 and have been released into the midlands along with the 10 birds imported from Spain. In order to minimize any impact on the southern Eng- land population, only the smallest young were taken from broods of two or three (broods of four occur only very occasionally) . In one case, the two smallest young were taken from a brood of three. With many species of birds of prey and owls, the smallest young are vulnerable to being eaten by their siblings if there is a shortage of food (Newton 1979, Cramp and Simmons 1980, Cramp 1985, Watson 1997), and this has been recorded in the Welsh kite population (Lovegrove et al. 1990). Most deaths due to aggression from siblings would be expected when young are still small, and deaths are much less likely when the chicks are four wk or older, the age at which the kites were collected from nests. Nevertheless, it is still possible that some young taken from southern England would not have fledged successfully if left in nests. In captivity, it was possible to provide the young with a sur- plus of food and prevent any problems due to food short- age. Because our model is purely deterministic, the year to year population changes were solely dependent on the set of population parameters derived from monitoring the southern England population. No possible effects of chance, stochastic, events acting on the population were considered. While chance events are unlikely to effect the relatively large southern England population signifi- cantly, this is certainly not the case in the early years in the midlands while the population is still small. An out- break of disease in the midlands could wipe out the tiny breeding population completely by causing the deaths of only a handful of adult kites. This emphasizes the im- portance of ensuring that the vulnerable midlands pop- ulation increases to a level at which such stochastic effects are less significant. The model considered the impact of the translocation options on Red Kites in southern England and the mid- lands separately. However, given the stated aims of the project to reestablish the Red Kite throughout Britain, the translocation of birds should not be viewed as a loss to one area and a gain for another. In the long-term, translocated birds will form part of a single, larger British population. Although unrealistic, it is interesting to run the model for a longer period of time. If this is done for the southern England population under the “do noth- ing’’ scenario, then the population would reach 14250 pairs in 2012 and 93 700 pairs by 2017. The population will clearly not reach such levels as quickly as predicted by the model because we can expect increasing compe- tition among kites as the population expands in both numbers and range. This would likely increase age of first breeding, reduce levels of breeding productivity and re- duce survival rates (Newton 1979), thus slowing the rate of population increase. Resumen. — Un modelo deterministico simple fue utili- zado para evaluar los resultados potenciales del traslado de Milvus milvus en Inglaterra como parte del proyecto de reintroduccion de la English Nature/Royal Society for the Protection of Birds (RSPB). El modelo utilizo datos de sobrevivencia y productividad para el monitoreo de la poblacion en expansion al sur de Inglaterra y demostro que el traslado de pequenos numeros de aves tienen poco efecto en esta poblacion, pero si para la poblacion de la region central. Dos escenarios incluyendo el tras- lado de distintos numeros de aves en periodos diferentes son presentados como ejemplos de como el modelo puede ser usado para evaluar las diferentes opciones. Al utilizar los resultados del modelo, la decision fue la de trasladar los primeros 10 juveniles del sur de Inglaterra a la region central en 1997. [Traduccion de Cesar Marquez] Acknowledgments The Red Kite Project in England is jointly funded by English Nature (as part of its Species Recovery Pro- gramme) and the RSPB. The project would not be pos- sible without the involvement of many other individuals and organizations including Forest Enterprise (Karl Ivens and others) , Dr. James Kirkwood and Dr. Sue Thornton at the London Institute of Zoology, the Joint Nature Con- servation Committee, the British Airways Assisting Con- servation Programme (Rod Hall MBE), the Institute of Terrestrial Ecology, Tony Cross at the Welsh Kite Trust, and many others. We are particularly grateful to Spanish workers in Direccion General del Medio Natural, Sego- via, within the Junta de Castilla y Leon and members of 254 Short Communications VoL. 33, No. 3 the Southern England Kite Group for helping to collect kite chicks for release in the midlands in 1997. The paper benefitted greatly from the comments of Dr. Ian Newton, Tom J. Cade and Harrison Tordoff. Literature Cited Carter, I,C., I.M. Evans and N. Crockford. 1995. The Red Kite re-introduction project in Britain — progress so far and future plans. Br. Wildl. 7:18-25. Cramp, S. 1985. Handbook of the birds of Europe, the Middle East and North Africa; the birds of the West- ern Palearctic. Vol. 4. Oxford Univ. Press, Oxford, U.K. and K.E.L. SiMMONS. 1980. Handbook of the birds of Europe, the Middle East and North Africa: the birds of the Western Palearctic. Vol. 2. Oxford Univ. Press, Oxford, U.K. Evans, I.M., M.W, Pienkowski and R.H. Dennis. 1991. Experimental re-introduction of Red Kites: report to 1990. NCC CSD Res. Rep. No. 1224, Peterborough, U.K. , R.H. Dennis, D.C. Orr-Ewing, N. Kjellen, P. An- dersson, M. Sylven, a. Senosiain and EC. Carbo. 1997. The re-establishment of Red Kite Milvus milvus breeding populations in Scotland and England. Br. Birds 90:123-138. Lovegrove, R., G. Elliot and K. Smith. 1990. The Red Kite in Britain. In C.J. Cadbury [Ed.], RSPB Conser- vation Review 1990. Sandy, U.K. Newton, I. 1979. Population ecology of raptors. T. & A. D. Poyser, London, U.K. , P.E. Davis and J.E. Davis. 1987. Age of first breeding, dispersal and survival of Red Kites Milvus milvus in Wales. Ibis 131:16-21. , and D. Moss. 1994. Philopatry and pop- ulation growth of Red Kites Milvus milvus in Wales. Proc. R. Soc. Land. B, 257:317-323. International Union for the Conservation of Na- ture AND Natural Resources. 1987. Translocation of living organisms: introductions, re-introductions and re-stocking. lUCN, Gland, U.K. Watson, J, 1997. The Golden Eagle. T. & A. D. Poyser, London, U.K. Received 29 August 1998; 11 April 1999 J. Raptor Res. 33(3):254-257 © 1999 The Raptor Research Foundation, Inc. First Recorded Polygynous Mating in the Red Kite {Milvus milvus) Hein Van Kleef^ and Javier Bustamante Department of Applied Biology, Estacion Biologica de Donana, CSIC, Avda. Maria Luisa s/n, 41013 Sevilla, Spain Key Words: Red Kite, Milvus milvus; bigamy; polygyny; Donana, Spain. Diurnal birds of prey are predominantly monogamous (Newton 1979). Alternative mating systems like polygyny, polyandry, or cooperative breeding are rare, but have been recorded in at least 16 species of raptors (Newton 1979, Faaborg and Bednarz 1990, Heredia and Donazar 1990, Telia 1993, Telia et al. 1996). Nonmonogamous re- lationships are easily overlooked when working with un- marked individuals and are almost certainly more wide- spread than published records show (Newton 1979). Polygyny in birds of prey has only been regularly ob- served in harriers {Circus spp.) although it has been oc- casionally recorded in another nine species (Newton 1979, Hiraldo et al. 1991, Telia et al. 1996) and seems to ^ Present address: Burghardt van den Berghstraat 163, 6512 DK, Nijmegen, The Netherlands. be associated with relatively productive habitats with an abundant food supply. The Red Kite {Milvus milvus) is considered a monoga- mous raptor and to our knowledge no instances of polyg- yny have been reported previously (Glutz von Blotzheim et al. 1971, Newton 1979, Cramp and Simmons 1980). Ac- cording to Glutz von Blotzheim et al. (1971) and Cramp and Simmons (1980), both adults build nests. Incubation is mainly done by females although males may incubate for short periods during the day. Males bring prey to fe- males and defend nest sites during incubation and the first two weeks after hatching while females brood and feed the young. Later, both members of pairs defend nest sites and bring food to nests, where the young feed themselves. On average, young fledge 55 d after hatching and are fed by both parents for another 26 d in the vicinity of the nest (Bustamante 1993). The entire hunting territory is not de- fended, but Red Kites defend areas surrounding nest sites at least until the young become independent (Bustamante and Hiraldo 1993). In 1996 and 1997, we recorded the presence of a po- September 1999 Short Communications 255 lygynous trio of Red Kites in the area around the “Palacio de Donana,” Dohana Biological Reserve, southwest Spain (37°N, 6°30'W; Valverde 1958, Rogers and Myers 1980) . The male could be identified by a white PVC band on the left leg with the code “2AU” that could be read with a telescope from approximately 200 m. This bird was banded as a nestling by L. Garcia on 6 May 1977 in a pine forest named “Pinar de San Agustin,” <1000 m from the “Palacio de Donana.” The bird was re trapped and observed several times during its life close to this pine forest and was documented breeding in this area at least since 1984 (B. Heredia, L. Garcia, G. Vilchez, F. Hiraldo, M. Pineda and R. Lafitte pers. comm.). One fe- male (FI) associated with this male was not banded and could only be identified due to her proximity to one of the nests. The other female (F2) using the territory had a metal band on the right leg, but we could not read this band. Due to the absence of differences in coloration or molting between the individuals, they could only be iden- tified when they were perched in plain sight. Observations in 1996 Observations in March 1996 indicated that there was a polygynous trio in the area around the “Palacio de Donana.” On 8 March 1996, from 0757-1200 H, we were simultaneously observing what we thought were two dif- ferent pairs of breeding Red Kites. -One pair was ob- served building a nest in a stone pine {Pinus pinea) inside a small pine forest named “Pinar de San Agustin.” One kite was banded with a white PVC band on the left leg that was read with the telescope as “2AU.” The other kite (FI) was not banded. During observations, both kites were observed bringing material to the nest. On one oc- casion, they copulated indicating they were pair-bonded and we deduced “2AU” was the male and FI the female. A second nest being observed was in a tall eucalyptus tree {Eucalyptus globulus) approximately 50 m east of the “Pa- lacio de Donana” and 800 m east of the first nest. Most of the time only one kite (F2) was observed in the area, flying and perching on the nest tree. This kite was band- ed with a metal band on the right leg. At 0934 H, a sec- ond kite, banded with white PVC band on the left leg, was seen perched 100 m from the nest. At 0959 and 1002 H, the kites copulated twice in another eucalyptus close to the nest tree. The copulation positions indicated that F2 was a female but we were not able to read the PVC band code of the male with a telescope. At 1005 H, the male with the white PVC band flew away. On 20 March 1996, F2 was observed copulating with the male with a white PVC band and both kites brought sticks to the nest in the eucalyptus tree. On 22 March 1996 at 1026 H, the nest was observed while F2 incubated. The male with the white PVC band flew to the nest and copulated with F2. On this occasion, the PVC band of the male was identi- fied as “2AU.” At 1234 H on the same day, a blind was set close to the nest occupied by FI in the pine forest. At 1353 H, the male “2AU” entered the nest in the stone pine with prey and began incubating. The male “2AU” was still incubating at 1415 H when observations were stopped. The nest tended by F2 could be checked from an ob- servation tower nearby. F2 laid one egg that failed to hatch for unknown reasons. We did not check the nest used by FI until 8 May 1996. We found it contained two young that hatched on 20 and 21 April based on the length of their 7th primary feathers (Bustamante 1990) Both young fledged. Observations in 1997 We made opportunistic observations of the polygynous trio from 9 February-9 March. The male “2AU” was seen copulating with both an unbanded female on three oc- casions (9, 10, and 11 February) and a female with a metal band on the right leg on three occasions (13 Feb- ruary, 9 and 13 March). We suspect they were the same females, FI and F2, from 1996. FI was seen copulating and perched close to the male “2AU” in the same pine forest as in 1996 and F2 was seen bringing nest material to a nest in a poplar {Populus alba) 100 m north of the 1996 F2 nest. Only on one instance was the male seen bringing nest material to the pine forest nest. The FI nest was located in a stone pine 50 m west from the 1996 nest The distance between FI and F2 nests was approximately 750 m. Both nests were observed on a regular basis from 24 April-26 June, during which time the young of both nests grew from a few days old to fledging. Observations were carried out from a blind and were restricted to days that were mostly without clouds and low wind speeds to min- imize the effects of weather conditions on the hunting success and feeding frequency of the kites. During this period, FI nest was observed for a total of 33.4 hr on 11 different days (x = 3.0 ± 2.0 hr, ±SD). F2 nest was ob- served for a total of 41.9 hr on 11 different days (;« = 3.8 ± 1.6 hr, ±SD). The FI nest contained two young that were banded on 30 April. We estimated that the young had hatched on 7 and 12 April. A total of 10 prey was observed delivered to the nest during 33.4 hr of observation; five were brought by the male “2AU,” two by the female FI. In three instances, it was not possible to identify the individ- ual bringing the prey. F2 nest contained one young that was banded on 17 May. We estimated hatching date to have been 24 April. We recorded eight prey delivered to the nest during 41.9 hr of observation and all were brought by the female F2. The male “2AU” was never seen at F2 nest. We approached both nests on several occasions. FI nest was defended by two adults (kites hov- ering above the nest) while F2 nest was defended by only one kite. Discussion Observations in 1996 clearly showed that the same male “2AU” copulated with two females that used differ- 256 Short Communications VoL. 33, No. 3 ent nests, helped both females with nest building and helped one of them (FI) with incubation. Both females laid eggs and only one bred successfully. We cannot com- pletely exclude from the 1996 observations that copula- tions with F2 were extra-pair copulations by “2AU” and that F2 may have had an undetected partner. Observa- tions in 1997 were detailed enough to show that “2AU” was the only male seen with both females during nest building, incubation and brood rearing. The male “2AU” copulated seemingly equally with both females and helped both during nest building. During brood rearing, the male’s attention was directed exclusively to the FI female and F2 raised its single nestling alone. Polygyny had not been previously documented in Red Kites, but as individuals tend to be very similar in color- ation this behavior could be very easily overlooked unless kites are marked. This polygynous trio was similar to po- lygynous bondings observed in Northern Harriers ( Circus cyaneus) (Simmons et al. 1986) in that there was a hier- archy between females. In the case of the Red Kites, the second female to lay eggs received no attention from the male once its young hatched. Nearest-neighbor distance between Red Kite nests (x = 893.1 m, range = 690-2250, A= 21, Bustamante and Hiraldo 1990) was much shorter than elsewhere in Europe (4-5 km, Davies and Davis 1973, Valet 1975), probably indicating that there was abundant food for breeding Red Kites in Dohana. The Dohana area includes a highly productive marsh. The short distance between nesting territories in Donana may have been a proximate cause facilitating this polygynous mating. Four main hypotheses have been proposed to explain how polygyny may be advantageous. The “polygyny threshold” hypothesis (Verner and Wilson 1966, Orians 1969) suggests that a female should choose to mate po- lygynously in a high-quality territory when her fitness would be equal or greater than if she mated monoga- mously in a lower-quality territory. The “sexy son” hy- pothesis (Weatherhead and Robertson 1979) suggests that the female choice of polygyny could be based on the quality of the male rather than his territory. A third pos- sibility is that if suitable breeding sites are limited and monopolized by a few males, females may be forced to mate polygynously rather than forgo breeding (Orians 1961). A skewed sex ratio (shortage of males) could be equivalent to limited breeding sites. Finally, the “decep- tion” hypothesis (Alatalo et al. 1981) suggests that the male may be able to conceal its mated status to the fe- males holding two separate territories. It is difficult to believe that the male was able to de- ceive the second female and hide its mated status consid- ering the polygynous trio was stable for two years and that the two nests were very close. In 1996 and 1997, we sur- veyed 32 nesting territories inside the Donana Biological Reserve (10 000 ha) that had been occupied by Red Kites at least once since 1981. For the period 1981-97, the average number of nests in which Red Kites layed eggs was 16 (maximum 19 nests). Red Kites layed eggs in 13 nests in 1996 and in 14 nests in 1997 (F. Hiraldo unpubl. data) . Nesting sites did not seem to be limited in any of the two years. We do not know if there could be an un- balanced sex ratio or a lack of bachelor males. The ter- ritory of the male “2AU” could be of a higher quality than other territories in the area. The fact that F2 was able to raise a young in 1997 without the help of the male suggests that food was abundant in the proximity of the nest. The F2 nest was very close to a group of inhabited houses, the “Palacio de Donana,” that constitutes a pre- dictable source of food for the kites. Also, we cannot ex- clude the possibility that the male “2AU” was of higher quality than other males. Male “2AU” was 19 years old in 1997, the oldest known-age Red Kite in our marked population (F. Hiraldo unpubl. data). The longevity of “2AU” may suggest this bird carried “good genes” and he may have been a high-quality male. Resumen. — Se describen las observaciones de un trio po- liginico de milanos reales {Milvus milvus) en la Reserva Biologica de Donana, durante dos anos consecutivos: 1996 y 1997. Un macho marcado, de edad conocida, se apareo con dos hembras, posiblemente las mismas ambos anos, que utilizaron para poner nidos distintos. En 1996 la primera hembra crio dos polios y la segunda fracaso. En 1997 la primera hembra en realizar la puesta crio dos polios y la segunda uno. Se comprobo este ano que el macho cebaba y defendia exclusivamente el nido de la primera hembra y que la segunda fue capaz de criar con exito su polio sin la ayuda del macho. [Traduccion de Autores] Acknowledgments We thank M. de la Riva, A. Fernandez, N. Lawson, D. Lawson, and G. Vilchez for contributing with their ob- servations, and F. Hiraldo for access to unpublished data. J.J. Negro, J.A. Donazar, R. Rodriguez-Estrella, J. Bednarz and K. Bildstein provided helpful comments on the man- uscript. Funding was provided by project PB96-0855 from the Direccion General de Ensenanza Superior e Investi- gacion Cfica, Ministerio de Educacion y Cultura, Spain, and by the Agencia de Medio Ambiente, Junta de An- dalucia, Spain. Literature Cited Alatalo, R.V., A. Carlson, A. Lundberg and S. Ulf- STRAND. 1981. The conflict between male polygyny and female monogamy: the case of the Pied Flycatch- er Ficedula hypoleuca. Am. Nat. 117:738-753. Bustamante, J. 1990. Condicionantes ecologicos del pe- riodo de emancipacion en Falconiformes. Ph.D. dis- sertation, Univ. Autonoma de Madrid, Madrid, Spain. . 1993. Post-fledging dependence period and de- velopment of flight and hunting behaviour in the Red Kite Milvus milvus. Bird Study 40:181-188. AND F. Hiraldo. 1990. Adoptions of fledglings by Black and Red Kites. Anim. Behav. 39:804—806. September 1999 Short Communications 257 AND . 1993. The function of aggressive chases by breeding Black and Red Kites Milvus mi- grans and M. milvus during the post-fledging depen- dence period. Ibis 135:139-147. Cramp, S. and K.E.L. Simmons. 1980. Handbook of the birds of Europe, the Middle East and North Africa. Vol. II. Oxford Univ. Press, Oxford, U.K. Davies, P.W. and P.E. Davis. 1973. The ecology and con- servation of the Red Kite in Wales. Br. Birds 66:183- 224. Faaborg, J. and J.C. Bednarz. 1990. Galapagos and Har- ris’ Hawks: divergent causes of sociality in two raptors. Pages 357-383 in P. Stacey and W.D. Koenig [Eds.], Cooperative breeding in birds. Cambridge Univ. Press, Cambridge, NY U.S.A. Glutz von Blotzheim, U.N., K.M. Bauer and E. Bezzel. 1971. Handbuch der Vogel Mitteleuropas. Vol. 4. Fal- coniformis. Vol. 4. Academische Verlagsgesellschaft AULA-verlag, Wiesbaden, Germany. Heredia, B. and J.A. Donazar. 1990. High frequency of polyandrous trios in an endangered population of Lammergeiers {Gypaetus barbatus). Biol. Conserv. 53: 163-171. Hiraldo, F., JJ- Negro and J.A. Donazar. 1991. Aborted polygyny in the Lesser Kestrel Falco naumanni {Aves, Falconidae) . Ethology 89:253—257. Newton, I. 1979. Population ecology of raptors. T. &; A.D. Poyser, Berkhamsted, U.K. Orians, G.H. 1961. The ecology of blackbird (Agelaius) social system. Ecol. Monogr. 31:285-312. . 1969. On the evolution of mating systems in birds and mammals. Am. Nat. 103:589-603. Rogers, P.M. and K. Myers. 1980. Animal distribution, lanscape classification and wildlife management. Goto de Donana, Spain./. Appl Ecol. 17:545-565. Simmons, R.E., R.E. Smith and R.B. MacWhirter. 1986. Hierachies among Northern Harriers ( Circus cyaneus) harems and the cost of polygyny./. Anim. Ecol. 55:755- 771. Tella, J.L. 1993. Polyandrous trios in a population of Egyptian Vultures, /. Raptor Res. 27:119—120. . J J- Negro, M. Villarroel, U. Kuhnlein, F. Hi- raldo, F. Donazar and D. Bird. 1996. DNA finger- printing reveals polygyny in the Lesser Kestrel {Falco naumanni). Auk 113:262—265. Valet, G. 1975. La sedentarisation du Milan Royal Milvus milvus en Auxois. Alauda 43:263-269. Valverde, J.A. 1958. An ecological sketch of the Goto Donana. Br. 51:1-23. Verner, J. and M.F. Wilson. 1966. The influence of hab- itat on mating system of North American passerine birds. Ecology 47:143—147. Weatherhead, PJ- AND RJ. Robertson. 1979. Offspring quality and the polygyny threshold: “the sexy son hy- pothesis”. Am. Nat. 113:201-208. Received 5 December 1998; accepted 20 April 1999 / Raptor Res. 33(3):257-260 © 1999 The Raptor Research Foundation, Inc. Mites in Greater Spotted Eagle Nests Dariusz j. Gwiazdowicz, Tadeusz Mizera and Maciej Skorupski Akademia Rolnicza, Katedra Ochrony Lasu, and Zoologii Srodowiska Przyrodniczego, ul. Wojska Polskiego 7lc, PL 60-625 Poznan, Poland Key Words: Spotted Eagle, Aquila clanga; mites-, Acari; Me- sostigmata; Biebrza National Park, Poland. The Greater Spotted Eagle {Aquila clanga) is among the rarest predatory birds in the world, nesting from the eastern areas of Poland, through Belarus and Russia to the Pacific coast (Meyburg 1994). In Poland, it is legally protected and classified in the Red Data Book as Endan- gered (Kr61 1992). It nests in impenetrable boggy woods and, in Poland, only about 12 pairs remain in Biebrza National Park (Maciorowski et al. 1996). Greater Spotted Eagles build their nests only in birches {Betula spp.), black beech {Alnus glutinosa) and, less frequently, in oaks {Querqus spp.) and willows {Salix spp.). While the mite fauna in the nests of several predatory birds has been studied in Europe (Nordberg 1936, Philips 1981, Wis- niewski and Hirschmann 1985, 1990, Mizera 1990, Masan 1993), the acarofauna of Greater Spotted Eagle nests has never been described. In this paper, we identify mites of the order Mesostigmata which inhabit the nests of Greater Spotted Eagles in Poland. These mites are small (adult length 0.4-1. 6 mm), free-living arachnids which have phoretic or other relationships with birds. Methods Samples were obtained at two Greater Spotted Eagle nests in Biebrza National Park in northeastern Poland. 258 Short Communications VoL. 33, No. 3 The lining of the first nest was collected in 1994. The sample contained branches and grass and weighed about 300 g. At the time of collection, there was one 7-wk-old nestling in the nest. Subsequent observation confirmed that this hird later successfully fledged. The nest was built m the crown of an aspen tree {Populus tremuloides) in 1991 and was reused each year. It was built mainly from branches and mistletoe. The sample from the second nest was collected in 1995. It had a similar composition and was approximately the same weight. This nest was in a birch tree. Mites were extracted from collected material by funnel- heat dessication and preserved in 70% alcohol. For mac- eration and bleaching, mites were placed in a drop of lacto-phenol on a slide and a coverslip was applied. Per- manent preparations were made in polyvinyl alcohol (Evans 1992). Results and Discussion Only mites of the order Mesostigmata were identified to genus and species level. Two subclasses of mites were rep- resented, the Anactinotrichida and Actinotrichida. In the subclass Anactinotrichida, we identified mites in seven families of the suborder Gamasina. Nest II contained one deutonymph of Parasitus consanguineus which is a rare species in the Parasitadae found in arable and meadow soils, in compost, stable manure, and in decomposing organic material (Karg 1993). This is a new species to the Polish fauna. We also found a Macrochelid, Macrocheles ancyleus, in both Nest I (104 females, 28 males, 21 deu- tonymphs, 9 protonymphs) and Nest II (242 females, 28 males, 15 deutonymphs, 5 protonymphs). It has been found previously in the decaying wood of aspen trees (Krauss 1970) hut it is also a new species to the Polish fauna. Macrochelid mites are not uncommon in birds nests and are known from other raptor nests (Amhros et al. 1992, Philips et al. 1983, Zeman and Jurik 1981). We also identified three species of mites in the family Ascidae: Arctoseius cetratus, Proctolaelaps pini, and P. pyg- maeus. We found one female and one protonymph of Arctoseius cetratus in Nest I and two females in Nest II. This species has previously been found in soils of arable fields, in compost, in deciduous and mixed forests, and humus between plant roots (Karg 1993). In Nest II, we also found 37 females, seven males, nine deutonymphs, and five protonymphs of P. pini. It is typically found in pine stumps and on bark beetles {Hylastes spp.) (Karg 1993). We also found five females, one male, and one deutonymph of P. pygmaeus in Nest II. It has been previ- ously found in soil, moss, decaying plants, and nests of small mammals (Bregetova 1977a). Three species of mites were in the family Laelapidae. One female, one male, and one deutonymph of Andro- laelaps casalis occurred in Nest II. It is a species typically found in the soils of meadows, humus between roots, hay and straw, and in nests of rodents and birds (Karg 1993), such as raptors (Philips 1981, Zeman and Jurik 1981). Also in Nest II, we found two females of Hypoaspis ( Cos- molaelaps) vacua which is typically encountered in decid- uous and conifer forests, in humus and moss, in decaying wood, and in nests of ants of the genus Lasius (Bregetova 1977b, Karg 1993). Lastly, we identified two females of Hypoaspis (Pneumolaelaps) lubrica in Nest I. It is typically found on rotten plants, in rodent nests, and on rodents themselves, as well as in the nests of Sand Martins {Ri- paria riparia) (Karg 1993). Three species in the family Digamasellidae occurred in both nests. A total of 40 females, four males, and 50 deu- tonymphs of Dendrolaelaps {Punctodendrolaelaps) fallax were in Nest I but only five females and one deutonymph were found in Nest II. This is a species that is commonly found in compost and stable manure (Hirschmann and Wisniewski 1982, Karg 1993). Twelve female, one male, three deutonymph, and one larva of Dendrolaelaps {Punc- todendrolaelaps) wengrisae were found in Nest II. It has been previously found in ant {Formica fusca) nests (Hirschmann and Wisniewski 1982). Only one female of Dendrolaelaps {Apophyseodendrolaelaps) zwoelferi was in Nest II. It has been previously found in nests of the Formica rufa, under pine bark, in the pathways of Dryocoetes auto- graphus, and in Sand Martin nests (Scerbak 1980). A female and a deutonymph of two additional species, Paragarmania sp. (Phytoseiidae) and Halolaelaps sp. (Hal- olaelapidae) were also found in Nest II. We found mites representing five families in the sub- order Uropodina. One species each in the families Trach- ytidae and Polyaspidae occurred in Nest I. We found one female Trachytes aegrota (Trachytidae). It is a species found in moss, soil, grass, often in litter of deciduous and conifer forests, in rotting wood, in the paths of bark bee- tles, in bird nests, nests of the mole Talpa europaea, and ant nests (Karg 1989, Wisniewski and Hirschmann 1993). A total of 44 females, 49 males, 51 deutonymphs, 38 pro- tonymphs, and one larva of Uroseius {Apionoseius) infirmus (Polyaspidae) were also identified from this nest. This species is typically found in moldy wood, in rotting parts of plants, in the litter of deciduous and coniferous for- ests, and in bird nests (Kadite and Petrova 1977, Karg 1989, Masan 1993, Wisniewski and Hirschmann 1993). Three species in the family Trematuridae were in both nests but they were not common. A female and male of Nenteria flcrralis, a species associated with dunghills and greenhouse soils (Wisniewski and Hirschmann 1993), were found in Nest II. This is a new species to the Polish fauna. Four females, three males, two deutonymphs, three protonymphs, and one larva of Nenteria pandioni were identified in Nest I and 74 females, 26 males, 229 deutonymphs, 48 protonymphs, and nine larvae were found in Nest II. This species has only been found in Osprey {Pandion haliaetus) and White-tailed Sea Eagle {Haliaeetus albidlla) nests (Wisniewski and Hirschmann 1985, 1990) (Fig. 1). Two females of Trichouropoda ovalis, a species commonly found in hay, moss, mushrooms, straw, ant nests, the paths of Scolytidae, on Cerambycidae and Scarabaeidae, in decaying wood, in hollow tree trunks, September 1999 Short Communications 259 Figure 1. A dorsal view of Nenteria pandioni (Wisniewski and Hirschmann 1985). and in bird and mole nests (Wisniewski and Hirschmann 1993), was also found in Nest II. Two species of mites in the family Urodinychidae, Di- nychus perforatus and Uroobovella obovata, were identified. One female of Dinychus perforatus found in Nest I and Uroobovella obovata found in both Nest I (two females, two males, and one deutonymph) and Nest II (seven fe- males, three males, and two deutonymphs) . Both species are usually encountered in rotting litter of deciduous for- ests, meadow soils, moss, decaying plant material, mole nests, and phoretically on birds (Karg 1989, Wisniewski and Hirschmann 1993). We found one representative of the family Uropodi- dae, Uropoda (Phaulodinychus) hamulifera. Ten deuto- nymphs and two protonymphs were found in both nests. This species is found in ant nests and stable manure (Karg 1989, Wisniewski and Hirschmann 1993). In the material collected, 1267 mites in the order Me- sostigmata and 646 in the subclass Actinotrichida were found. Mites in the order Mesostigmata represented 21 species and 12 families. Eleven and 16 species of mites were found in Nests I and II, respectively. Six species were common to both nests. The most frequently found spe- cies in Nest I were U. infirmus (183 individuals), M. an- cyleus (162), D. fallax (94), and N. pandioni (13). In Nest II, the most frequent species were N. pandioni (386), M. ancyleus (290), P. pini (58), and D. wengrisae (17). Species new to the Polish fauna were M. ancyleus, N. floralis, and P. consanguineus. It is difficult to evaluate the relationship between free- living mites and predatory birds; however, some species of mites occur, almost exclusively, in nests of these birds. Previously N. pandioni was found only in the nests of Os- preys and White-tailed Sea Eagles (Wisniewski and Hirschmann 1985). The large numbers of all stages of development of N. pandioni that we found in Greater Spotted Eagle nests appear to confirm that eagle nests provide a specific microhabitat for this mite. Resumen. — Los nidos de aves rapaces proveen un micro- habitat para pequenos invertebrados, particularmente para piojos. Estudiamos la acarofauna del orden Mesotig- mata que habita en nidos de Aquila clanga en el parque Nacional de Biebrza en el noreste de Polonia. Veintiuna especies de piojos fueron identificados. Una especie: Nen- teria pandioni, aparentemente es dependiente del habitat de anidacion de Haliaeetus albicilla y de Pandion haliaetus Las otras especies fueron encontradas en madera podri- da, basura, compost y nidos de aves. [Traduccion de Cesar Marquez] Literature Cited Ambros, M.,J. Kristofikand Z. Sustek. 1992. The mites {Acari, Mesostigmata) in the birds’ nests in Slovakia. Biologia (Bratislava) 47(5):369-381. Bregetova, N.G. 1977a. Semejstvo Aceosejidae. Pages 169- 476 in M.S. Giljarov and N.G. Bregetova [Eds.], Opre- delitiel obitajuscih v pocvie klescej Mesostigmata. Nau- ka, Leningrad, Russia. . 1977b. Semejstvo Laelaptidae. Pages 483-554 in M.S. Giljarov and N.G. Bregetova [Eds.], Opredehtiel obitajuscih v pocvie klescej Mesostigmata. Nauka, Len- ingrad, Russia. 260 Short Communications VoL. 33, No. 3 Evans, G.O. 1992. Principles of acarology. C.A.B. Inter- national, Wallingford, U.K. Hirschmann, W. and J. Wisniewski. 1982. Weltweite Re- vision der Gattungen Dendrolaelaps Halbert 1915 und Longoseius Chant 1961 {Parasitiformes) . Acarologie (Ntirnberg) 29, 1:1-190, 11:1-48. Kadite, B.A. AND A.D. Petrova. 1977. Kohorta Uropodina, family Uropodidae. Pages 632-690 in M.S. Giljarov and N.G. Bregetova [Eds.], Opredelitiel obitajuscih v po- cvie klescej Mesostigmata. Nauka, Leningrad, Russia. Karg, W. 1989. Acari {Acarina), Milben Unterordnung Parasitiformes (Anactinochaeta) Uropodina Kramer, Schildkrotenmilben. Gustav Fischer Verlag, Jena, Ger- many. . 1993. Acari {Acarina), Milben Parasitiformes {An- actinochaeta), Cohors GamasinaUeach. Raubenmilben. Gustav Fischer Verlag, Jena, Germany. Krauss, W. 1970. Die europaischen Arten der Gattungen Macrocheles UaXxeiWc 1829 und Geholaspis^crXese 1918. Acarologie (Niirnberg) 14:2-43. Krol, W. 1992. Orlik grubodzioby clanga (Greater Spotted Eagle). Pages 129-131 in Z. Glowacinski [Ed.], Polish Red Data Book of animals. State Pub- lishing House of Agricultural and Forestry, Warszawa, Poland. Maciorowski, G., B.-U. Meyburg, J. Matthes and T. Miz- ERA. 1996. Breeding biology of the Greater Spotted Eagle {Aquila clanga) in Poland. Pages 35-36 in Sec- ond International Conference on Raptors, Raptor Re- search Foundation, Urbino, Italy. Masan, P. 1993. Mites {Acarina) associated with species of Trox {Coleoptera: Scarabaeidae) . Eur. J. Entomol. 90: 359-364. Mevburg, B.-U. 1994. Greater Spotted Eagle Aquila clan- ga. Page 193 in J. Hoyo, A. Elliott and J. Sargatal [Eds.], Handbook of the birds of the world. Vol. 2. Lynx Edicions, Barcelona, Spain. Mizera, T. 1990. The White-tailed Sea Eagle Haliaeetus alhidlla in Poland: present status and prospects. Pages 17-23 mj. Viksne and I. Vilks [Eds.], Baltic birds 5, ecology, migration and protection of Baltic birds. Pro- ceedings of the Fifth Conference on the Study and Conservation of Migratory Birds of the Baltic Basin, Ziname, Riga, Russia. Nordberg, S. 1936. Biologisch-okologische untersuchun- gen uber die Vogelnidicolen. Acta Zool. Fenn. 21:1- 168. Philips, J.R. 1981. Mites {Acarina) from nests of Norwe- gian birds of prey. Fauna Norv. Ser. B. 28:44—47. , M. Root and P. DeSimone. 1983. Arthropods from a Saw-whet Owl {Aegolius acadicus) nest in Con- necticut. Entomol. News 2:60-64. SCERBAK, G. I. 1980. Klescej semejstva Rhodacaridae Pa- learktyki. Naukova Dumka, Kijev, Ukraine. Wisniewski, J. and W. Hirschmann. 1985. Studium einer neuen Nenteria-Art der stammeri-Gruppe aus einem Fi- schadlernest in Polen {Trichouropodini, Uropodinae). Acarologie (Nurnberg) 32:161-163. AND . 1990. Erganzungsbeschreibung von Nenteria pandioni Wisniewski et Hirschmann, 1985 {Ac- arina, Uropodina) aus Polen. Ann. Zool. 43:259-269. AND . 1993. Katalog der Ganggattungen, Untergattungen, Gruppen und Arten der Uropodiden der Erde. Acarologie (Niirnberg 40:1-220. Zeman, P. and M. Jurik. 1981. A contribution to the knowledge of fauna and ecology of Gamasoid mites in cavity nests of birds in Czechoslovakia. Folia paras- itol. (Praha) 28:265-271. Received 15 April 1998; accepted 20 April 1999 J Raptor Res. 33(3):260-264 © 1999 The Raptor Research Foundation, Inc. Food Habits of the Rufous-legged Owl {Strix rufipes) in the Mediterranean SCLEROPHYLLOUS FOREST OF CENTRAL CHILE IvAn Diaz Laboratorio de Sistemdtica y Ecologia Vegetal, Facultad de Ciendas, Universidad de Chile, Casilla 653, Santiago, Chile Key Words: Rufous-legged Owl; Strix rufipes; food habits; Mediterranean sclerophyllous forest, central Chile. The Rufous-legged Owl {Strix rufipes) inhabits old- growth temperate rainforest of southern South America (Housse 1945, Goodall et al. 1946, Johnson 1967, Vuil- leumier 1985, Martinez and Jaksic 1996, 1997). Infor- mation on the species is limited, and only plumage, hab- itat and distribution have been described (Housse 1945, Goodall et al. 1946, Johnson 1967, Vuilleumier 1985, Ar- aya and Millie 1989). Martinez (1993a, 1996) and Mar- tinez and Jaksic (1996, 1997) have recently reported on population density, specific habitat requirements and diet of the Rufous-legged Owl in temperate rainforests of September 1999 Short Communications 261 southern Chile. They pointed out similarities in habitat requirements of this species and the Northern Spotted Owl {Strix occidentalis) of North America. In Chile, the range of the Rufous-legged Owl extends from Santiago to Magallanes, (33°S-55°S) (Goodall et al. 1946, Araya and Millie 1989). In the northern limit of distribution, it inhabits hydrophyllous and sclerophyllous forest of the Mediterranean zone. Relative to southern Chile, abundance in central Chile is lower (Housse 1945, Jaksic and Jimenez 1986), and numbers appear to be de- clining as a result of habitat loss (Jaksic and Jimenez 1986). From a conservation viewpoint, the Rufous-legged Owl has been catalogued as inadequately studied and in need of further research (Glade 1993). Historical records doc- umenting the presence of Rufous-legged Owls in moun- tainous old-growth forest of central Chile (Housse 1945) date from the past century (specimens present at the Na- tional Museum of Natural History from Central Chile are MNHM 153 Aculeo 1865, MNHM 150 San Francisco del Monte 1869, MNHM 4338 Quebrada de Ramon 1947). Since the arrival of Spanish colonials, forests of the Cen- tral Valley and Andean range in Chile have largely been destroyed (Elizalde 1970, Mooney et al. 1972, Hoffmann 1998). During this century, forests near Santiago have virtually disappeared (Lara 1998), having been cut for firewood in Santiago and nearby towns. Trees older than 250 yr are scarce and may only be found along deep ravines in remote mountainous areas. This study is the first report of the food habits of Ru- fous-legged Owl in the Mediterranean sclerophyllous for- est near the northern limit of its distribution in Chile. Study Area and Methods The study was conducted in Rio Clarillo National Re- serve (33°45'S, 70°26'W), located 45 km southeast of San- tiago in the foothills of the Andean range. The reserve covers 10 185 ha of mountainous landscapes, with slopes of approximately S45° and 870-3050 m in altitude (ICSA 1983). The climate at 870 m is Mediterranean with mean annual rainfall of 648 mm and a long dry season from September to May (CONAF 1996). The reserve supports sclerophyllous forest <1500 m that are dominated by Cryptocarya alba (Lauraceae), Per- sea lingue (Lauraceae), Drimys udnteri (Winteraceae), Lith- raea caustica ( Anacardiaceae) , Quillaja saponaria (Rosa- ceae), and Acacia caven (Mimosaceae) (CONAF 1996, ICSA 1983). Between 1500—2000 m, Kageneckia angustifol- ia (Rosaceae) and Austrocedrus chilensis (Cupressaceae) are dominant. Above 2000 m, shrubs such as Chuquiraga sp. and Mulinum sp. grow to a height of 50 cm (ICSA 1983). According to local inhabitants, the location of the reserve was very remote until the middle of the century. However, around the turn of the century, the forest be- gan to be exploited for firewood and charcoal, and open areas were used for cattle grazing during the summer (CONAF 1996, ICSA 1983). Exploitation was intensified in 1972, when indiscriminate tree removal began, and many parts of the reserve were burned and cleared (CONAF 1996). In 1978, hunting was prohibited and, in 1982, the area was designated a national reserve (CONAF 1996, ICSA 1983). A total of eight surveys were conducted on a periodic basis (one survey each two or three months) from Feb- ruary 1994— October 1995, in a forest patch less than 1 ha in size located at 1100 m altitude that was occupied by a Rufous-legged Owl pair. In all, 210 pellets were col- lected and their contents were identified to species level through comparisons with the mammal skull collection in the National Museum of Natural History, Santiago, Chile. A reference collection was also used to identify insects found in pellets. Data were pooled by season and the results expressed .separately for each season as a per- centage of the total prey mass consumed. The following equation was used to calculate the percentage of the total prey mass consumed represented by each species in the diet (Martinez 1993a), assuming that the entire body mass of each individual prey species caught was con- sumed by the owl; B% = (lOOSpiNi)/ S Sp,N, where Sp^ is the mass of the ith species, A; is the number of individuals of ith species consumed and B% is the per- centage of the total prey mass consumed for each season contributed by the ith species. The mass of each prey species was obtained by capturing and weighing individ- uals of different taxa present in the Rufous-legged Owl diet. For mammals, body mass was obtained from Jaksic (1986). Results and Discussion The diet of Rufous-legged Owl consisted of 10 mam- malian species, one lizard and four invertebrate families (Table 1). In spite of the large number of invertebrates consumed, they composed only 2.1% of the total prey biomass. Mammals made up 98% of prey biomass in the Rufous-legged Owl diet (62% of the total of number of prey items consumed) . The species most frequently con- sumed were Abrocoma bennetti and Octodon lunatus, com- prising 16.5% and 15.2% of the total prey biomass, re- spectively. In addition, 29% of the prey biomass (13% of individual prey items identified) was attributed to mam- mals identified from remains retrieved from pellets which did not contain skulls or teeth. These pellets con- tained large femurs or humeri (x = 32.2 ± 0.7 and 27.6 ± 0.3 mm, respectively) identified as adult A. bennetti and O. lunatus, the only large rodents in the area (CONAF 1996, Diaz and Sarmiento 1999). These species consti- tuted 35% of the total number of prey items identified, 51% of the mammals captured and the 61% of the total prey biomass consumed. Both are primarily terrestrial species which live in caves among rocks covered by shrubs (Mann 1978), and occasionally climb trees (C. Ve- loso pers. comm.). Lizards in central Chile are diurnal (Donoso-Barros 1966), and were probably consumed only occasionally. In contrast to data presented by Martinez (1993a) and Martinez and Jaksic (1996, 1997) on Rufous-legged Owls in forests of southern Chile, prey species in this study 262 Short Communications VoL. 33, No. 3 Table 1. Food habits of the Rufous-legged Owl (Strix rufipes) in sclerophyllous forest in Rio Clarillo National Reserve, central Chile. Data from 1994 and 1995 are pooled by season. B% is a percent of total prey mass consumed and {N) IS prey number by season. Prey Mass ( g ) Summer Autumn Winter Spring TOTAL 71 % {N) B% (AO 5 % (AO B% (AO fi% (AO Mammals Abrocoma bennetti 219 =* 30.0 ( 15 ) 20.0 ( 9 ) 0 ( 0 ) 14.0 ( 5 ) 19.0 ( 29 ) Abrocoma bennetti* legi * 6.2 ( 4 ) 1.8 ( 1 ) 27.0 ( 8 ) 4.2 ( 2 ) 7.5 ( 15 ) Octodon lunatus ISC ’* 9.9 ( 6 ) 26.0 ( 14 ) 7.2 ( 2 ) 11.0 ( 5 ) 14.0 ( 27 ) Octodon lunatus* 14 Q 1 * 0 ( 0 ) 0 ( 0 ) 8.4 ( 3 ) 0 ( 0 ) 1.2 ( 3 ) Oryzomys longicaudatus 45 =* 0.4 ( 1 ) 1.9 ( 4 ) 1.8 ( 2 ) 5.0 ( 9 ) 2.1 ( 16 ) Abrothrix longipilis 76 =* 2.8 ( 4 ) 3.9 ( 5 ) 0 ( 0 ) 1.9 ( 2 ) 2.5 ( 11 ) Abrothrix olivaceus 40 =* 0.4 ( 1 ) 0.4 ( 1 ) 3.2 ( 4 ) 2.0 ( 4 ) 1.2 ( 10 ) Phyllotis darwini 66 =* 0.6 ( 1 ) 4.1 ( 6 ) 7.9 ( 6 ) 4.1 ( 5 ) 3.5 ( 18 ) Cricetidae unident. 40 ^ 0.4 ( 1 ) 0.4 ( 1 ) 2.4 ( 3 ) 0.5 ( 1 ) 0.7 ( 6 ) Rattus norvegicus 220 4.0 ( 2 ) 0 ( 0 ) 0 ( 0 ) 0 ( 0 ) 1.3 ( 2 ) Rattus rattus 158 =* 2.9 ( 2 ) 3.3 ( 2 ) 3.1 ( 1 ) 0 ( 0 ) 2.3 ( 5 ) Thylamis elegans 40 =* 2.2 ( 6 ) 2.9 (V) 3.2 ( 4 ) 3.0 ( 6 ) 2.7 ( 23 ) Oryctolagus cuniculus* 408 ** 7.5 ( 2 ) 0 ( 0 ) 16.0 ( 2 ) 0 ( 0 ) 4.9 ( 4 ) Unident, large mammals 200 ** 24.0 ( 13 ) 27.0 ( 13 ) 16.0 ( 4 ) 45.0 ( 18 ) 29.0 ( 48 ) Unident, small mammals 58 ** 6.4 ( 12 ) 5.4 ( 9 ) 2.3 ( 2 ) 7.9 ( 11 ) 5.9 ( 34 ) Subtotal mammals 97.8 ( 70 ) 97.6 ( 72 ) 98.5 ( 41 ) 97.8 ( 68 ) 97.8 ( 251 ) Reptilians Squamata unident. 10 ** 0 ( 0 ) 0 ( 0 ) 0.2 ( 1 ) 0 ( 0 ) 0.03 ( 1 ) Subtotal reptilians 0 ( 0 ) 0 ( 0 ) 0.2 ( 1 ) 0 ( 0 ) 0.03 ( 1 ) Insects Cerambicidae Acanthinodera cummingii 11 ** 1.1 ( 11 ) 0.6 ( 5 ) 0.4 ( 2 ) 0.7 ( 5 ) 0.8 ( 23 ) Scarabeidae 0 . 3 ** 0 ( 0 ) 0 ( 0 ) 0 ( 0 ) 0 ( 9 ) 0 ( 9 ) Gryllacridiidae 3 . 2 ** 0.4 ( 15 ) 1.7 ( 50 ) 0 ( 0 ) 0.8 ( 20 ) 0.8 ( 85 ) Insecta unident. 4 ** 0.6 ( 15 ) 0.1 ( 3 ) 0.9 ( 11 ) 0.7 ( 14 ) 0.5 ( 43 ) Subtotal insects 2.1 ( 41 ) 2.4 ( 58 ) 1.3 ( 13 ) 2.2 ( 48 ) 2.1 ( 160 ) Scorpions 0 . 8 ** 0.1 ( 11 ) 0 ( 3 ) 0 ( 1 ) 0 ( 1 ) 0 ( 16 ) Total prey items 122 133 56 117 428 Total prey mass 10 882 9643 5025 8077 33 627 Total pellets 59 55 29 67 210 ‘‘Prey biomass obtained fromjaksic (1986). Prey biomass obtained from I. Diaz (unpubl. data). * Young specimens. were larger and primarily terrestrial. In southern Chile, 72.4% of the total prey mass consumed were scansorial mammals weighing 26-42 g (Martinez 1993a) while, in this study, adults of the dominant prey species, A. bennetti and O. lunatus, weighed between 180-219 g, and young individuals weighed over 100 g. In central Chile, Abrothrix longipilis, A. olivaceus, and Oryzomys lorigicaudatus made up a relatively small proportion of species consumed by Ru- fous-legged Owls, in contrast to southern Chile where they were consumed in large quantities (Martinez 1993a, Martinez and Jaksic 1996, 1997). The sizes of prey species identified in this study are larger than the optimal prey size predicted by Bozinovic and Medel (1988) for owls of a similar size. Therefore, it is probable that individuals of Rufous-legged Owl did not select prey species according to optimal size, but accord- ing to prey species abundance. Large prey may have been eaten in pieces, instead of being swallowed whole. I have observed this behavior in Great Horned Owls {Bubo vir- ginianus) when they caught and ate European rabbits ( Oryctolagus cuniculus) . Less palatable body parts or heads of large mammals which may have been too large to swal- September 1999 Short Communications 263 low may have been avoided. This could explain the high percentage of pellets found which did not contain teeth or skulls (Table 1). The forest patch in Rio Clarillo provided an open un- derstory and forest floor which may have enhanced Ru- fous-legged Owls’ ability to see and capture terrestrial mammals. The opposite is true in the southern Chilean rainforest, where dense understory provides cover for ter- restrial mammals, and owls forage mainly on arboreal and scansorial mammals (Martinez 1993a, 1993b, Marti- nez andjaksic 1997). In the Rio Clarillo National Reserve forest patch, some Persea lingue and Cryptocarya alba trees are >200 years old and reach 13-15 m height. They pro- vide 70% canopy cover, similar to that reported by Mar- tinez andjaksic (1996) in southern Chile. The Rufous- legged Owl pair studied in Rio Clarillo was frequently observed outside forest patches during night hours, giv- ing territorial vocalizations in open shrublands with a canopy <6 m in height. However, during the day they were always observed inside the forest patch and pellets were found only in the forest. It is my opinion that Ru- fous-legged Owls use forest patches for nesting and for protection and shelter during daylight hours, and forage in forest and shrublands during the night. Resumen. — Este es el primer analisis cuantitativo de la dieta del Concon (Strix rufipes) en un bosque esclerofilo, cerca del limite norte de la distribucion de esta especie en Chile. Se analizo el contenido de 210 egagropilas re- colectadas en un remanente de bosque esclerofilo entre 1994 y 1995. Los resultados indican que el Concon con- sumio mayoritariamente los roedores Abrocoma bennetti y Octodon lunatus, ademas de otras ocho especies de mami- feros, un lagarto y cuatro familias de artropodos. El ta- mano de las presas consumidas es de tres a cuatro veces mayor que el tamano de las presas consumidas por esta misma especie en el sur de Chile. Las presas consumidas son fundamentalmente terricolas, a diferencia de las es- pecies arboricolas y escansoriales consumidas en el sur. Probablemente el Concon captura estas grandes presas y las consume por trozos. Se han detectado ejemplares fuera de los remanentes de bosque, en matorrales, por lo cual el Concon podria usar los bosques para descansar durante el dia y anidar, y por la noche capturaria sus presas en bosques y matorrales vecinos. [Traduccion del Autor] Acknowledgments I appreciate the company and assistance in the field provided by Mauricio Soto, Andres Sepulveda, Eduardo Soto, the late Luis Cavieres, Christian Laytte, Hector Mel- lado and Claudia Ron. Our gratitude is also extended to Rogelio Moreira, Luis Ulloa, Carlos Pena, Carlos Sar- miento and Jorge Naranjo of the Chilean Forest Service for authorizing and providing facilities for conducting the field work. David Martinez provided important com- ments about the field work and Javier Simonetti give me assistance at the beginning of the project. I thank Jaime Solervicens and Ignacio Mellado for their assistance m weighing invertebrates and J.C. Torres-Mura for access to the mammal skull collection of the Museo Nacional de Historia Natural, Chile. R. Nespolo, J.C. Torres-Mura, C. Marti, D.W. Holt, M. Rosenmann, and one anonymous referee provided helpful and constructive comments on previous versions of the manuscript. I am grateful to Kate Matthews and Traci Darnell for providing assistance in the translation of this manuscript. This work was funded by Fondo Universitario de Desarrollo Estudiantil, Univer- sity of Chile. Literature Cited Araya, B. and G. Millie. 1989. Guia de campo de las aves de Chile. Editorial Universitaria, Santiago, Chile. Bozinovic, F. and R. Medel. 1988. Body size, energetic and foraging mode of raptors in central Chile: an in- ference. Oecologia (Berlin) 75:456-458. CONAF (CoRPORACiON Nacional Forestal). 1996. Plan de manejo Reserva Nacional Rio Clarillo. Documento de trabajo n” 247, Unidad de gestion Patrimonio Sil- vestre, Corporacion Nacional Forestal, Santiago, Chile. Diaz, I. and C, Sarmiento. 1999. Vertebrados terrestres de la Reserva Nacional Rio Clarillo: representatividad y conservacion. Documento tecnico, Departamento de Patrimonio Silvestre, Corporacion Nacional Fores- tal, Santiago, Chile. Donoso-Barros, R. 1966. Reptiles de Chile. Ediciones de la Universidad de Chile, Santiago, Chile. Elizalde, P. 1970. La sobrevivencia de Chile. Ministerio de Agricultura, El Escudo Impresores, Santiago, Chile. Glade, A. 1993. Libro rojo de los vertebrados terrestres de Chile. Corporacion Nacional Forestal, Santiago, Chile. Goodall, J.D., A.W. Johnson and R.A. Philippi. 1946 Las aves de Chile, su conocimiento y sus costumbres. Platt Establecimientos Graficos S.A., Buenos Aires, Ar- gentina. Hoffmann, A. 1998. La tragedia del bosque chileno Ocho Libros Editores, Santiago, Chile. Housse, R. 1945. Las aves de Chile en su clasificacion moderna: su vida i sus costumbres. Ediciones Univer- sidad de Chile, Santiago, Chile. ICSA (Ingenieros Consultores S.A.). 1983. Plan de ma- nejo de la Reserva Nacional Rio Clarillo. Corporacion Nacional Forestal, Santiago, Chile. Jaksic, F.M. 1986. Predation upon small mammals in shrublands and grasslands of southern South Ameri- ca: ecological correlates and presumable consequenc- es. Rev. Chil. Hist. Nat. 59:209-221. and J. Jimenez. 1986. The conservation status of raptors in Chile. Birds of Prey Bull. 3:95-104. Johnson, A.W. 1967. The birds of Chile and adjacent re- gions of Argentina, Bolivia and Peru. Vol. II. Platt Es- tablecimiento Graficos, Buenos Aires, Argentina. Lara, A. 1998. El catastro de lavegetacion nativa: lafuer- 264 Short Communications VoL. 33, No. 3 za de los resultados. Pages 118—119 m A. Hoffmann [Ed.], La tragedia del bosque chileno. Ocho Libros Editores, Santiago, Chile. Mann, G. 1978. Los pequehos mamiferos de Chile. Guya- na Zool. 40, 342 pp. Martinez, D. 1993a. Food-habits of the Rufous-legged Owl {Strix rufipes) in temperate rainforest of Southern Chile./. Raptor Res. 27:214—216. . 1993b. Depredacion selecdva de roedores por zorros y buhos en la pluviselva valdiviana. Resumenes del III Congreso Internacional de Gestion en Recur- sos Naturales, Pucon, Chile. . 1996. Ecologia trofica del Concon {Strix rufipes, King 1828) y su respuesta a manipulaciones de frac- cion de sus presas en el bosque de “San Martin,” Valdivia, Chile. Tesis de Magister, Facultad de Cien- cias, Universidad Austral de Chile, Valdivia, Chile. AND F.M. jAKSic. 1996. Habitat, abundance and diet of the Rufous Legged Owls {Strix rufipes) in tem- perate forests remnants of southern Chile. Ecoscience 3:259-263. AND . 1997. Selective predation on scanso- rial and arboreal mammals by Rufous-Legged Owl {Strix rufipes) in southern Chilean rainforest./. Raptor Res. 31:370-375. Mooney, H.A., E.L. Dunn, F. Shrospshire and L. Song, Jr. 1972. Land use history of California and Chile as related to the structure of the sclerophyll scrub veg- etations. Madrono 21:305-319. Vuilleumier, F. 1985. Forest birds of Patagonia: ecologi- cal geography, speciation, endemism, and faunal his- tory. Pages 255—304 in P.A. Buckley, M.S. Foster, E.S. MortOn, R.S. Ridgely and EG. Buckley [Eds.], Neo- tropical ornithology. Ornithol. Monogr. 36. Received 7 October 1998; accepted 10 April 1999 /. Raptor Res. 33(3):264-267 © 1999 The Raptor Research Foundation, Inc. Food Habits of the Cinereous Harrier ( Circus cinereus) in the Araucania, Southern Chile Ricardo A. Figueroa R. and Ema S. Corales S. Laboratorio de Ecologia, Departamento de Ciencias Bdsicas, Universidad de los Lagos, Casilla 933, Osorno, Chile Key Words: Cinereous Harrier, Circus cinereus; food; Chile. The Cinereous Harrier {Circus cinereus) is widespread in South America, ranging from northern Colombia to Tierra del Fuego in open areas such as marshes, grass- lands, shrublands, and steppes (Jimenez and Jaksic 1988, and references therein). In Chile, it is a resident along its entire distributional range from Copiapo (30°S) to Ti- erra del Fuego (57°S) (Hellmayr 1932, Johnson 1965). Its status varies from “rare” in northern Chile, to “fre- quent” in central and southern Chile, to “abundant” in southernmost Chile (Jaksic and Jimenez 1986). Although historical forest cutting and establishment of croplands has favorably affected the Cinereous Harrier in southern Chile, current increases of exotic pine and eucalyptus plantations and marsh draining for urban and agricul- tural development appear to negatively impact harriers by decreasing their hunting habitat, prey availability and nest sites (Jaksic and Jimenez 1986). Little is known about the biology of this South Amer- ican harrier, except for studies conducted by Jimenez and Jaksic (1988) in Chile and by Saggese and De Lucca (1995) in Argentina. Data on its diet in the Magellanic steppe were reported by Jimenez and Jaksic (1988) and reanalyzed by Iriarte et al. (1990). Here, we report the first quantitative data on food habits of the Cinereous Harrier in an agricultural landscape of the Araucania re- gion (ca. 38°S) in southern Chile. Study Area and Methods We studied two pairs of Cinereous Harriers in a 200- ha area at Tricauco Farm, located approximately 6 km south of Traiguen city (38°14'S, 72°38'W) in the Arau- cania Region of Chile. The landscape comprised crop- lands of wheat and corn, grasslands, marshlands, exotic tree plantations of Pinus spp. and Eucalyptus spp. and remnants of the original deciduous Nothofagus forest. The climate is moist-temperate with a Mediterranean influ- ence (di Castri and Hajek 1976) and mean annual rain- fall and temperature are 1400 mm and 12°C, respectively. Although originally the Araucania region extended from Maule Province to Seno Reloncavl, the current Admin- istrative Araucania Region extends from the Renaico Riv- er (37°30'S) to Calafquen Lake (38°47'S). During August 1996 (austral winter), we collected 68 pellets in abandoned pastures and marshes where harri- ers hunted and perched. Pellets were collected both from open ground in pasture-marshes and under trees, fences and stumps used as perches by harriers. Avian prey were identified mainly on the basis of feathers, using two com- September 1999 Short Communications 265 Table 1. Food habits of Cinereous Harriers ( Circus cinereus) in an agricultural landscape of the Araucania, southern Chile. Prey Mass (g) Frequency (%) Biomass (%) Mammals 21.4 26.3 Rodentia Abrothrix olivaceus 23 4.4 0.9 Oligoryzomys longicaudatus 26 2.7 0.6 Mus musculus 21 0.9 0.2 Unidentified rodents 23 10.7 2.2 Lagomorpha Oryctolagus cuniculu^-^ 800 0.9 6.4 Lepus europaeu^'^ 1000 1.8 16.0 Birds 67.0 73.3 Tinamiformes Nothoprocta perdicaria 160 2.7 3.8 Ciconiiformes Nycticorax nycticorao^ 600 0.9 4.8 Anseriformes Anas flavirostris 400 0.9 3.2 Anas georgica VOO 1.8 11.2 Falconiformes Milvago chimang(P 100 0.9 0.8 Galliformes Callipepla californicd° 64 2.7 1.5 Charadriiformes Vanellus chilensis 270 3.5 8.6 Columbiformes Columba araucana 300 2.7 7.2 Zenaida auriculata 137 9.8 12.1 Passeriformes Troglodytes aedon 10 0.9 0.1 Turdus falcklandii 90 14.3 11.5 Sicalis luteola 16 15.2 2.2 Sturnella loyca 96 3.5 3.0 Curaeus curaeus 95 0.9 0.8 Phry^lus patagonicus 38 0.9 0.3 Carduelis barbata 15 0.9 0.1 Unident, passeriformes 51 4.5 2.0 Reptiles 4.5 0.3 Liolaemus spp. 8 4.5 0.3 Insects 7.1 0.03 Coleoptera 0.5 3.5 0.01 Orthoptera 0.5 1.8 0.01 Unident, insects 0.5 1.8 0.01 Total prey items {N) 112 Total biomass (g) 12502 Total pellets (A/) 68 ® Likely juveniles. Introduced. 266 Short Communications VoL. 33, No. 3 plementary methods; microscopic analysis of feather structures such as nodes and barbules (Reyes 1992) and comparison of feather coloration patterns with voucher specimens deposited in the Zoology Department, Univer- sidad Austral de Chile at Valdivia and in the National Museum of Natural History at Santiago. Because we ob- served harriers eating birds entirely, we used the pres- ence of feathers of a given species in a pellet as repre- senting only one individual. Mammals were identified and quantified on the basis of skulls or dentary pairs, whichever gave the highest count following keys in Reise (1973). Reptiles were identified and quantified by the presence of scales and/ or teeth, and insects by head cap- sules, mandibles or elytra following keys in Donoso-Bar- ros (1966) and Pena (1986), respectively. We identified prey items to the finest possible taxonomic category. Biomass contribution was estimated following Marti (1987). Masses of birds and small mammals were ob- tained for birds using Jaksic et al. (1983), Morgado et al. (1987), Jimenez and Jaksic (1989) and Egli (1996) and for mammals using Greer (1968), Pearson (1983) and Martinez (1993). Masses of lagomorphs were obtained from juveniles found killed by raptors such as Cinereous Harriers and Short-eared Owls {Asia flammeus) . Masses of lizards and insects were obtained from unpublished data of the authors. We assumed that masses of unidentified prey were similar to the mean mass of the most closely related identified taxon. Although it was not possible to obtain simultaneous data on relative abundance of bird species, during the winter of 1997 we estimated bird abundance using eight parallel, fixed-band (100 m wide) line transects (Bibby et al 1993) placed 400 m apart in the hunting areas of harriers. To test whether harriers took birds selectively or opportunistically, we compared prey frequency distribu- tion in pellets with bird abundance using Spearman’s rank correlations as recommended by Jaksic (1979) for coarse comparisons between prey consumption and avail- ability. Results and Discussion Pellets averaged 28.2 ±1.9 (±SE) mm in length and 14 9 ± 0.9 mm in width {N = 50) and had a mean dry weight of 1.0 ± 0.6 g {N = 68). We identified 112 prey Items in these pellets including 16 species of birds, five species of mammals, one species of reptile and two or- ders of insects. Birds were the staple prey comprising 67% by frequency and 73% by biomass of the diet. Most birds eaten were Passeriformes (41.1%) with Gras.sland Yellow-Finches {Sicalis luteola) and Austral Thrushes {Tur- dus falcklandii) being the most frequently eaten (Table 1) Eared Doves {Zenaida auriculata) ranked third by fre- quency but, by biomass, Eared Doves and Austral Thrush- es were the most important prey. This agreed with an- ecdotal reports by Housse (1945), who indicated that Eared Doves were the preferred prey of Cinereous Har- riers in southcentral Chile. Mammals were the second most common prey with olivaceous field mice {Abrothrix ohvaceus) and long-tailed rice rats {Oligoryzomys longicau- datus) the most frequent. However, because of their larg- er size, lagomorphs accounted for almost all of the mam- malian biomass (22.4%). The incidence of reptiles and insects was negligible both by number and biomass. Spearman rank correlation coefficients were u = 0.72 {P < 0.01) when all bird prey species were considered, and r^ = 0.94 (P = 0.02) when passerines alone were considered. These trends indicated that Cinereous Har- riers hunted avian prey opportunistically rather than se- lectively (i.e., they took avian prey in proportion to their availability) . Overall, our results agree with those of Jimenez and Jaksic (1988), who reported that birds were the main prey of Cinereous Harriers during the breeding season in steppe marshes of Magallanes, southernmost Chile. Also, the diet of these harriers was similar to that of Long- winged Harriers {Circus buffoni) in Argentina (Bo et al. 1996), indicating that South American harriers are essen- tially bird predators. Resumen. — Determinamos la dieta del vari {Circus ciner- eus) en agroecosistemas de la Region de la Araucania, sur de Chile, analizando 68 egagropilas colectadas en aso- ciaciones de pastizal-humedal durante el invierno de 1996. Los varis consumieron un amplio espectro de pre- sas que incluyo aves, mamiferos, reptiles e insectos. Sin embargo, las aves fueron la base de la dieta (16 especies presa), alcanzando 67% en frecuenciay 73% en biomasa. Entre estas, las mas frecuentes fueron Sicalis luteola (15%), Turdus falcklandii (14%) y Zenaida auriculata (10%). En terminos de biomasa, Z. auriculata y T. falck- landii hicieron una mayor contribucion a la dieta (ca. 12% cada uno). Los mamiferos alcanzaron el 21% de la frecuencia y contribuyeron con el 26% de la biomasa. La contribucion de reptiles e insectos fue poco relevante. En general, nuestros resultados son coincidentes con otros autores, quienes mencionan a las aves como las principales presas de los varis en la Region de Magalla- nes. [Traduccion de Autores] Acknowledgments We thank the Zenhder Stappung family for permission to work in their lands and we are especially grateful to Olga Stappung for partially funding this study, and to Lilian Stappung for logistical support. In addition, we wish to thank Roberto Schlatter and Juan C. Torres-Mura who kindly permitted access to the collections in the Zo- ology Department of Universidad Austral and Chile’s Na- tional Museum of Natural History, respectively. We also are grateful to David R, Martinez who allowed us to use his facilities and to Jaime R. Ran for access to his espe- cialized literature. D.R. Martinez and J.R. Rau made use- ful suggestions and criticisms on earlier drafts and two anonymous referees made valuable comments on the manuscript. Literature Cited Bibby, C., N. Burges and D. Hill. 1993. Bird census tech- niques. Academic Press, London, U.K. September 1999 Short Communications 267 Bo, M.S., S.M. CiCCHiNO AND M.M. Martinez. 1996. Diet of Long-winged Harrier { Circus buffoni) in southeast- ern Buenos Aires province, Argentina, /. Raptor Res. 30:237-239. Di Castri, F. and E. Hajek. 1976. Bioclimatologia de Chile. Universidad Catolica de Chile, Santiago, Chile. Donoso-Barros, R. 1966. Reptiles de Chile. Ediciones Universidad de Chile, Santiago, Chile. Egli, G. 1996. Biomorfologia de algunas aves de Chile Central. Bol. Chil. Ornitol. 3:2-9. Greer, J.K. 1968. Mamiferos de la provincia de Malleco. Publicacion del Museo Dillman S. Bullock, El Vergel, Angol, Chile. Hellmayr, C.E. 1932. The birds of Chile. Field Mus. Nat. Hist. Zool. Ser. 19:1-472. Housse, R. 1945. Las aves de Chile en su clasificacion moderna. Ediciones Universidad de Chile, Santiago, Chile. iRiARTE, J.A., W.L. Franklin and W.E. Johnson. 1990. Diet of sympatric raptors in southern Chile./. Raptor Res. 24:41-46. Jaksic, F.M. 1979. Tecnicas estadisticas simples para eva- luar selectividad dietaria en Strigiformes. Medio Am- biente (Chile) 4:114—118. AND J.E. Jimenez. 1986. The conservation status of raptors in Chile. Birds of Prey Bull. 3:95-104. , J.L. Yanez AND J.R. Rau. 1983. Trophic relations of the southernmost population of Dusycion in Chile. J. Mammal. 64:693-697. Jimenez, J.E. and F.M. Jaksic. 1988. Ecology and behavior of southern South American Cinereous Harriers, Cir- cus cinereus. Rev. Chil. Hist. Nat. 61:199-208. AND . 1989. Biology of the Austral Pygmy- Owl. Wilson Bull. 10:377-389. Johnson, A.W. 1965. The birds of Chile and adjacent re- gions of Argentina, Bolivia and Peru. Vol. 1. Platt Es- tablecimientos Graficos, Buenos Aires, Argentina. Marti, C. 1987. Raptor food habits studies. Pages 67-79 in B.A. Pendleton, B.A. Millsap, K.W. Cline and D M Bird [Eds.], Raptor management techniques manual. Natl. Wildl, Fed., Sci. and Tech. Ser. 10, Washington DC U.S.A. Martinez, D.R. 1993. Food habits of the Rufous-legged Owl (Strix rufipes) in temperate rainforests of south- ern Chile./. Raptor Res. 27:214—216. Morgado, E., B. Gunther and U. Gonzalez. 1987. On the allometry of wings. Rev. Chil. Hist. Nat. 60:71-79 Pearson, O.P. 1983. Characteristics of a mammalian fau- na from forests in Patagonia, southern Argentina. / Mammal. 64:476—492. Pena, L.E. 1986. Introduccion a los insectos de Chile Editorial Universitaria, Santiago, Chile. Reise, D. 1973. Clave para la determinacion de los cra- neos de marsupiales y roedores chilenos. Guyana Zool. 27:1-20. Reyes, C.S. 1992. Clave para la identificacion de los 6r- denes de aves chilenas: microestructura de los nodos de las barbulas. Seminario para optar al titulo de Pro- fesor de Estado en Biologia y Ciencias Naturales, Universidad de los Lagos, Osorno, Chile. Saggese, M.D. and E.R. De Lucca. 1995. Reproduccion del gavilan ceniciento Circus cinereus en la Patagonia Argentina. Hornero 14:21-26. Received 7 October 1998; accepted 11 April 1999 J. Raptor Res. 33(3):268-269 © 1999 The Raptor Research Foundation, Inc. Striated Caracara {Phalcoboenus australis) Population at Staten and Ano Nuevo Islands Esteban Frere and Alejandro Travaini Centro de Investigaciones de Puerto Deseado, Universidad Nadonal de la Patagonia Austral, Alte. Brown y Colon s/n, (9050) Puerto Deseado, Santa Cruz, Argentina Anibal Pt\RERA Fundacion Vida Silvestre Argentina, Defensa 245 6° piso, (1065) Buenos Aires, Argentina Adrian Schiavini Centro Austral de Investigaciones Cientificas, CC 92, (9410) Ushuaia, Tierra delFuego, Argentina Key Words: Striated Caracara; Phalcoboenus australis; population sizer, Argentina. The Striated Caracara (Phalcoboenus australis) is cur- rently considered to be globally Near-threatened (Collar et al. 1994). Distributed on southern islands off southern South America (Narosky and Yzurieta 1987), its heavy persecution on the Islas Malvinas (Falkland Islands), where it was believed to attack weak or stranded sheep (del Hoyo et al. 1994), led to its local extinction and isolation on the smaller islands of the group. Its present status is currently unknown and surveys are needed, es- pecially in the islands where Striated Caracaras are sup- posedly locally numerous (del Hoyo et al. 1994). Here, we report on the first exhaustive Striated Caracara census conducted at Staten and Aho Nuevo islands during a sci- entific survey aimed at evaluating the islands’ faunal con- servation status (Parera et al. 1997), Both islands (63 000 ha) are located at the southern extreme of South America, 30 km away from Isla Grande de Tierra del Fuego which is shared by both Argentina and Chile across the Le Maire Straits (Fig. 1). The to- pography of Staten Island is rugged with a peak elevation of 823 m at Monte Bove and the coastline is highly irreg- ular with many fiords and bays (Kiihnemann 1976). The Ano Nuevo Islands are a group of islets with similar to- pography, north of Staten Island. The weather on both islands is cold with a mean temperature of 6-8 and — 2- 4°C during summer and winter, respectively. Oceanic hu- mid winds are responsible for an annual rainfall of 2000 mm (Kuhnemann 1976). The survey was conducted from 12 November-9 De- cember 1995. Counts were conducted by three experi- enced observers from a 12-m long sailboat (maximum 110 m off the coastline), a rubber boat and ground searching by foot when possible. Binoculars (10X50) and a 12-40X spotting scope were used for bird identifica- tion. The position of detected birds was obtained with a Global Positioning System (Garmin 50, GPS). Thirty-seven different Striated Caracaras were observed at 12 locations along 196 km of coastline survey (68% of total island coastline. Fig. 1) for a linear density of one caracara for every 5.3 km. Two Andean Condors (Vultur gryphus), 17 Turkey Vultures (Cathartes aura), one Grey Buzzard-Eagle (Geranoaetus melanoleucus) , six Crested Ca- racaras (Caracara plancus), and one Red-backed Hawk (Buteo polyosoma) were also observed during surveys. Stri- ated Caracaras were the most abundant raptors account- ing for 59% of total observations and were as common as Crested Caracaras in continental habitats (one individ- ual for every 4.4 km; Travaini et al. 1995). Crested Caracara and Turkey Vulture counts at Staten Island were similar to those obtained for the same species in Patagonia where linear densities were 40 and 14 km/ individual, respectively (Travaini et al. 1995). Relative abundances for the other three raptor species we observed were quite low when compared to figures obtained in con- tinental Patagonia where linear densities of Andean Con- dors, Grey Buzzard-Eagles, and Red-backed Hawks ranged between 40-60 km/individual (Donazar et al. 1993). Striated Caracaras were more abundant in our study area than in the Peninsula Mitre in Tierra del Fuego (Clark 1984) . All the Striated Caracaras we observed were at or in close proximity to colonies of marine mammals and seabirds. We observed 68 seabird and marine mam- mal colonies along 196 km of coastline at Staten and Ano Nuevo Islands compared to only 12 colonies along 170 km of coastline at Peninsula Mitre, Tierra del Fuego (Pa- rera et al. 1997, Schiavini unpubl. data). Differential availability of food sources could explain the higher abundance of Striated Caracaras. Staten and Ano Nuevo Islands should be considered key areas to protect this geographically isolated raptorial species. Resumen. — Aunque el estado de conservacion del Caran- cho austral (Phalcoboenus australis) se considera incierto, la fuerte persecucion de sus poblaciones por el hombre hace suponer que podrfa estar amenazada. Aqui describimos los resultados de un censo realizado para esta especie en Isla de los Estados e Islas de Ano Nuevo (54°45'S, 64°98'W). Sobre un recorrido de 196 km, el 68% de la linea de costa total de las islas, se contaron 37 caranchos australes (un individuo cada 5.3 km). El carancho austral resulto ser la rapaz mas abundante del area estudiada, alcanzando sus densidades valores similares al de poblaciones bien conser- vadas del Carancho comun ( Caracara plancus) en areas con- tinentales. La abundancia del Carancho austral registrada en nuestro estudio es de las mas altas descritas para esta especie. La cantidad de colonias de aves y mamiferos ma- rinos presentes en las islas y la fuerte asociacion de la es- 268 September 1999 Short Communications 269 Figure 1, The location of Staten and Aho Nuevo Islands. The dotted line shows the coastline that was surveyed. pecie con estas colonias, parecen ser las causas del buen estado de conservacion de la poblacion de Carancho aus- tral. Isla del los estados e Islas de Aho Nuevo deben ser consideradas como ^eas prioritarias para la conservacion de esta especie. [Traduccion de Autores] Acknowledgments We acknowledge Fundacion Patagonia Natural and the Wildlife Conservation Society and by National Geograph- ic Society (Grant 5548/95) for financial support, and thank Julio Brunet, Sergio Balbino and Alejandro Chiz- zini for their help in the field work. Javier Bustamante, Miguel Delibes, David Martinez, and Fabian Jaksic pro- vided helpful comments on the manuscript. Literature Cited Clark, R.R. 1984. Notas sobre aves de Peninsula Mitre, Isla Grande de Tierra del Fuego, Argentina. Hornero 12:212-218. Collar, N.J., M.J. Crosby and A.J. Statterseield. 1994. Birds to watch 2. The world list of threatened birds. Birdlife International, Cambridge, U.K. Donazar, J.A., O. Ceballos, A. Travaini and F. Hiraldo 1993. Roadside raptor surveys in the Argentinean Pa- tagonia. /. Raptor Res. 27:106-110. DEL Hoyo, J., a. Elliot and J. Sargatal. 1994. Hand- book of the birds of the world. Vol. 2. Lynx Editions, Barcelona, Spain. Kuhnemann, O. 1976. Observaciones ecologicas sobre la vegetacion marina y terrestre de la Isla de los Estado (Tierra del Euego, Argentina). Ecoittr 3:121-248. Narosky, T. and D. Yzurieta. 1987. Guia para la identi- ficacion de las aves de Argentina y Uruguay. Vazquez Mazzini Editores, Buenos Aires, Argentina. Parera, a., a. SCHIAVINI and E. Frere. 1997. Relevamien- to ecologico de la Isla de los Estados. Tech. Bull. No. 38, Fundacion Vida Silvestre Argentina, Buenos Aires, Argentina. Travaini, A, A. Rodriguez, O. Ceballos, J.A. Donazar and F. Hiraldo. 1995. Roadside raptor surveys in cen- tral Argentina, 1995. Hornero 14:64—66. Received 29 August 1998; accepted 13 March 1999 Letters /. Raptor Res. 33(3) :270 © 1999 The Raptor Research Foundation, Inc. Foraging of Yellow-headed Caracaras in the Fur of a Three-toed Sloth Both the Yellow-headed Caracara {Milvago chimachima) and the brown-throated three-toed sloth (Bradypus variega- tus) inhabit low elevation areas from southern Central America to northern Argentina (Emmons 1990, Neotropical rainforest mammals: a held guide. Univ. Chicago Press, Chicago, IL U.S.A.; Sick 1993, Birds in Brazil, a natural history. Princeton Univ. Press, Princeton, NJ U.S.A.). The sloth is generally limited to forested environments (Emmons 1990), while the caracara prefers more open habitat (Haverschmidt 1962, Condor 64:154-158; Sick 1993). Under natural circumstances, most interactions between these two species should be limited to forest edges. In metropolitan settings, however, species may be forced together in parks encircled by urban surroundings. In these situations, it may be easier to observe behaviors between species that remain undetected in more remote areas. Here, we describe the foraging on a brown-throated three-toed sloth by Yellow-headed Caracaras. We observed the animals in the botanical garden of the Ciudad Universitaria in downtown Caracas, Venezuela, at approximately 1600 H on 16 November 1996. The sloth was 15-20 m above the ground in an isolated Cecropia tree, and was attended by two juvenile caracaras. The caracaras appeared to be picking from the fur of the sloth; our observational distance precluded us from seeing anything in the birds’ beaks. The sloth showed no sign of defen- siveness or aggression toward the caracaras even when they were foraging on its head and neck. It assumed a relaxed posture, reclining on a branch with its front legs extended behind its head. The behavior continued for 5-10 min, until we moved closer. At this point, the caracaras stopped their grooming behavior but remained within a few meters of the sloth. The caracaras appeared to be members of a larger group; we observed an adult and another juvenile in nearby trees. Although previously undescribed, the nature of this interaction is not surprising. Yellow-headed Caracaras are known to forage by picking ticks and botflies from domestic livestock. When no source of appropriate ectoparasites IS available, caracaras are opportunistic and may consume a wide variety of items including insects, fruit and carrion (Haverschmidt 1962). Thus, their opportunistic foraging in the fur of other slow-moving mammals may not be entirely unexpected. Sloths may represent a particularly attractive foraging substrate, as their fur contains a rich fauna of invertebrates, especially Lepidoptera (Waage and Montgomery 1976, Science 193:157-158). We thank Jack Clinton-Eitniear, Joan Morrison, and Mel Sunquist for help in finding appropriate references, as well as Jack Clinton-Eitniear, David Houston, Joan Morrison, Mike Wallace, David Whitacre and an anonymous re- viewer for comments on the manuscript. — Alan H. Krakauer, Museum of Vertebrate Zoology, 3101 Valley Life Sciences Building, University of California, Berkeley, CA 94720-3160 U.S A., and Thomas H. Krakauer, North Carol in a Museum of Life and Science, 433 Murray Avenue, P.O. Box 15190, Durham, NC 27704 U.SA. J Raptor Res. 33(3) :270-27l © 1999 The Raptor Research Foundation, Inc. Common Black-Hawk Nesting in West-central Texas Common Black-Hawks {Buteogallus anthracinus) are obligate riparian nesters of the southwestern U.S. Although not federally listed, it is listed as Endangered in New Mexico and Threatened in Texas (Schnell 1994, Common Black- Hawk, Birds of North America, No. 122, A. Poole and F. Gill [Eds.], The Birds of North America, Inc., Philadelphia, PA U.S.A.). Once a regular nester along the lower Rio Grande Valley of Texas, all local breeders were extirpated by 1940, apparently the result of extensive habitat loss. In 1970, a small breeding population (about 10 pairs) was discovered along Lympia Creek of the Davis Mountains in Jeff Davis County, Texas (Oberholser 1974, The bird life of Texas. Vol. 1., Univ. Texas Press, Austin, TX U.S.A.; Schnell 1994). To date, this remains the only known regular 270 September 1999 Letters 271 breeding population in Texas. Extralimital breeding attempts have been reported in Val Verde (Lasley and Sexton 1988, Am. Birds 42:456-462), Lubbock (Texas Ornithological Society 1995, Checklist of Texas birds, Capital Printing, Inc., Austin, TX U.S.A.), and Brewster (Lasley and Lockwood, Texas Bird Records Committee, Texas Ornithological Society, pers. comm.) counties and the lower Rio Grande Valley (Webster 1976, Am. Birds 23:975-978). Only in the Davis Mountains has successful breeding been documented for consecutive years. Nesting attempts by a Common Black-Hawk pair in Big Bend National Park, Brewster County, Texas, were unsuccessful in 1997 and 1998 (M. Lock- wood pers. comm.). We observed the successful nesting of a pair of Common Black-Hawks in Tom Green County, Texas from 1996- 98, at a location 345 km northeast of the nearest breeding population at Lympia Creek. A single pair of Common Black-Hawks was initially found during June 1996 nesting in a narrow (<300 m wide) band of riparian woodland along the spring-fed Cole Creek, a tributary of the South Concho River (2.5 km south, 0.62 km west Christoval, Texas). The 67 ha riparian woodland was dominated by pecan trees {Carya illinoensis\ 49%), bur oak {Quercus macro- carpa; 14%) and plateau liveoak (<2- virginiana-, 7%). Seven other tree species were present, but in small numbers Mean tree density and mean canopy height, as determined by a point-quarter technique (Cottam and Curtis 1956, Ecology 37:451-460), was 383 ha ^ and 21.4 m, respectively. Adjacent vegetation consisted of a patchwork of mixed mesquite {Prosopis glandulosa) brushlands and juniper (Juniperus spp.)4iveoak savanna. When they were found, the adults were carrying food (an unidentified snake) back to a nest containing one nestling. The nest was in the forked primary limb of a pecan tree of greater than local average size. The nest tree was 26 m tall, had a dbh of 107.7 cm and a maximum canopy width of 14.6 m. It was positioned 78 m from the edge of the riparian woodland and 125 m from Cole Creek. The closest tree of equal or greater dbh was 103 m away. Before nest measurements could be taken, the limb holding the nest fell during a storm following fledging. In 1997, a nest was constructed in the fork of a primary limb of the same pecan tree at a height of 18 m and 4.5 m from canopy edge. The limb had an estimated diameter of 26 cm at the base of the fork; the three branches arising from the fork had a diameter of about 14 cm This same nest was subsequently reused in 1998. Irregular visits were made to the site throughout the summer of 1996 and again from April through August of 1997 and 1998. During 1997 and 1998, a pair was observed nesting in the same tree as 1996. We observed the successful fledging of one individual each year. Attempts to band adults were unsuccessful, so we do not know if the nesting pair was the same each year. Although the nest was close to water in Cole Creek (125 m away), all foraging we observed was from the South Concho River at a man-made rock dam (1 km away). We were unable to identify food items carried from the South Concho River, but potential prey observed at the rock dam included crayfish {Procambarus clarki) , largemouth bass {Micropterus salmoides), gizzard shad {Darosoma cepedianum) , channel {Ictalurus punctatus) , blue (I. furcatus) andflathead {Pylodictis olivaris) catfish, green {Lepomis cyanellus) and long-eared {L. megalotis) sunfish, bluegill (L. macrochirus) , Rio Grande {Rana berlahdieri) and plains {R. blairi) leopard frogs, cricket frog {Acris crepitans), checkered garter snake {Thamnophis marcianus), ribbon snake {T. proximus), plain-bellied {Nerodia erythrogaster) and diamond-backed {N. rhom- bifera) watersnakes, and cottonmouth {Agkistrodon piscivorus) . While local tree species composition can be considerably different, structural characteristics of this location were similar to those reported at other nesting sites in Arizona, New Mexico and Texas (Schnell 1994). It is the isolation and consistent breeding success at this location that makes these records important. To our knowledge, this is the first known incident of successful breeding of Common Black-Hawks outside of the Davis Mountains in Texas since 1939. The chance discovery of this nesting site brings up the issue of potential range expansion by a species consid- ered Threatened in Texas. While similar pockets of riparian woodlands are widespread in west-central Texas, poten- tially providing suitable nesting habitat, many of these sites are on private land holdings making extensive surveys for more nests difficult. We thank the Boulware family for generously allowing us continued access to the nesting site and the summer 1998 Ecology class at Angelo State University for collecting local vegetation data. We also thank J.H. Schnell, R. Glinski and R Bloom for helpful comments on earlier versions of this manuscript. — ^Terry C. Maxwell, Department of Biology, Angelo State University, San Angelo, Texas 76909 U.S.A. and Michael S. Husak, Department of Biological Sciences, Mississippi State University, Mississippi State, Mississippi 39762 U.SA. BOOK REVIEW Edited by Jeffrey S. Marks J Raptor Res. 33(3) :272 © 1999 The Raptor Research Foundation, Inc. Holarctic Birds of Prey. Edited by B.-U. Mey- burg, R.D. Chancellor, and JJ. Ferrero. 1998. World Working Group on Birds of Prey and Owls, Berlin, Germany. 680 pp., numerous figures and tables. ISBN 84-605-7398-2. Paper, $35. — In con- junction with ADENEX (Asociacion para la Defen- sa de la Naturaleza y los Recursos de Extremadu- ra), the World Working Group on Birds of Prey and Owls (WWGBP) held a conference in Badajoz, Spain, from 17-22 April 1995. Of the 100 or so oral and poster presentations, 59 are included in this proceedings (16 in Spanish, 43 in English), which is organized into eight parts: “Molecular System- atics of Holarctic Birds of Prey and Owls” (3 pa- pers), “Breeding Biology of Raptors” (7), “Biology and Conservation of Diurnal Raptors” (26), “Sup- porting Raptor Populations in Danger of Extinc- tion via Captive Breeding” (4), “Biology and Con- servation of Holarctic Owls” (5), “Raptors: Contamination and Development” (4), “Socio- economic Aspects of Raptor Conservation” (3), and “Migration of Birds of Prey” (7). Rather than attempt to mention all of the papers, I will note a few (among the many) that I found to be especially interesting. Papers by Michael Wink and associates present data on molecular systematics of hawks and owls based on the cytochrome-^> gene. The latter paper (with P. Heidrich) contains two notable tidbits: (1) the genetic distance is 5-7% between Little Owls {Athene noctua) from Israel versus Europe, suggest- ing that two species are involved; and (2) the “white-faced scops-owls,” formerly regarded as two taxa of Otus, actually are closer to Asia and belong in the genus Ptilopsis. Among the handful of contributions from North America, Charles Henny presents a review of chemical contamination that focuses on falconi- forms, James Enderson et al. provide an overview of the captive-breeding program for Peregrine Fal- cons {Falco peregrinus) in the U.S.A. and Canada, and Keith Bildstein challenges raptor biologists who study migration to form partnerships with mainstream ecologists and conservation biologists to broaden efforts to understand raptor migration on a global scale. Another thought-provoking pa- per, by Clayton White and Lloyd Kiff, discusses how efforts to list the Northern Goshawk {Accipiter gen- tilis) and to avoid delisting the Peregrine Falcon potentially weaken the Endangered Species Act by clouding the definition of “endangered” and by lowering the credibility of biologists who study these species. This paper should be read by every- one interested in the conservation of rare and threatened species. Jeff Watson reports that Golden Eagles {Aquila chrysaetos) in western Scodand take a higher diver- sity of prey species than those in eastern Scotland and have concomitandy lower reproductive suc- cess. The difference in diet results from reduced numbers of medium-sized herbivores in western Scodand. Watson suggests that Golden Eagles do best when a few species of prey of favored size are “particularly abundant” and that they fair poorly when forced to widen their diet to include species outside of the preferred range of prey size. Ian Newton reviews migration patterns of falconiforms in the western Palearctic, drawing a correlation be- tween migration distance and diet. Raptors that specialize on birds and mammals tend to migrate short distances and winter in the Palearctic, where- as those that eat mainly cold-blooded prey (i.e., reptiles and invertebrates) are long-distance mi- grants that winter in sub-Saharan Africa. Holarctic Birds of Prey continues the line of worth- while publications on raptor biology edited by Chancellor, Meyburg, and their associates and pro- duced by the WWGBP. As such, it would make a valuable addition to one’s library. — Jeff Marks, Montana Cooperative Wildlife Research Unit, Uni- versity of Montana, Missoula, MT 59812 U.S.A. 272 Erratum J. Raptor Res. 33(3):273 © 1999 The Raptor Research Foundation, Inc. Diet Composition and Reproductive Success of Mexican Spotted Owls The following correction should be made to the article “Diet Composition and Reproductive Success of Mexican Spotted Owls,” Journal of Raptor Research 33(2): 143-1 48. Page 143, abstract, third sentence of abstract should read “This was contrary to previously observed diet patterns m Northern (S. o. caurina) and California Spotted Owls (5. o. ocddentalis) . Mammals comprised 88.2% of the dietary biomass in Arizona and 94.0% in New Mexico.” 273 BUTEO BOOKS The following Birds of North America Species Accounts are available through Buteo Books, 3130 Laurel Road, Shipman, VA 22971. TOLL-FREE ORDERING: 1-800-722-2460; FAX; (804) 263-4842. E-mail: alien® buteobooks.com Barn Owl (1). Carl D. Marti. 1992. 16 pp. Boreal Owl (63). G.D. Hayward and P.H. Hayward. 1993. 20 pp. Broad-winged Hawk. (218). L.J. Goodrich, S.C. Crocoll and S.E. Senner. 1996. 28 pp. Burrowing Owl (61). E.A. Haug, B.A. Millsap and M.S. Martell. 1993. 20 pp. Common Black-Hawk (122). Jay H. Schnell. 1994. 20 pp. Cooper’s Hawk (75). R.N. Rosenfield and J. Bielefeldt. 1993. 24 pp. Crested Caracara (249). Joan L. Morrison. 1996. 28 pp. Eastern Screech-owl (165). Frederick R. Gehlbach. 1995. 24 pp. Ferruginous Hawk (172). MarcJ. Bechard and Josef K. Schmutz. 1995. 20 pp. Flammulated Owl (93). D. Archibald McCallum. 1994. 24 pp. Great Gray Owl (41). Evelyn L. Bull and James R. Duncan. 1993. 16 pp. Great Horned Owl (372). C. Stuart Houston, Dwight G. Smith, and Christoph Rohner. 1998. 28 pp. Gyrfalcon (114). Nancy J. Clum and Tom J. Cade. 1994. 28 pp. Harris’ Hawk (146). James C. Bednarz. 1995. 24 pp. Long-eared Owl (133). J.S. Marks, D.L. Evans and D.W. Holt. 1994. 24 pp. Merlin (44). N.S. Sodhi, L. Oliphant, P. James and I. Warkentin. 1993. 20 pp. Mississippi Kite (402). James W. Parker. 1999. 28 pp. Northern Saw-whet Owl (42). Richard J. Cannings. 1993. 20 pp. Northern Goshawk (298). John R. Squires and Richard T. Reynolds. 1997. 32 pp. Northern Harrier (210). R. Bruce MacWhirter and Keith L. Bildstein. 1996. 32 pp. Northern Hawk Owl (356). James R. Duncan and Patricia A. Duncan. 1998. 28 pp. Red-shouldered Hawk (107). Scott T. Crocoll. 1994. 20 pp. Red-tailed Hawk (52). C.R. Preston and R.D. Beane. 1993. 24 pp. Short-eared Owl (62). D.W. Holt and S.M. Leasure. 1993. 24 pp. Snail Kite (l7l). P.W. Sykes, Jr., J. A. Rodgers, Jr, and R.E. Bennetts. 1995. 32 pp. Snowy Owl (10). David E. Parmelee. 1992. 20 pp. Spotted Owl (179). R.J. Gutierrez, A.B. Franklin and W.S. Lahaye. 1995. 28 pp. Swainson’s Hawk (265). A. Sidney England, MarcJ. Bechard and C. Stuart Houston. 1997. 28 pp. Swallow-tailed Kite (138). Kenneth D. Meyer. 1995. 24 pp. Turkey Vulture (339). David A. Kirk and Michael J. Mossman. 1998. 32 pp. White-tailed Hawk (30). C. Craig Farquhar. 1992. 20 pp. White-tailed Kite (178). Jeffrey R. Dunk. 1995. 16 pp. Buteo Books stocks all published species accounts, not only those covering raptors. The current list in taxo- nomic order may be viewed at: http://www.buteobooks.com Buteo Books stocks the Handbook of the Birds of the World. The first five volumes of this projected 12-volume work have been published including; Volume 2; New World Vultures to Guineafowl (1994) covering the diurnal raptors and Volume 5: Barn Owls to Hummingbirds (1999) covering owls. These volumes are priced at $185 each plus shipping and handling. Usually available from Buteo Books, the classic reference on diurnal birds of prey: Brown, Leslie and Dean Amadon. Eagles, Hawks and Ealcons of the World. Country Life Books, 1968. Two volumes. First English edition in brown cloth. Fine in slipcase. $300.00 and other editions at lesser prices. video Systems AUTO-COLOR SENTINEL SYSTEM Installed in an artificial eagle's nest, this surveillance system combines near-infrared and color cameras for day and nighttime recording. PEEPER VIDEO PROBE Ideal for studying burrowing owls. TREETOP PEEPER II Access a variety of nests with this 50-foot elevating video system. Christensen Designs (209) 239-8090 PST www.Peeperpeople.com • email: Arm@Peeperpeople.com WILDLIFE RESEARCH EQUIPMENT Vocalizers • Video triggers • Solar /battery saving devices IR spotlights • Digital equipment • Custom designs Equipment Grants for graduate students ARDEOLA The Official Journal of the Spanish Ornithological Society ARDEOLA VOL. 40(1) JUNIO 19^ SOCIEDAD ESPANOLA DE ORNITOLOGIA An Invitation to Contribute! ARDEOlX has recently expanded its scope to provide a renewed forum for publication of ornithological studies, both empirical and experimental, carried out both in Spain and elsewhere. Papers dealing with ornithology in the areas of Mediterranean climate and in the Neotropics, as well as with the conservation of birds and their habitats, are especially acknowledged. The Editorial staff encourages the use of English to prospective authors, and will promote rapid publication of peer-reviewed manuscripst. Please address manuscript submissions, editorial queries and comments to: Dr. Mario Diaz Departamento de Ciencias Medioambientales Facultad de Ciencias del Medio Ambiente. Universidad de Castilla-La Mancha E-45071 Toledo (Spain) e-mail mariod@amb-to.uclm.es Subscription rates to ARDEOLA: Annual subscription for SEO members: 2.000 pesetas + 5.000 pesetas of membership fee (US $15 + 45) Annual subscription for non-members: 10.000 pesetas (US $95) For more information about SEO memebership, subcribing to ARDEOLA or receiving other SEO publications please contact the Society 's headquarters: www.seo.org SEO/BirdLife, Ctra. de Humera, 63-1, 28224 Pozuelo de Alarcon, Madrid (Spain). Fax: 34 1 351 1386. e-mail bmdina(5> seo.org THE RAPTOR RESEARCH FOUNDATION, INC. (Founded 1966) PRESIDENT: Michael N. Kochert VICE-PRESIDENT: Keith L. Bildstein OFFICERS SECRETARY: Patricia A. H alt. TREASURER: Jim Fitzpatrick BOARD NORTH AMERICAN DIRECTOR #1: Brian A. Millsap NORTH AMERICAN DIRECTOR #2: Petra Bohall Wood North American director #3: Robert Lehman INTERNATIONAL DIRECTOR #1: Massimo Pandolfi INTERNATIONAL DIRECTOR #2: Reuven Yosef OF DIRECTORS INTERNATIONAL DIRECTOR #3; Beatriz Arrovo DIRECTOR AT LARGE #1: Jemima ParryJones DIRECTOR AT LARGE #2: John A. Smallwood DIRECTOR AT LARGE #3: James C. Bednarz DIRECTOR AT LARGE #4: Miguel Ferrer DIRECTOR AT LARGE #5: Lloyd Kief DIRECTOR AT LARGE #6: Robert Kenward EDITORIAL STAFF EDITOR: Marc J. Bechard, Department of Biology, Boise State University, Boise, ID 83725 U.S.A. ASSOCIATE EDITORS Allen M. Fish Fabian Jaksic Gary R. Bortolotti Daniel E. Varland Charles J. Henny Cole Crocker-Bedford BOOK REVIEW EDITOR: Jeffrey S. Marks, Montana Cooperative Research Unit, University of Montana, Missoula, MT 59812 U.S.A. SPECIAL PUBLICATIONS EDITOR: Daniel E. Varland, Rayonier, 3033 Ingram Street, Hoquiam, WA 98550 SPANISH EDITOR: Cesar Marquez Reyes, Instituto Humboldt, Colombia, AA. 094766, Bogota 8, Colombia The Journal of Raptor Research is distributed quarterly to all current members. Original manuscripts dealing with the biology and conservation of diurnal and nocturnal birds of prey are welcomed from throughout the world, but must be written in English. Submissions can be in the form of research articles, letters to the editor, thesis abstracts and book reviews. Contributors should submit a typewritten original and three copies to the Editor. All submissions must be typewritten and double-spaced on one side of 216 X 278 mm (SVi X II in.) or standard international, white, bond paper, with 25 mm (1 in.) margins. The cover page should contain a title, the author’s full name(s) and address(es). Name and address should be centered on the cover page. If the current address is different, indicate this via a footnote. A short version of the title, not exceeding 35 characters, should be provided for a running head. An abstract of about 250 words should accompany all research articles on a separate page. Tables, one to a page, should be double-spaced throughout and be assigned consecutive Arabic numer- als. Collect all figure legends on a separate page. Each illustration should be centered on a single page and be no smaller than final size and no larger than twice final size. The name of the author(s) and figure number, assigned consecutively using Arabic numerals, should be pencilled on the back of each figure. Names for birds should follow the A.O.U. Checklist of North American Birds (7th ed., 1998) or another authoritative source for other regions. Subspecific identification should be cited only when pertinent to the material presented. Metric units should be used for all measurements. Use the 24-hour clock (e.g., 0830 H and 2030 H) and “continental” dating (e.g., 1 January 1990). Refer to a recent issue of the journal for details in format. Explicit instructions and publication policy are outlined in “Information for contributors,”/. Raptor Res., Vol. 32(4), and are available from the editor. 1999 ANNUAL MEETING The Raptor Research Foundation, Inc. 1999 annual meeting will be held on 3—7 November in La Paz, Mexico. For information about the meeting contact Ricardo Rodriguez Estrella, Centro de Investigaciones Biologicas del Noreste, Division de Biologica Terrestrie, KM 1 Carretera, San Juan de LaCosta, La Paz 23000, Mexico. Telephone 112-536-33, FAX 112-553-43, E-mail estrella@cibnor. mex. Persons interested in predatory birds are invited to join The Raptor Research Foundation, Inc. Send requests for information concerning membership, subscriptions, special publications, or change of address to OSNA, P.O. Box 1897, Lawrence, KS 66044-8897, U.S.A. The Journal of Raptor Research (ISSN 0892-1016) is published quarterly and available to individuals for $33.00 per year and to libraries and institutions for $50.00 per year from The Raptor Research Foundation, Inc., 14377 117th Street South, Hastings, Minnesota 55033, U.S.A. (Add $3 for destinations outside of the continental United States.) Periodicals postage paid at Hastings, Minnesota, and additional mailing offices. POSTMASTER: Send address changes to The Journal of Raptor Research, OSNA, P.O. Box 1897, Lawrence, KS 66044-8897, U.S.A. Printed by Allen Press, Inc., I^wrence, Kansas, U.S.A. Copyright 1999 by The Raptor Research Foundation, Inc. Printed in U.S.A. 0 This paper meets the requirements of ANSI/NISO Z39.48-1992 (Permanence of Paper). Raptor Research Foundation, Inc., Awards Recognition for Significant Contributions^ The Dean Amadon Award recognizes an individual who has made significant contributions in the field of systematics or distribution of raptors. Contact: Dr. Clayton White, 161 WIDB, Department of Zoology, Brigham Young University, Provo, UT 84602 U.SA,. Deadline August 15. The Tom Cade Award recognizes an individual who has made significant advances in the area of captive propagation and reintroduction of raptors. Contact: Dr. Brian Walton, Predatory Bird Research Group, Lower Quarry, University of California, Santa Cruz, CA 95064 U.SA. Deadline: August 15. The Fran and Frederick Hamerstrom Award recognizes an individual who has contributed signiBcantly to the understanding of raptor ecology and natural history. Contact: Dr. David E. Andersen, Department of Fisheries and Wildlife, 200 Hodson Hall, 1980 Folwell Avenue, University of Minnesota, St. Paul, MN 55108 U.S.A. Deadline: August 15. Recognition and Travel Assistance The James R. Kophn Travel Award is given to a student who is the senior author of the paper to be presented at the meeting for which travel funds are requested. Contact: Dr. Petra Wood, West Virginia Cooperative Fish and Wildlife Research Unit, P.O. Box 6125, Percival Hall, Room 333, Morgantown, WV 26506-6125 U.SA. Deadline: established for conference paper abstracts. The WUham C. Andersen Memorial Award is given to the student who presents the best paper at the annual Raptor Research Foundation Meeting. Contact: Ms. Laurie Goodrich, Hawk Mountain Sanctuary, Rural Route 2, Box 191, Kempton, PA 19529-9449 U.SA. Deadline: Deadline established for meeting paper abstracts. Grants^ The Stephen R. Tully Memorial Grant for $500 is given to support research, management and conservation of raptors, especially to students and amateurs with limited access to alternative funding. Contact: Dr. Kimberly Titus, Alaska Division of Wildlife Conservation, P.O. Box 20, Douglas, AK 99824 U.SA Dead- line: September 10. The Leslie Brown Memorial Grant for $500-$l,000 is given to support research and/or the dissemination of information on raptors, especially to individuals carrying out work in Africa. Contact: Dr. Jeffrey L. Lincer, 1220 Rosecrans St. #315, San Diego, CA 92106 U.SA. Deadline: September 15. ’ Nominations should include: (1) the name, title and address of both nominee and nominator, (2) the names of three persons qualihed to evaluate the nominee’s scientific contribution, (3) a brief (one page) summary of the scientific contribution of the nominee. ^ Send 5 copies of a proposal (^5 pages) describing the applicant’s background, study goals and methods, anticipated budget, and other funding.